United States

Department of 

Agriculture

Forest Service

Northern 

Research Station

General Technical

Report NRS-80

Assessment of Nitrogen 

Deposition Effects and 

Empirical Critical Loads 

of Nitrogen for Ecoregions 

of the United States

L.H. Pardo, M.J. Robin-Abbott, C.T. Driscoll, editors
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Abstract

Human activity in the last century has led to a substantial increase in nitrogen (N) emissions 

and deposition. This N deposition has reached a level that has caused or is likely to cause 

alterations to the structure and function of many ecosystems across the United States. One 

approach for quantifying the level of pollution that would be harmful to ecosystems is the 

critical loads approach. The critical load is defi ned as the level of a pollutant below which no 

detrimental ecological effect occurs over the long term according to present knowledge.

The objective of this project was to synthesize current research relating atmospheric N 

deposition to effects on terrestrial and aquatic ecosystems in the United States and to 

identify empirical critical loads for atmospheric N deposition. The receptors that we evaluated 

included freshwater diatoms, mycorrhizal fungi and other soil microbes, lichens, herbaceous 

plants, shrubs, and trees. The main responses reported fell into two categories: (1) 

biogeochemical, and (2) individual species, population, and community responses.

The range of critical loads for nutrient N reported for U.S. ecoregions, inland surface waters, 

and freshwater wetlands is 1 to 39 kg N ha-1 y-1. This broad range spans the range of N 

deposition observed over most of the country. The empirical critical loads for N tend to 

increase in the following sequence for different life forms: diatoms, lichens and bryophytes, 

mycorrhizal fungi, herbaceous plants and shrubs, trees.

The critical loads approach is an ecosystem assessment tool with great potential to simplify 

complex scientifi c information and effectively communicate with the policy community and 

the public. This synthesis represents the fi rst comprehensive assessment of empirical critical 

loads of N for ecoregions across the United States.

Manuscript received for publication 25 February 2010

Cover Photos

Front, left to right, top: elevated nitrogen inputs to a prairie grassland in Minnesota (control) 

resulted in a decrease in species richness and an increase in invasive grasses (N addition). 

Photos by David Tilman, University of Minnesota, used with permission.

Front, left to right, bottom: elevated nitrogen inputs to a high elevation spruce fi r forest 

(control) in Vermont resulted in decreased growth and increased mortality (high treatment). 

Photos by Linda Pardo, U.S. Forest Service.

Back top: The endangered checkerspot butterfl y (Euphydryas editha bayensis). Photo by 

Stuart Weiss, used with permission.

Back bottom: The threatened purple pitcher plant (Sarracenia purpurea L.). Photo by Lingli 

Liu, used with permission.
Assessment of Nitrogen Deposition Effects 

and Empirical Critical Loads of Nitrogen 

for Ecoregions of the United States

Editors

L.H. Pardo

M.J. Robin-Abbott

C.T. Driscoll

LINDA H. PARDO, environmental engineer, U.S. Forest Service, Northern 

Research Station, 705 Spear St., South Burlington, VT 05403; lpardo@fs.fed.us

MOLLY J. ROBIN-ABBOTT, research assistant, Syracuse University, located 

at U.S. Forest Service, Northern Research Station, 705 S. Spear St., South 

Burlington, VT 05403

CHARLES T. DRISCOLL, professor, Syracuse University, Department of Civil 

and Environmental Engineering, 151 Link Hall, Syracuse, NY 13244.
Acknowledgments

This project was funded, in part, by the U.S. Environmental Protection Agency’s Clean Air 

Markets Division. Initial work on this project was funded by the U.S. Forest Service Air Resource 

Management program. Funding was provided by the UCR Center for Conservation Biology and 

NSF grant DEB 04-21530. We appreciate the reviews of Suraj Ahuja, Allison Aldous, Mary Arthur, 

Jayne Belnap, John Blair, Tamara Blett, Matthew Brooks, Andrew Burton, William Currie, Daniela 

Cusack, Robert Edmonds, Heather Erickson, Ilka Feller, Benjamín Sánchez Gimeno, Rick Graw, 

Rick Haeuber, Cindy Huber, Randy Kolka, Steven Lindberg, Kathleen Lohse, Michelle Mack, Bill 

McDowell, Steve McNulty, Anne Mebane, Tom Meixner, John Melack, Kristi Morris, Bengt Nihlgård, 

Annika Nordin, Rock Ouimet, Michael Pace, Gareth Phoenix, Cindy Prescott, James Sickman, Jack 

Triepke, David Wedin, Trent Wickman, and Mark Williams. We would also like to thank Scott Collins, 

Robin Dennis, Amanda Elliot Lindsey, Ray Gomez, Robert Johnson, Duan Lei, and Bethany Zinni.
CONTRIBUTING AUTHORS

EDITH B. ALLEN, professor of plant ecology, Department of Botany and Plant Sciences and Center for 

Conservation Biology, University of California, Riverside, CA

JILL S. BARON, research ecologist, U.S. Geological Survey, Natural Resource Ecology Laboratory, 

Colorado State University, Fort Collins, CO

ROLAND BOBBINK, director/ senior consultant, B-WARE Research Centre B.V., Radboud Universiteit, 

Nijmegen, NL

WILLIAM D. BOWMAN, professor, Department of Ecology and Evolutionary Biology and Institute of 

Arctic and Alpine Research, University of Colorado, Boulder, CO

CHRISTOPHER M. CLARK, AAAS science and technology policy fellow, Offi ce of Research and 

Development, National Center for Environmental Assessment, Global Change Research Program, 

U.S. Environmental Protection Agency, Washington, D.C.

CHARLES T. DRISCOLL, professor, Syracuse University, Department of Civil and Environmental 

Engineering, Syracuse, NY

BRIDGET EMMETT, head of site, Centre for Ecology and Hydrology, Environment Centre Wales, 

Bangor, Gwynedd, UK

MARK E. FENN, research plant pathologist, U.S. Forest Service, Pacifi c Southwest Research Station, 

Forest Fire Laboratory, Riverside, CA

LINDA H. GEISER, ecologist, U.S. Forest Service, Pacifi c Northwest Region Air Program, Corvallis, OR

FRANK S. GILLIAM, professor, Department of Biological Sciences, Marshall University, Huntington, WV

CHRISTINE L. GOODALE, associate professor, Department of Ecology and Evolutionary Biology, 

Cornell University, Ithaca, NY

TARA GREAVER, ecologist, Environmental Media Assessment Group, National Center for 

Environmental Assessment, Offi ce of Research and Development, U.S. Environmental Protection 

Agency, Research Triangle Park, NC

SHARON J. HALL, assistant professor, Ecology, Evolution, and Environmental Science, School of Life 

Sciences, Arizona State University, Tempe, AZ

ERIK LILLESKOV, research ecologist, U.S. Forest Service, Northern Research Station, Forestry 

Sciences Laboratory, Houghton, MI

LINGLI LIU, ORISE post-doctoral fellow, Environmental Media Assessment Group, National Center 

for Environmental Assessment, Offi ce of Research and Development, U.S. Environmental Protection 

Agency, Research Triangle Park, NC

JASON A. LYNCH, ecologist, Clean Air Markets Division, U.S. Environmental Protection Agency, 

Washington, D.C.

KNUTE NADELHOFFER, professor, Department of Ecology and Evolutionary Biology, University of 

Michigan, Ann Arbor, MI

LINDA H. PARDO, environmental engineer, U.S. Forest Service, South Burlington, VT

STEVEN S. PERAKIS, research ecologist, U.S. Geological Survey, Forest and Rangeland Ecosystem 

Science Center, Corvallis, OR

MOLLY J. ROBIN-ABBOTT, research assistant, Syracuse University, located at U.S. Forest Service, 

South Burlington, VT

JOHN L. STODDARD, research life scientist, U.S. Environmental Protection Agency, Western Ecology 

Division, Corvallis, OR

KATHLEEN C. WEATHERS, senior scientist, ecosystem ecologist, Cary Institute of Ecosystem 

Studies, Millbrook, NY
CONTENTS

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

L.H. Pardo, C.T. Driscoll, C.L. Goodale

2 Ecosystem Classifi cation ............................................................................... 9

M.J. Robin-Abbott, L.H. Pardo

3 Deposition ...................................................................................................... 15

K.C. Weathers, J.A. Lynch

4 Methods .......................................................................................................... 25

L.H. Pardo, E.A. Lilleskov, L.H. Geiser, M.J. Robin-Abbott

5 Tundra ............................................................................................................ 37

K. Nadelhoffer, L.H. Geiser

6 Taiga ............................................................................................................... 49

L.H. Geiser, K. Nadelhoffer

7 Northern Forest ............................................................................................. 61

 L.H. Pardo, C.L. Goodale, E.A. Lilleskov, L.H. Geiser

8 Northwestern Forested Mountains .............................................................. 75

W.D. Bowman, J.S. Baron, L.H. Geiser, M.E. Fenn, E.A. Lilleskov

9 Marine West Coast Forest ............................................................................ 89

S.S. Perakis, L.H. Geiser, E.A. Lilleskov

10 Eastern Temperate Forest .......................................................................... 99

F.S. Gilliam, C.L. Goodale, L.H. Pardo, L.H. Geiser, E.A. Lilleskov

11 Great Plains.................................................................................................117

C.M. Clark

12 North American Deserts ........................................................................... 133

E.B. Allen, L.H. Geiser

13 Mediterranean California .......................................................................... 143

M.E. Fenn, E.B. Allen, L.H. Geiser

14 Southern Semi-Arid Highlands ................................................................ 171

M.E. Fenn, E.B. Allen

15 Temperate Sierra ....................................................................................... 175

M.E. Fenn, L.H. Geiser

16 Tropical and Subtropical Humid Forests ................................................ 181

S.J. Hall

17 Wetlands..................................................................................................... 193

T. Greaver, L. Liu, R. Bobbink
18 Inland Surface Water ................................................................................. 209

J.S. Baron, C.T. Driscoll, J.L. Stoddard

19 Synthesis.................................................................................................... 229

L.H. Pardo, L.H. Geiser, M.E. Fenn, C.T. Driscoll, C.L. Goodale, E.B. Allen, 

J.S. Baron, R. Bobbink, W.D. Bowman, C.M. Clark, B. Emmett, F.S. Gilliam, 

T. Greaver, S.J. Hall, E.A. Lilleskov, L. Liu, J.A. Lynch, K. Nadelhoffer, 

S.S. Perakis, M.J. Robin-Abbott, J.L. Stoddard, K.C. Weathers.

Appendix 1 Plant Species Common and Scientifi c Names ........................ 285

Appendix 2 Chemical Appendix .................................................................... 291

Chapter 1—Introduction GTR-NRS-80 1

1 INTRODUCTION

L.H. Pardo, C.T. Driscoll, C.L. Goodale

events can represent the bulk of N deposition, mean 

annual N deposition of 71 kg ha-1 y-1 was reported 

(Fenn et al. 2008). Nitrogen deposition has increased 

signifi cantly in the last decade in some regions of the 

United States and is projected to increase further in 

certain regions of the country (Lehmann et al. 2005, 

Nilles and Conley 2001). Estimates of N deposition are 

discussed in more detail in Chapter 3.

1.2.2 Eff ects of N deposition

Th e increases in atmospheric N deposition after the 

middle of the 20th century initially raised few concerns 

about detrimental ecosystem impacts. Many terrestrial 

ecosystems are N limited. Hence, additional N inputs 

could have a fertilizing eff ect, which was perceived 

as benefi cial for some ecosystems. Increased tree 

growth due to N deposition has been demonstrated 

in northeastern U.S. forests (Th omas et al. 2010). 

However, substantial debate continues over the 

existence and magnitude of deposition-induced growth 

enhancement in forests globally (DeVries et al. 2008, 

Magnani et al. 2007, Nadelhoff er et al. 1999, Sutton 

et al. 2008). Furthermore, elevated N inputs can lead 

to detrimental eff ects on ecosystems (Nihlgård 1985), 

including soil and surface water acidifi cation, plant 

nutrient imbalances, declines in plant health, changes 

in species composition, increases in invasive species, 

increased susceptibility to secondary stresses such as 

freezing, drought, and insect outbreaks, as well as 

eutrophication of fresh and coastal waters (Galloway et 

al. 2003). High concentrations of ammonia (NH3) can 

be directly toxic to plants (Krupa 2003).

Nitrogen saturation can be defi ned as the condition 

when available N exceeds plant and microbial demand. 

Aber et al. (1989, 1998) described four stages of N 

saturation in forest ecosystems (Fig. 1.1): background 

conditions (stage 0), an initial fertilization response 

(stage 1), fl attened response of N mineralization but 

increased net nitrifi cation (stage 2), and detrimental 

eff ects on plant health and growth and general decrease 

1.1 Objectives

Th is publication provides a scientifi c synthesis of the 

current state of research and knowledge about the 

response of terrestrial and aquatic ecosystems to nitrogen 

(N) inputs (N deposition or N additions), and, where 

possible, identifi es critical loads for atmospheric N 

deposition. It also targets policy makers and resource 

managers who are seeking a scientifi c basis for making 

decisions about the potential broad ecological impacts of 

air pollution, as well as the impact of specifi c pollution 

sources on particular ecosystems.

1.2 Background

1.2.1 Historical N Deposition

Human activity in the past century has led to an 

exponential increase in N emissions (Galloway et al. 

2003). Primary sources of anthropogenic N emissions 

are high temperature combustion of fossil fuels 

(i.e., electric power plants, motor vehicles), and the 

production and use of fertilizers and animal manure 

(Galloway et al. 2003). Deposition of N has paralleled 

emissions (Butler et al. 2003; US EPA 2009a, 2009b) 

so that, in most areas, current rates of N deposition 

are an order of magnitude higher than pre-industrial 

levels (Dentener et al. 2006, Holland et al. 1999). 

Pre-industrial atmospheric N deposition is thought to 

have been 0.4 to 0.7 kg ha-1 y-1 (Holland et al. 1999). 

Contemporary N deposition levels in the United States 

are on average one-fi fth of those observed in Europe 

(Holland et al. 2005). Wet N deposition reported by 

the National Atmospheric Deposition Program (NADP) 

ranges from <1 to >7 kg ha-1 y-1 for 2008 (NADP 

2008), and dry deposition can supply large quantities of 

additional N, but is far more diffi cult to quantify (e.g., 

Dentener et al. 2006). Regional maximum values can 

be substantially higher than the wet deposition reported 

by NADP. In the Northeast, maximum N deposition of 

30 kg ha-1 y-1 has been reported for high elevation sites 

(Miller 2000); in the Southeast, deposition was modeled 

as high as 31 kg N ha-1 y-1 (Weathers et al. 2006); and 

in the Pacifi c Southwest, where dry deposition and fog 
2 Chapter 1—Introduction GTR-NRS-80

in N retention (stage 3). When vegetation is no longer 

N limited, but before damage is incurred, nitrate (NO3

-

) 

immobilization will be reduced and NO3

-

 export is likely 

to increase gradually. Th e increase in NO3

-

 export may 

occur earlier in this progression of N saturation than 

initially thought (Emmett 2007). Adverse consequences 

on plant health and growth occur in the last stage of 

N saturation, resulting from some combination of 

insuffi cient allocation of plant carbon to roots and 

mycorrhizae and soil acidifi cation induced by NO3

-

leaching. Nutrient imbalances (e.g., elevated N:Ca 

or N:Mg) may also aff ect plant health. In addition, 

alterations in an ecosystem’s N status may lead to 

increased susceptibility to secondary stresses, including 

freezing injury (Schaberg et al. 2002), pest outbreak, 

and drought (Bailey et al. 2004).

At the catchment level, N saturation is indicated by 

increased NO3

-

 leaching, especially during the growing 

season (Aber et al. 2003, Stoddard 1994). Stoddard 

(1994) described the stages of surface water response to 

N saturation (Fig. 1.2).

At Stage 1, seasonal losses of NO3

-

 occur during periods 

of hydrologic fl ushing (e.g., snowmelt); in stage 2, 

episodic NO3

-

 losses remain high, and growing season 

losses increase as biotic demand for N is depressed; 

in stage 3, the base fl ow NO3

-

 concentration becomes 

elevated to such an extent that seasonal patterns are no 

longer signifi cant (Stoddard 1994).

Among the most signifi cant indicators of N saturation 

are changes in species composition or community 

structure in terrestrial and aquatic ecosystems. 

Alterations in species composition may result from shifts 

in dominance of species present, for epiphytic lichens, 

for example (Geiser and Neitlich 2007, Jovan 2008, 

McCune and Geiser 2009), or increased dominance 

of invasive species, and may lead to signifi cant changes 

in vegetation. Nitrogen deposition may aff ect species 

richness and has been implicated in dramatic declines 

in species richness, for example, in coastal sage scrub1

, 

grasslands (Stevens et al. 2004).

1.2.3 Critical Loads of N Deposition

A critical load is the level of input of a pollutant below 

which no harmful ecological eff ect occurs over the long 

Figure 1.1—Stages of N saturation for forest ecosystems (reprinted with permission in modifi ed 

format from Aber et al. 1998).

0 1 2 3

Stage

0

100

200

Nitrification

NPP

Foliar N

N Mineralization

NO3

- Leaching

N2O Efflux

Ca: Al and 

Mg: N 

ratios

1

Allen, E.B. Unpublished data. Professor and natural resources 

extension specialist, Department of Botany and Plant 

Sciences and Center for Conservation Biology, University of 

California, Riverside, CA 92521.
Chapter 1—Introduction GTR-NRS-80 3 Figure 1.2—Stages of N saturation for surface water (reprinted with permission in modifi ed format from Stoddard 1994, as published by the American Chemical Society).
4 Chapter 1—Introduction GTR-NRS-80

term (UBA 2004). Critical loads have been defi ned 

for Europe (Posch et al. 1995, Posch et al. 2001) and 

have been used as a tool in the process of negotiating 

air pollution emission reductions in Europe. In the 

United States, critical loads have been calculated for 

specifi c regions (Dupont et al. 2005, Fenn et al. 2008, 

NEG ECP 2003, Williams and Tonnessen 2000), and 

are of interest to policy makers for assessing emission 

reduction programs and to resource managers as a 

tool to evaluate the potential impact of new pollution 

sources (Burns et al. 2008, Porter et al. 2005).

Th ere are three main approaches for calculating critical 

loads: empirical, simple mass balance, and dynamic 

modeling approaches (Pardo 2010). Empirical 

approaches are based on observations of the response 

of an ecosystem or ecosystem component (e.g., foliage, 

lichens, soil) to a given, observed deposition level. 

Empirical critical loads are calculated using data 

obtained from a single site or multiple sites; generally, 

they are applied to similar sites where such data are not 

available.

Simple mass balance approaches are based on estimating 

the net loss or accumulation of N based on inputs and 

outputs. Simple mass balance methods are steady-state

models that calculate the critical load of deposition to 

an ecosystem over the long term (i.e., one rotation in 

land managed for timber, 100+ years in wilderness). 

Th ey are based on defi ning acceptable values for fl uxes 

into and out of the ecosystem (soil N accumulation, N 

leaching, or acceptable/critical NO3

-

 concentration in 

leachate).

Dynamic models use a mass balance approach 

expanded by incorporating internal feedbacks, such as 

accumulation of N in the system. Dynamic models can 

predict the amount of time it takes for N deposition 

to damage an ecosystem, as well as the amount of time 

necessary for recovery. Dynamic models that capture 

the complexity of the N cycle are currently unavailable 

for broad-scale use across multiple ecosystems in the 

United States. Th is review focuses on the data available 

to identify empirical critical loads for ecoregions of the 

United States.

1.3 Report Organization

Th is document is organized by ecoregion. Th e 

ecoregions are described in detail in Chapter 2. Issues 

and challenges in estimating total N deposition are 

described in Chapter 3. Th e methods are described in 

Chapter 4 for mycorrhizal fungi, lichens, herbaceous 

plants and shrubs, and forest ecosystems. For other 

receptors or ecosystem types, the methods are described 

in the individual chapters. Explanation of the response 

parameters and details of how the critical load was set 

are given in each ecoregion chapter. For each ecoregion, 

the ranges of responses reported are presented, and 

estimates for critical loads are made, where possible; the 

critical load values are compared to values presented for 

similar ecosystems types outside the United States (when 

possible); and the steps necessary to identify or improve 

critical loads estimates are described.

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Posch, M.; de Smet, P.A.M.; Hettelingh, J.-P.; 

Downing, R.J., eds. 2001. Modelling and mapping 

of critical thresholds in Europe. Status report 

2001. RIVM Rep. No. 259101010. Bilthoven, 

Th e Netherlands: Coordination Center for Eff ects, 

National Institute for Public Health and the 

Environment. Available at http://www.rivm.nl/

bibliotheek/digitaaldepot/PBL_CCE_SR01.pdf. 

(Accessed May 12, 2010).

Schaberg, P.G.; DeHayes, D.H.; Hawley, G.J.; 

Murakami, P.F.; Strimbeck, G.R.; McNulty, S.G. 

2002. Eff ects of chronic N fertilization on foliar 

membranes, cold tolerance, and carbon storage 

in montane red spruce. Canadian Journal of Forest 

Research. 32:1351-1359.

Stevens, C.J.; Dise, N.B.; Mountford, J.O.; Gowing, 

D.J. 2004. Impact of nitrogen deposition on the 

species richness of grasslands. Science. 303: 1876-

1879.

Stoddard, J. 1994. Long-term changes in watershed 

retention of nitrogen: its causes and aquatic 

consequences. In: Baker, L.A., ed. Environmental 

chemistry of lakes and reservoirs. Washington, DC: 

American Chemical Society: 223-284.
Chapter 1—Introduction GTR-NRS-80 7

Sutton, M.A.; Simpson, D.; Levy, P.E.; Smith, 

R.I.; Reis, S.; van Oijen, M.; deVries, W. 2008. 

Uncertainties in the relationship between 

atmospheric nitrogen deposition and forest carbon 

sequestration. Global Change Biology. 14: 1-7.

Th omas, R.Q.; Canham, C.D.; Weathers, K.C.; 

Goodale, C.L. 2010. Increased tree carbon storage 

in response to nitrogen deposition in the US.

Nature Geoscience. 3: 13-17.

UBA (UmweltBundesAmt). 2004. Manual on 

methodologies and criteria for modeling 

and mapping critical loads and levels and air 

pollution eff ects, risks, and trends. Berlin: Federal 

Environmental Agency (UmweltBundesAmt). 

Available at http://icpmapping.org/cms/

zeigeBereich/11/manual-english.html (Accessed May 

12, 2010).

US EPA (U.S. Environmental Protection Agency). 

2009a. Clean air markets - data and maps.

Washington, DC: U.S. Environmental Protection 

Agency. Available at http://camddataandmaps.epa.

gov/gdm (Accessed May 12, 2010).

US EPA (U.S. Environmental Protection Agency). 

2009b. Clean air markets - progress reports.

Washington, DC: U.S. Environmental Protection 

Agency. Available at http://www.epa.gov/airmarkets/

progress/progress-reports.html (Accessed May 12, 

2010).

Weathers, K.C.; Simkin, S.M.; Lovett, G.M.; Lindberg, 

S.E. 2006. Empirical modeling of atmospheric 

deposition in mountainous landscapes. Ecological 

Applications. 16(4): 1590-1607

Williams, M.W.; Tonnessen, K.A. 2000. Critical loads 

for inorganic nitrogen deposition in the Colorado 

Front Range, USA. Ecological Applications. 10(6): 

1648-1665.
8
Chapter 2—Ecosystem Classifi cation GTR-NRS-80 9

2 ECOSYSTEM CLASSIFICATION

M.J. Robin-Abbott and L.H. Pardo

Major river valleys support scattered clumps of stunted 

spruce trees. Typical shrubs include dwarf birch (Betula 

nana), willows (Salix spp.), and heath species commonly 

mixed with various herbs and lichens. Wetlands are 

common in the low-lying areas, mainly supporting sedge 

and moss covers.

Taiga

Th e Taiga ecoregion contains elements of both Tundra 

and Northern Forests ecoregions. It is characterized by 

innumerable lakes, bogs, other wetlands, and northern 

boreal forests, interwoven with open shrublands and 

sedge meadows more typical of the tundra. From south 

to north, forests become open and form woodlands with 

a characteristic groundcover of lichens, which merge 

into areas of tundra. Along the northern edge of this 

ecological region the latitudinal limits of tree growth 

are reached and dwarf birch, Labrador tea (Ledum spp.), 

willow, bearberry (Arctostaphylos alpina), mosses, and 

sedges are dominant. Further south, the region contains 

open stands of stunted black spruce (Picea mariana) 

and jack pine (Pinus banksiana) accompanied by alder 

(Alnus spp.), willow, and tamarack (Larix larcinia) in 

the fens and bogs. Mixed wood associations of white 

(Picea glauca) and black spruce, lodgepole pine (Pinus 

contorta), quaking aspen (Populus tremuloides), balsam 

poplar (Populus balsamifera), and white birch (Betula 

papyrifera) are found on well drained and warm upland 

sites, as well as along rivers and streams. Along the 

nutrient-rich alluvial fl ats of the larger rivers, white 

spruce and balsam poplar grow to sizes comparable to 

the largest in the boreal forests to the south.

Northern Forests

Over 80 percent forested, the ecoregion is dominated 

by conifers, largely white and black spruce, jack pine, 

balsam fi r (Abies balsamea), and tamarack. Toward the 

south and the Canadian Maritimes provinces, there 

is a wider distribution of white birch, quaking aspen, 

balsam poplar, eastern white pine (Pinus strobus), red 

pine (Pinus resinosa), sugar maple (Acer saccharum), 

Th e ecosystem classifi cation in this report is based on 

the ecoregions developed through the Commission for 

Environmental Cooperation (CEC) for North America 

(CEC 1997). Only ecosystems that occur in the United 

States are included. CEC ecoregions are described, 

with slight modifi cations, below (CEC 1997) and 

shown in Figures 2.1 and 2.2. We chose this ecosystem 

classifi cation system because it is easily accessible, 

hierarchical, and based on biological, rather than purely 

climatic characteristics. An ecoregion is a geographically 

defi ned area that contains distinct physical features 

and biological communities. Th e descriptions include 

the whole ecoregion as it occurs in North America. 

Descriptions in the individual chapters refl ect the 

ecoregion as represented in the United States.

Within the ecoregions, where data are available, we 

present the results by subcategory (e.g., for forest, 

chaparral and coastal sage scrub ecosystems within 

the Mediterranean California ecoregion). Certain life 

forms that occur across many ecosystem types, such as 

lichens and mycorrhizal fungi, are presented within each 

ecosystem for which data are available, and also, in some 

detail, in Chapter 19. Common names for species are 

used throughout this report; scientifi c names are given 

at fi rst mention in each chapter, and are also provided in 

Appendix 1. Wetlands and inland surface waters occur 

across most ecoregions in the United States and are 

addressed extensively in Chapters 17 and 18. Ecoregion 

chapters include a cursory discussion of wetlands and 

surface water where data are available.

2.1 Ecosystem Descriptions

Tundra

Th is ecoregion is a mosaic of alpine meadows, foothills, 

mesas, low-lying coastal plains, river corridors, and 

deltas. A key feature is that soils are frozen in permafrost 

and contain large stores of organic carbon. Vegetation 

is characterized by dwarf shrubs that decrease in size 

moving north, with very low and fl attened plants being 

most characteristic of the northern and central locales. 
10 Chapter 2—Ecosystem Classifi cation GTR-NRS-80

Figure 2.1—Ecological Regions of North America, Level I. From the Commission for Environmental Cooperation (1997).

Level I Regions

Arctic Cordillera

Tundra

Taiga

Hudson Plains

Northern Forests

Northwestern Forested Mountains

Marine West Coast Forests

Eastern Temperate Forests

Great Plains

North American Deserts

Mediterranean California

Southern Semi-Arid Highlands

Temperate Sierras

Tropical Dry Forests

Tropical Humid Forests

500 0 500 1,000 1,500 2,000 km

1

5

4

3

2

11

15

14

13

12

6

10

9

8

7
Chapter 2—Ecosystem Classifi cation GTR-NRS-80 11

Figure 2.2—Ecological Regions of North America, Level II. From the Commission for Environmental Cooperation (1997).

6.2

9.4

9.3

8.3

9.2

10.1

3.1

8.1

8.4

8.5

8.2

2.2

2.3

7.1

5.2

13.1

1.1

5.3

15.4

6.1

2.1

3.4

5.1

Ecoregions

 2.2 Alaska Tundra

 2.3 Brooks Range Tundra

 3.1 Alaska Boreal Interior

 3.2 Taiga Cordillera

 5.2 Mixed Wood Shield

 5.3 Atlantic Highlands

 6.1 Boreal Cordillera

 6.2 Western Cordillera

 7.1 Marine West Coast Forest

 8.1 Mixed Wood Plains

 8.2 Central Plains

 8.3 Southeastern Plains

 8.4 Ozark, Ouachita-Appalachian Forest

 8.5 Mississippi Alluvial & SE Coastal Plains

 9.2 Temperate Prairies

 9.3 West-Central Semi-Arid Prairies

 9.4 South Central Semi-Arid Prairies

 9.5 Texas-Louisiana Coastal Plain

 9.6 Tamaulipas-Texas Semi-Arid Plain

10.1 Cold Deserts

10.2 Warm Deserts

11.1 Mediterranean California

12.1 Western Sierra Madre Piedmont

13.1 Upper Gila Mountains

15.4 Everglades

Ecoregions of the US

Level II

500 0 500 1,000 1,500 2,000 km

8.4

8.5

10.2

10.1

7.1

6.2

8.1

8.1

8.3

6.2

6.1

7.1

11.1

10.2

9.5

9.6

5.3
12 Chapter 2—Ecosystem Classifi cation GTR-NRS-80

beech (Fagus spp.), red spruce (Picea rubens), and 

various species of oak (Quercus spp.). Areas of shallow 

soils and exposed bedrock are common and tend to be 

covered with a range of plant communities, dominated 

by lichens, shrubs and forbs. It is underlain by ancient 

Canadian Shield bedrock interspersed with glacial 

moraine deposits, creating a hilly terrain dotted with 

numerous lakes. Soils derived from the bedrock are 

generally coarse textured and nutrient-poor. Peatlands 

are extensive in northern Minnesota. Th e climate is 

characterized by long, cold winters (means of -20.5 °C in 

the west to -1 °C in the east) and short, warm summers 

(means of 11 to 18 °C). Annual precipitation ranges 

from 400 to 1,000 mm.

Northwestern Forested Mountains

Vegetative cover is extremely diverse, with distinct 

elevational bands: alpine environments contain various 

herb, lichen, and shrub associations; whereas the 

subalpine forest has species such as lodgepole pine (Pinus 

contorta), subalpine fi r (Abies lasiocarpa), and Engelmann 

spruce (Picea engelmannii) in interior assemblages; 

and Pacifi c silver fi r (Abies amabilis), grand fi r (Abies 

grandis), lodgepole pine, and Engelmann spruce in 

the west. Mid-elevation forests are characterized by 

ponderosa pine (Pinus ponderosa), Rocky Mountain 

Douglas-fi r (Pseudotsuga menziesii var. glauca), lodgepole 

pine, and quaking aspen in much of the southeast 

and central portions; and western hemlock (Tsuga 

heterophylla), western red cedar (Th uja plicata), Douglasfi r (Pseudotsuga menziesii ) and western white pine 

(Pinus monticola) in the west and southwest. White and 

black spruce dominate the plateaus of the north. Shrub 

vegetation found in the dry southern interior includes 

big sagebrush (Artemisia tridentata), rabbit brush 

(Chrysothamnus spp.), and antelope bitterbrush (Purshia 

tridentata). Most of the natural grasslands that existed in 

the dry south have been replaced by agriculture.

Marine West Coast Forests

 Variations in altitude create widely contrasting 

ecological zones within the region. Th ey range from 

mild, humid coastal rain forest to cool boreal forests and 

alpine conditions at higher elevations. Th e maritime 

infl uence of the Pacifi c Ocean causes high precipitation 

(600 to 5000 mm), a long growing season, and 

moderate mean annual temperatures (5 to 9 °C). Th e 

temperate coastal forests are composed of mixtures of 

western red cedar, Nootka Cypress (Alaska yellow cedar; 

Callitropsis nootkatensis), western hemlock, Douglasfi r, Pacifi c silver fi r, white spruce, Sitka spruce (Picea 

sitchensis), California redwood (Sequoia sempervirens), 

and red alder (Alnus rubra). Many of these trees reach 

very large dimensions and live to great age, forming 

ancient or old growth. In the drier rain-shadow areas, 

Oregon white oak (Garry oak; Quercus garryana) and 

Pacifi c madrone (Arbutus menziesii) occur with Douglasfi r. Sub-alpine forests are characterized by mountain 

hemlock (Tsuga mertensiana) and Pacifi c silver fi r. 

Alpine tundra conditions are too severe for growth of 

most woody plants except in dwarf form. Th is zone is 

dominated by shrubs, herbs, mosses, and lichens.

Eastern Temperate Forests

Th e Eastern Temperate Forests form a dense forest 

canopy consisting mostly of tall, broadleaf, deciduous 

trees and needle-leaf conifers. Beech-maple (Fagus-Acer)

and maple-basswood (Acer-Tilia) forest types occur 

widely, especially in the eastern reaches of this region; 

mixed oak-hickory (Quercus-Carya) associations are 

common in the Upper Midwest, changing into oakhickory-pine (Quercus-Carya-Pinus) mixed forests in 

the south and the Appalachians. Th ese forests have a 

diversity of tree, shrub, vine, and herb layers. While 

various species of oaks, hickories (Carya spp.), maples 

(Acer spp.), and pines (Pinus spp.) are common, other 

wide-ranging tree species include ashes (Fraxinus spp.), 

elms (Ulmus spp.), black cherry (Prunus serotina), 

yellow-poplar (Liriodendron tulipifera), sweet gum 

(Liquidambar styracifl ua), basswood (Tila spp.), 

hackberry (Celtis occidentalis), persimmon (Diospyros 

virginiana), eastern red cedar (Juniperus virginiana), and 

fl owering dogwood (Cornus fl orida). A key tree species, 

the American chestnut (Castanea dentata), was virtually 

eliminated from the Eastern Temperate Forests in the 

fi rst half of the 20th century by an introduced fungus.

Great Plains

Th e Great Plains ecoregion was once covered with 

natural grasslands that supported rich and highly 
Chapter 2—Ecosystem Classifi cation GTR-NRS-80 13

specialized plant and animal communities. Th e 

interaction of climate, fi re, and grazing infl uenced the 

development and maintenance of the Great Plains. 

Rainfall increases from west to east, defi ning diff erent 

types of native prairies. Short-grass prairie occurs in 

the west, in the rain shadow of the Rocky Mountains, 

with mixed-grass prairie in the central Great Plains and 

tall-grass prairie in the wetter eastern region. Because 

of the suitability of the Great Plains for agricultural 

production, many native prairie vegetation types have 

been radically altered. Th e short-, mixed-, and tallgrass 

prairies now correspond to the western rangelands, the 

wheat belt, and the corn/soybean regions, respectively. 

Drier sites are home to prickly pear (Opuntia spp.), with 

sagebrush (Artemisia spp.) also abundant.

Th e Aspen Parkland, the northern transition zone to the 

boreal forest, has expanded south into former grasslands 

since settlement eff ectively stopped prairie fi res. In the 

United States, native prairie vegetation ranges from 

grama grass (Bouteloua spp.), wheatgrass (Agropyron

spp.), and bluestem (Andropogon spp.) prairie in the 

north to diff erent shrub and grassland combinations 

(e.g., mesquite-acacia (Prosopis-Acacia) savanna and 

mesquite-Texas live oak (Prosopis-Quercus fusiformis)

savanna) and grassland and forest combinations (e.g., 

juniper-oak [Juniperus-Quercus] savanna and mesquitebuff alo grass [Prosopis-Buchloe dactyloides]) in the 

south. Th ere are also patches of blackland prairie, 

bluestem-sacahuista (Andropogon-Nolina) and southern 

cordgrass (Spartina) prairie in the southern United 

States. Th e eastern border of the region, stretching from 

central Iowa to Texas, shows patterns of grassland and 

forest combinations mixed with oak-hickory forest. 

Th roughout the remainder of the Great Plains there 

are few native deciduous trees that occur, except in 

the eastern regions or in very sheltered locations along 

waterways or at upper elevations.

North American Deserts

In this ecological region of altitudinal, latitudinal, and 

landform diversity, there are a variety of vegetation 

types, but low-growing shrubs and grasses predominate. 

In the northern Palouse and Snake River Basin areas, 

grasslands and sagebrush steppes were once common. 

However, most of these northern grasslands have 

been converted to agriculture, and in some areas, the 

sagebrush steppe is being invaded by western juniper 

(Juniperus occidentalis) and cheatgrass (Bromus tectorum). 

Th e Great Basin and greater Colorado Plateau are 

characterized by sagebrush (Artemisia spp.), with 

saltbush (Atriplex spp.) and greasewood (Sarcobatus spp.) 

on more alkaline soils.

In the warmer southern deserts, creosote bush (Larrea 

tridentata) is common; the Mojave Desert also contains 

areas of the distinctive Joshua tree (Yucca brevifolia). 

Th e Sonoran Desert has greater structural diversity in 

its vegetation than the other North American deserts. 

Its paloverde (Parkinsonia spp.), cactus, and shrub 

vegetation includes various types of succulents, such 

as saguaro (Cereus gigantea), cholla (Cylindropuntia

spp.), and agave (Agave spp.). Th e Chihuahuan Desert 

is characterized by smaller leaved, shorter statured 

vegetation such as American tarwort (Flourensia cernua), 

creosote bush, and intermixed grasses.

Mediterranean California

Th e Mediterranean California region is characterized 

by a mostly evergreen shrub vegetation called chaparral, 

plus patches of oak woodland, grassland, and coniferous 

forest on upper mountain slopes. Th e chaparral has 

thickened, hardened foliage resistant to water loss, 

and forms a cover of closely spaced shrubs 1 to 4 m 

tall. Common shrubs include chamise (Adenostoma 

fasciculatum), buckbrush (Ceanothus cuneatus), and 

manzanita (Arctostaphylos spp.). Coastal sagebrush 

(Artemisia californica), summer-deciduous plants 

that tolerate more xeric, or dry, conditions than the 

evergreen chaparral, are found at lower elevations. 

About 80 percent of the presettlement coastal sage scrub 

in southern California has been displaced, primarily by 

residential development and agriculture. Several dozen 

threatened and endangered species occur in the coastal 

sage scrub community. To the north, the chaparral is 

less continuous, occurring in a mosaic with grassland, 

as well as broadleaf and coniferous forests. A blue 

oak-California foothills pine (Quercus douglasii –Pinus 

sabiniana) woodland community forms a ring around 

the Central Valley, which itself once had extensive 

grasslands and riparian forests. Th e southern oak 

woodland extends into the transverse and peninsular 
14 Chapter 2—Ecosystem Classifi cation GTR-NRS-80

ranges and includes California walnut (Juglans 

californica) and Engelmann oak (Quercus englemannii). 

Th e high elevation mixed pine forests are dominated 

by Jeff rey pine (Pinus jeff reyi) and ponderosa pine, 

with incense cedar (Calocedrus decurrrens), black oak 

(Quercus velutina), white fi r (Abies concolor), and others.

Southern Semi-Arid Highlands

Th e characteristic natural vegetation, which presently 

is very diminished or altered, consists of grasslands 

and combinations of grasslands with scrublands, 

and woodlands and forests in the higher elevations. 

Certain species of grasses are dominant, particularly 

bluestem, threeawn (Aristida spp.), galleta (Hilaria 

jamesii), and muhly (Muhlenbergia spp.) grasses. 

Among the shortgrasses, blue grama (Bouteloua gracilis) 

is important in Arizona, New Mexico, and in the 

northern Mexican states at the foot of the Western 

Sierra Madre. On deep clay soils, mesquite groves are 

the most conspicuous plant community, and mesquite 

occurs with acacia in some sites, especially in Mexico. 

Oak and western juniper are common at the foot of the 

Sierras. Where the climate is warmer in Mexico, one 

fi nds subtropical scrublands with a diversity of shrub 

species.

Temperate Sierras

 Vegetation can be evergreen or deciduous, composed 

primarily of conifers and oaks. Th ey grow from 10 to 

30 m, sometimes reaching 50 m. Th is vegetative cover 

may comprise from one to three tree layers, one or two 

shrub layers, and an herbaceous stratum. Th is forest 

community is characterized by about 3,000 vascular 

plant species, 30 percent of which are endemic to 

Mexico. Overall most of this vegetation type occurs in 

Mexico, where there is a high diversity of oaks, pine 

and other species, and includes mountain cloud forests 

that do not occur in the United States.

Tropical and Subtropical Humid Forests

Th is ecoregion includes tropical humid forests at the 

southern tip of the Florida Peninsula in the United 

States, as well as subtropical humid forests in coastal, 

low-elevation, and lower montane regions on the 

Hawaiian Islands and in Puerto Rico. Th e region 

spans from sea level to approximately 1,800 m of 

altitude. For subtropical and tropical humid forests, 

including those in the lower montane regions, average 

annual precipitation typically exceeds 1,000 mm and 

year round temperatures average between 16 °C and 

26 °C. Evergreen and semi-deciduous forests are the 

most characteristic plant communities of this region, 

which is one of the most biodiverse zones in the 

world. Forest stands are typically of mixed ages with 

a great abundance of air plants (epiphytes), including 

bromeliads, ferns, and orchids. Th e mature tree layer 

may attain heights up to 30 m or more. In the tropical 

humid forest of the Florida peninsula, fl ooded marshes 

and swamps (both saltwater and freshwater) are 

widespread, with characteristic mangrove vegetation 

and tropical tree islands. Native subtropical humid 

forests in Hawaii include the common dominant trees 

Ohi’a lehua (Metrosideros polymorpha) and Koa (Acacia 

koa), as well as tree ferns and a diverse understory of 

epiphytes and shrubs. Puerto Rican forests on volcanic 

soils include Tabonuco, Colorado, and dwarf cloud 

forests; sclerophyllous, evergreen vegetation is found on 

serpentine and limestone derived soils.

LITERATURE CITED

CEC (Commission for Environmental Cooperation). 

1997. Ecological regions of North America. Toward 

a common perspective. Montreal, Canada: 

Commission for Environmental Cooperation. 

71 p. Available at http://www.cec.org/fi les/pdf/

BIODIVERSITY/eco-eng_EN.pdf (Accessed May 

24, 2010).
Chapter 3—Deposition GTR-NRS-80 15

3 DEPOSITION

K.C. Weathers and J.A. Lynch

Th e purposes of this chapter are threefold. First, 

it presents an overview of atmospheric deposition 

to ecosystems and provides a brief summary of the 

processes and challenges of measuring and estimating 

total (wet, dry, cloud, or fog) atmospheric deposition. 

Second, we present a map (Fig. 3.1) of wet plus dry 

deposition for the more than 3200 U.S. locations for 

which we report ecological responses to N deposition in 

subsequent chapters. Th ird, we compare two methods 

of estimating deposition at the national scale (Weathers 

and Lynch 2008).

3.1 OVERVIEW

To determine the eff ects of air pollution on ecological 

systems using the critical load approach, accurate estimates 

of total nitrogen (N) deposition are essential. Empirical 

critical loads are set by relating observed ecosystem 

responses to N deposition (measured, experimentally 

manipulated, or modeled). Because many manipulation 

studies are done under natural conditions where the total 

N deposition is not known, the empirical critical load 

is often crudely estimated. However, making accurate 

measurements or models of N deposition is diffi cult, 

and such estimates are not available at the fi ne spatial 

resolution required for determining empirical critical loads.

Figure 3.1—Estimates of wet plus dry N deposition (including, as N, wet ammonium (NH4

+

) and nitrate (NO3

-

), and dry 

nitric acid (HNO3), ammonia (NH3), NH4

+

, and NO3

-

) generated by the CMAQ 2001 model for more than 3200 locations 

across the United States for which we report ecological responses to N deposition in subsequent chapters.
16 Chapter 3—Deposition GTR-NRS-80

3.2 Background on total N deposition

Th e primary forms of inorganic N deposition that 

have been measured routinely are secondary products 

of nitrogen oxide (NOx) emissions—nitrate (NO3

-

; 

particulate and aqueous) and nitric acid vapor (HNO3) 

—and ammonium (NH4

+

; particulate and aqueous). 

Th e primary anthropogenic source of NOx is fossil fuel 

combustion. Transportation accounted for more than 

50 percent of the total NOx emissions in 2002 (Driscoll 

et al. 2003, Weathers et al. 2007), while the primary 

sources of ammonia (NH3) and NH4

+

 are agricultural 

activities (Driscoll et al. 2003, Weathers et al. 2007). 

(Chemical formulas and common names for elements 

and compounds are provided in Appendix 2.)

Nitrogen is deposited from the atmosphere to the 

Earth’s surface in the form of wet (rain and snow), 

dry (gases and particles), and cloud and/or fog 

deposition (Fenn et al. 2009; Lovett 1994; Weathers 

et al. 1988, 2000a, 2000b, 2006). Wet N deposition 

can be measured directly, but often only some of the 

components are measured. Dry deposition cannot be 

measured directly. Moreover, the estimates that exist are 

either site specifi c, or are modeled (e.g., dry deposition) 

assuming fl at, homogeneous terrain, e.g. Clean Air 

Status and Trends (CASTNET) sites (US EPA 2010). 

For this reason, a combination of monitoring data, 

deposition models (e.g., the Multi-Layer Model [MLM] 

for estimating dry deposition), empirical measures 

of atmospheric deposition (e.g., throughfall), and 

emissions-based modeling outputs, are often necessary 

to estimate total N deposition. In the ecoregion chapters 

that follow, we consider total deposition to include wet 

(rain and snow) plus dry and, in some cases, cloud/fog.

Th ere are several factors that reduce the accuracy of total 

N deposition estimates:

• Not all forms (wet, dry, and cloud) of 

atmospheric deposition are measured or 

modeled routinely. Th us, estimates that are 

based on wet deposition only, for example, 

result in an underestimation of the total 

N delivered to an ecosystem (Lovett 1994, 

Weathers et al. 2007). Dry and cloud N 

deposition can contribute from 50 to 80 

percent of total N deposition (Johnson and 

Lindberg 1993); these forms of deposition are 

not often estimated.

• Not all of the chemical species of N that are 

deposited to ecosystems (e.g., NH3

 and organic 

N) are currently measured and/or modeled, 

which also leads to an underestimate of N 

deposition.

• Estimates of dry and cloud deposition depend 

upon measured air chemistry and modeled 

deposition. Th ese estimates are therefore likely 

to be imprecise. For example, the infl uences of 

local meteorology, complex terrain (elevation 

and slope), as well as canopy architecture 

all aff ect rates of dry deposition, yet these 

characteristics of landscapes and canopies are 

extremely diffi cult to represent in models, and 

thus represent serious limitations to accurate 

estimates of dry N deposition (Lovett 1994; 

Weathers et al. 1992, 1995, 2000a, 2000b, 

2006). Total deposition is therefore often 

underestimated for high elevation regions and/

or for complex terrain (Anderson et al. 1999; 

Johnson and Lindberg 1993; Weathers et al. 

1992, 2000a, 2000b, 2006).

• Estimates of deposition are also highly 

uncertain for regions with high snowfall, such 

as the montane west, and regions with a high 

proportion of dry deposition, such as the arid 

south and west (see Fenn et al. 2009 for a 

complete description of issues and potential 

solutions for these regions).

• Finally, inadequate number and spatial extent 

of monitoring stations is a crucial issue that 

adversely impacts deposition estimates across 

space, particularly in complex terrain. Th is is 

a signifi cant problem in the western United 

States (Fenn et al. 2009), where few monitoring 

stations exist.

Inorganic species of N most commonly measured or 

modeled in the fi eld (Fig. 3.2) include: NO3

-

, NH4

+

, and 

nitric acid (HNO3

). In recent years, dissolved organic 

nitrogen (DON) has also been shown to be prevalent 

in precipitation (e.g., Neff et al. 2002) and in clouds 
Chapter 3—Deposition GTR-NRS-80 17

(Weathers et al. 2000a). Other potentially important 

N species not addressed here, primarily because they 

are not measured at National Atmospheric Deposition 

Program (NADP) or CASTNET monitoring stations 

(see descriptions, Chapter 4), are wet and dry NH3, 

NO, NO2, N2O5

, peroxyacetal nitrate (PAN) (US EPA 

2001). We note that fl uxes of these species are estimated 

by the Community Multiscale Air Quality (CMAQ) 

model, however (see description, Chapter 4). One of 

the most important species that is not measured at 

CASTNET and NADP monitoring stations is NH3

. 

Th is form of N is currently increasing in the atmosphere 

and will be critically important to measure in both the 

short and long term (Driscoll et al. 2001, 2003).

In sum, estimates of total N deposition are still highly 

uncertain for several reasons, including the facts that 

(1) deposition models cannot accommodate complex 

terrain; (2) important N species are not often measured; 

(3) snow, dry, fog, or cloud deposition are not often 

considered, yet represent potentially important vectors 

for N deposition in some locations; and (4) there is 

inadequate monitoring (e.g., Driscoll et al. 2001, Lovett 

1994, Weathers et al. 2000b, 2006).

3.3 Measurements and Estimates of 

N Deposition

3.3.1 Wet Inorganic N Deposition

Wet deposition is measured at approximately 250 

monitoring stations across the United States as part 

of the NADP network. Wet deposition is the sum 

of NO3

-

-N and NH4

+

-N delivered via rain and snow. 

Wet deposition is also estimated at various resolutions 

(e.g., 36 km x 36 km; 12 km x 12 km) via versions of 

the CMAQ model (Byun and Ching 1999, Byun and 

Schere 2006, US EPA 2009; Fig 3.1). Th e CMAQ 

model derives estimates of deposition from emissions, 

atmospheric chemistry, meteorology, and deposition 

velocities as infl uenced by land cover (US EPA 2009). 

Figure 3.2—Total (wet plus dry, including wet NH4

+

 and NO3

-

, dry HNO3, particulate NO3

-

, and NH4

+

) mean N deposition 

for 85 CASTNET site locations for 2004 to 2008. Size of circle indicates relative total deposition. Proportion of wet and 

dry species for each are shown within circle. Note that HNO3-N is the largest proportion of dry deposition.
18 Chapter 3—Deposition GTR-NRS-80

Th e chemical species we consider here from the CMAQ 

model are wet NH4

+

 and NO3

-

.

3.3.2 Cloud Deposition Monitoring

Th ere exists only one cloud deposition monitoring 

network, the Mountain Atmospheric Deposition 

Program (MADPro), where cloud water samples are 

collected and measured for chemistry (NO3

-

 and NH4

+

). 

Th ese data are subsequently used to model cloud 

deposition for a few sites in the eastern United States 

(Anderson et al. 2006, Baumgartner et al. 2003). Cloud 

deposition estimates are not routinely extrapolated 

beyond the mountain tops where they are measured, 

such as Clingman’s Dome, North Carolina, Whitetop 

Mountain, Virginia, and Whiteface Mountain, New 

York. Obtaining accurate estimates of cloud/fog 

deposition across the landscape is a signifi cant problem 

(Fenn et al. 2009; Johnson and Lindberg 1993; Lovett 

1994; Weathers et al. 1988, 2000a, 2000b, 2006). 

Where it is frequent, cloud/fog deposition can represent 

up to 80 percent of total N deposition (Anderson et al. 

1999, Fenn et al. 2009, Johnson and Lindberg 1993).

3.3.3 Dry Deposition

Dry deposition is estimated at approximately 80 

monitoring locations across the United States (US 

EPA 2010; Fig 3.2). For the ground-based CASTNET 

and AIRMoN (Atmospheric Integrated Research 

Monitoring Network) programs, air N chemistry is 

measured (HNO3

, particulate NO3

-

, and particulate 

NH4

+

), and dry deposition is modeled based on 

meteorologic conditions and multiple resistances of the 

receptor surfaces (CASTNET and AIRMoN; NADP 

2009, US EPA 2010). Because protocols diff er between 

the CASTNET and AIRMoN programs and there are 

fewer than 10 AIRMoN sites, we restrict our analysis 

to CASTNET data. Researchers have used CASTNET 

air concentration data to estimate deposition across the 

landscape (e.g., Fenn et al. 2009, Ollinger et al. 1993, 

NEG/ECP 2003, Weathers and Lynch 2008). Th e 

Environmental Protection Agency’s CMAQ model also 

estimates dry deposition based on emissions, combined 

with air chemistry, meteorology and land cover-based 

deposition velocities. Th e chemical species we consider 

here from the CMAQ model are NH4

+

-N, NO3

-

-N, and 

HNO3-N.

3.3.4 Total Deposition

Total (wet plus dry plus cloud) inorganic N deposition 

has rarely been estimated across the U.S. landscape. In 

cases in which it has, indices of deposition or estimated 

air chemistry and meteorologic conditions have been 

used in combination with statistical and spatial tools 

(e.g., geographic information systems; NEG/ECP 2003; 

Weathers et al. 2000b, 2006). Wet plus dry (not cloud) 

deposition has been estimated across the country via 

the CMAQ modeling eff ort (Fig 3.1) as well as through 

statistical techniques (Ollinger et al. 1993), and for 

CASTNET locations using CASTNET and NADP 

data (Fig 3.2). Total fl ux to the forest fl oor (including 

cloud; throughfall) has been measured directly in 

many locations using bulk precipitation samplers (e.g. 

Simkin et al. 2004, Fenn et al. 2004). By comparing 

throughfall to wet-only deposition, estimates of dry 

(plus cloud, where relevant) deposition can be made 

(e.g., Ewing et al. 2009, Lovett 1994). However, often 

it is diffi cult to interpret these data vis-à-vis atmospheric 

deposition since canopy processing (uptake, leaching, 

and transformation) of N is common (Lovett 1994, 

Weathers et al. 2006).

As noted above, organic N has been identifi ed as an 

important component of total N deposition for some 

ecosystems (Neff et al. 2002, Weathers et al. 2000a). 

However, few data are available to assess its temporal 

or spatial extent or importance in contributing to total 

deposition across the United States. Th us, throughout 

this document, unless otherwise noted, we will refer to 

total inorganic N deposition (or total N deposition) 

as the sum of NO3

-

-N, NH4

+

-N, and HNO3-N for dry 

deposition, deposited via wet and dry deposition (cloud 

deposition only when noted).

3.3.5 Temporal Trends in Wet plus Dry N 

Deposition Across the United States

Figure 3.3 shows wet plus dry N concentration and 

deposition across the United States for the 18-year 

period from 1990 through 2007, based on data from 

34 CASTNET and NADP locations (US EPA 2007). 

While no trend lines are reported for these data, a 17 

percent decrease in total inorganic N deposition over 

the period was reported. Th e proportion of wet to dry 
Chapter 3—Deposition GTR-NRS-80 19

deposition also varies commensurate with changes in 

annual precipitation; when there is more precipitation, 

wet:dry ratios are greater (US EPA 2007).

3.4 Atmospheric N Deposition Across 

the Landscape

In this report, we have used the EPA’s 2001 CMAQ 

model v. 4.3 (which use 2001 reported data and is 

hereafter referred to as CMAQ 2001 model) to provide 

comparative estimates of wet plus dry deposition to the 

more than 3200 locations included in this assessment of 

critical loads. We chose to use the CMAQ 2001 model 

(using the 36 km x 36 km resolution version) so that 

we could make comparisons across the country using 

the same method. For some regions, a 12 km x 12 km 

resolution version of CMAQ is available, but since it is 

not available for the whole country, we have not used it 

here. It is not possible to extrapolate either CASTNET 

or AIRMoN estimates of dry deposition beyond a 1 

km x 1 km region surrounding the stations because 

the parameters required for the model are site-specifi c 

(US EPA 2010). It is also well known that estimates 

of total deposition that do not suffi ciently consider 

landscape heterogeneity are likely to underestimate 

total deposition (Driscoll et al. 2001; Fenn and 

Bytnerowicz 1997; Fenn et al. 1998, 2000, 2009; 

Holland et al. 2005; Ito et al. 2002; Lovett 1994; Miller 

et al. 1993; Nanus et al. 2003; Ollinger et al. 1993; 

Weathers et al. 1992, 1995, 2000a, 2006). Th ese issues 

notwithstanding, the CMAQ 2001 model represents 

a current state-of-the-art model to create cross-site 

deposition comparisons.

3.5 CMAQ Deposition Across Critical 

Loads Sites/Comparison Between 

CMAQ and Monitoring Stations

Average annual deposition across the more than 3200 

critical load sites varied more than 30-fold, from 0.5 

to 17 kg N ha-1 yr-1 (Fig 3.1). Total N deposition 

estimates based on the CMAQ 2001 model for 

CASTNET site locations compared well to those 

generated using CASTNET dry plus NADP wet data 

(Fig. 3.4, R2

=0.75). Note that in order to compare 

these two estimates, total N deposition included wet 

NO3

-

 and NH4

+

 and dry HNO3 and NH4

+

 as N, only 

since NH3

 data were not available for CASTNET 

sites.) However, air concentrations of HNO3

-N, the 

largest proportion of dry deposition, show a consistent 

diff erence (CMAQ < CASTNET), with a positive linear 

relationship (Fig 3.5; R2

=0.68). Th is is in contrast to 

estimates of dry deposition of HNO3

-N, which are 

also positively correlated (Fig 3.6; R2

=0.45), but with 

Figure 3.3—Total (solid black line), wet (gray area), dry (black area) deposition (kg N ha-1 yr-1) 

and wet concentration (dotted line in mg L-1) for 34 sites from 1990 to 2007 as estimated and 

measured by the NADP and CASTNET programs. Figure from US EPA 2007.
20 Chapter 3—Deposition GTR-NRS-80

Figure 3.4—Comparison between wet plus dry 

deposition estimates for 2001 (includes wet 

NH4

+

 and NO3

-

, dry HNO3, particulate NO3

-

 and 

NH4

+

) from CASTNET and NADP (NADP site 

nearest to CASTNET locations) and CMAQ 

2001 model for 77 CASTNET site locations. 

y = 0.9281x + 0.6818, R2

 = 0.75. 1:1 line is 

shown (from Weathers and Lynch 2008).

Figure 3.5—Mean air concentration 

(μg m-3) of HNO3-N during 2001 measured 

at 77 CASTNET sites vs. the mean air 

concentrations generated by the CMAQ 2001 

model for those sites during 2001. y = 0.3324x 

+ 0.034, R2

 = 0.6832. The 1:1 line is shown 

(from Weathers and Lynch 2008).

Figure 3.6—N dry deposition (kg N ha-1 yr-1) 

during 2001 modeled at 77 CASTNET sites vs. 

the N dry deposition generated by the CMAQ 

2001 model for those sites during 2001. y 

= 1.008x + 0.869, R2

 = 0.45. The 1:1 line is 

shown (from Weathers and Lynch 2008).
Chapter 3—Deposition GTR-NRS-80 21

CMAQ estimates greater than CASTNET estimates, 

suggesting that dry deposition velocities applied to air 

concentration data for the two diff erent methods vary 

(these analyses based on Weathers and Lynch 2008).

We note, however, that the CMAQ 2001 model that 

was available to us at the time of this analysis could 

not resolve deposition levels below 36 km x 36 km 

grids compared to deposition that has been measured 

in the fi eld. For example, Fenn et al. (2000, 2008) 

have measured 71 kg N ha-1 yr-1 deposition in the San 

Bernardino Mountains, an area that, according to 

the CMAQ 2001 model, would receive only 3 to 5 

kg N ha-1 yr-1. Similarly, Weathers et al. (2006) have 

estimated N deposition to be as high as 31 kg N ha-1

yr-1 at Clingman’s Dome, North Carolina (Fig. 3.7), 

whereas both CMAQ and NADP plus CASTNET 

deposition for that same location and year would be 

less than 10 kg N ha-1yr-1. In these cases, the CMAQ 

estimates are several times lower than on-site measures 

of deposition that include wet plus dry plus cloud 

deposition.

In summary, the comparison of N deposition estimates 

from the CMAQ model to CASTNET plus NADP 

suggest that either method could be used to generate a 

comparative estimate of wet plus dry deposition to fl at, 

homogenous terrain (Fig. 3.4). Th e diff erences between 

CMAQ and other estimates of N deposition (Fenn et 

al. 2003, Weathers et al. 2006) are not surprising, given 

the inadequacy of dry deposition models in handling 

complex terrain and the lack of cloud deposition 

estimates (Weathers et al. 2006), as well as the coarse 

scale (36 km x 36 km) over which the CMAQ model 

predicts deposition (Weathers and Lynch 2008).

Figure 3.7—Total (wet, dry, and cloud) inorganic N (and S) deposition estimated for Great Smoky Mountain National Park 

for the year 2000 (from Weathers et al. 2006).
22 Chapter 3—Deposition GTR-NRS-80

3.6 Future Research Directions and 

Gaps in Data

Despite decades of research, accurate estimates of 

total N deposition are lacking. While various methods 

for estimating deposition (see above) exist, almost all 

methods estimate only portions of total N inputs to 

ecosystems. Not all chemical species of N are measured 

or monitored, nor are all forms of deposition estimated. 

Priorities for future deposition research include: 

expanding monitoring station spatial coverage (especially 

in high elevation and arid sites and in the West); 

measuring more chemical species of N; generating 

better estimates of dry (and, where it is an important 

component of atmospheric deposition, cloud/fog) 

inputs; creation of models and maps of deposition across 

heterogeneous landscapes that incorporate or reference 

existing modeling and/or monitoring data (e.g., Ollinger 

et al. 1993, Holland et al. 2005, Weathers et al. 2006); 

expanding the use of passive sampling devices for 

chemical species such as NH3 (e.g., Fenn et al. 2009); 

coupling empirical measurements (such as throughfall) 

with monitoring data (e.g., Weathers et al. 2006); 

and comparing estimates of deposition using diff erent 

methods.

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Chapter 4—Methods GTR-NRS-80 25

4 METHODS

L.H. Pardo, E.A. Lilleskov, L.H. Geiser, M.J. Robin-Abbott

climate, and biogeochemical factors may also infl uence 

the response. For lichens (see section 4.3), these factors 

were taken into account. For critical loads derived 

from fertilization studies, the eff ects of added N are 

isolated through controlled experimentation; however, 

the scale of these experiments is often fairly small and 

their numbers are limited, making extrapolation to the 

ecoregion level diffi cult. Another pitfall of the empirical 

approach is that the observed response may be unique to 

the site at which it was measured (because of locationspecifi c disturbance history or site characteristics such 

as soil thickness) and may not be truly representative 

of the ecosystem type. Observations used to determine 

or estimate empirical critical loads are generally made 

over a relatively short term. In contrast, the residence 

time of N within major ecosystem compartments 

(e.g., soil, sediment, forest biomass) is generally on the 

order of decades to centuries. As a result, empirical 

critical loads provide a limited understanding of true 

ecosystem response to sustained elevated N inputs or 

their response to changes in these inputs. Critical loads 

dynamic modeling addresses some of the disadvantages 

of the empirical critical loads approach. However, 

current modeling capabilities do not enable large scale 

mechanistic modeling of critical loads for ecoregions 

across the United States. Th us we utilize multiple 

sources of current empirical information (multiple 

gradient and fertilization studies) to determine critical 

loads. To address these uncertainties, we present the 

ranges of responses observed. Where we have estimated 

an empirical critical load for N, we have made some 

assessment of the reliability of that estimate. We used 

the same categories for reliability as those used in 

European assessments (Bobbink et al. 2003, UBA 2004):

• reliable ##: when a number of published papers 

of various studies show comparable results

• fairly reliable #: when the results of some studies 

are comparable

• expert judgment (#): when very few or no 

empirical data are available for this type of 

ecosystem. Th e N critical load is then based 

4.1 Approach

We present an up-to-date review of current knowledge 

relating nitrogen (N) inputs to ecological eff ects, 

from published literature, reports, and some recent 

unpublished studies. Our goal is to summarize research 

on ecosystems across the United States. We include N 

gradient, long-term, and fertilizer studies; for fertilizer 

studies, we focus on those with low N additions 

(i.e., close to ambient deposition). We include both 

publications that report a response and those that report 

the lack of response.

Empirical critical loads are determined by recording the 

deposition level at which detrimental ecological eff ects 

have occurred in the fi eld. Th e lowest deposition level 

at which a detrimental response occurs is considered the 

critical load. In cases where there is a variation in the 

level of N input that causes a given response, a range 

for the critical load is often reported (e.g., Bobbink et 

al. 2003). Empirical critical loads are useful because 

they can be used to determine the critical load based 

on the best available information for an ecosystem type 

when data are not available at a given site. Under the 

best circumstances, empirical critical loads link observed 

ecosystem response to deposition.

Disadvantages of empirical critical loads stem from their 

reliance on observed cause-and-eff ect responses rather 

than an understanding of a process or mechanism. 

Because the observed responses to a given deposition 

level are based on past deposition scenarios, they may 

not be indicative of the response that might occur in the 

future. For example, if a response is observed to a given 

rate of deposition (or fertilization) after a certain number 

of years, it is possible that a lower deposition over a 

longer time period would cause the same detrimental 

eff ect. For critical loads derived from gradient studies, 

generally, the assumption is that the N deposition 

rate is the major factor contributing to changes in 

a particular response (e.g., foliar N concentrations, 

species richness) along the gradient, even though soil, 
26 Chapter 4—Methods GTR-NRS-80

upon expert judgment and knowledge of 

ecosystems, which are likely to be more or less 

comparable with this ecosystem.

In all cases, accurate estimates of N deposition are 

essential for developing useful empirical critical loads of 

N.

In presenting the ecosystem responses and critical loads, 

we identify the response parameter and the threshold. 

Within an ecosystem, there could be very diff erent types 

of response parameters; hence the critical load or input 

which induces each response may vary considerably. 

Th is poses a challenge in setting a critical load for the 

ecosystem. In some cases, it may be appropriate simply 

to choose the lowest critical load value, while in other 

cases, there may be some responses more characteristic 

or of greater concern which should be considered the 

primary indicator for that ecosystem. Furthermore, 

comparing critical loads that were set using diff erent 

critical responses presents a challenge.

Several criteria should be determined before assigning 

empirical critical loads to an ecosystem type. Th ere are 

four key criteria to consider: what response variables are 

relevant, what types of studies are acceptable, how to 

choose the value for the critical load, and if and when 

we can extrapolate conclusions beyond the study area.

Approaches for setting empirical critical loads for 

N for specifi c ecosystem components are discussed 

below. Approaches for setting critical loads for aquatic 

ecosystem components are presented in Chapters 17 

and 18.

4.2 Mycorrhizal Fungi

4.2.1 Response Variables 

Th e impacts of N deposition on arbuscular mycorrhizal 

fungi (AMF) and ectomycorrhizal fungi (EMF) have 

been determined primarily by examining the changes in 

percent root colonization, soil spore density (AMF) or 

sporocarp production (EMF), spore or sporocarp species 

composition, DNA or immunologic assays of species 

composition and structure on root tips or of hyphae in 

soils, and functional analysis of fungal communities. 

Data from naturally occurring, anthropogenic N 

deposition gradients have been particularly useful in 

determining critical loads because of the ability to fi t 

response curves. In deciding if N inputs are aff ecting 

mycorrhizal fungi, we focused mostly on community 

response (species richness and community structure 

changes) and functional responses. However, where 

these are lacking, we considered biomass responses, 

giving them less weight than the other factors.

4.2.2 Acceptable Studies

For information sources, we put more weight on results 

of long-term (>10 yr) fertilizations and gradient studies 

rather than short-term fertilizations, as the latter are 

unlikely to represent a steady-state response to N. 

Gradients have the advantage over typical two-level 

experiments in that response curves and threshold 

responses can be derived. However, even a two-level 

study with only a control and treatment can put an 

upper bound on critical loads if there is a response 

and lower bound if there is no response. Th erefore we 

have included some data from experimental sources, 

recognizing their limitations.

4.2.3 How the Critical Load Value was Chosen 

For our purposes, any detectable deviation in mycorrhizal 

community structure or composition from the baseline 

state is considered to have crossed a critical load 

threshold. For example, if the best-fi t response curve 

in an N deposition gradient indicates a linear response 

starting from the baseline, then any addition above the 

baseline would exceed the critical load. If, in contrast, 

the data suggest a nonlinear response with a threshold 

below which no directional community change can be 

detected, then the critical load is set at this threshold. 

Th e criterion of any detectable directional change is a 

very strict criterion that could be open to debate, but it 

has the merit of being relatively easy to defi ne.

One justifi cation for using the community change 

criterion is that there appear to be functional 

consequences for host plants arising from N-driven 

community change. Th is has been determined by 

growing plants using inoculum that has been subject to 

elevated N. Nitrogen-aff ected inoculum not only has 

lower spore density, but also has lower diversity and 
Chapter 4—Methods GTR-NRS-80 27

shifts in species composition (Egerton-Warburton and 

Allen 2000, Egerton-Warburton et al. 2007, Sigüenza 

et al. 2006). Growth response studies showed that 

native late successional plants had reduced biomass with 

N-aff ected inoculum, while early successional plants 

were unaff ected or sometimes had increased biomass. 

Th is indicates that long-term trends in losses of late seral 

(successional) vegetation and replacement by early seral 

or invasive species were in part promoted by shifts in 

mycorrhizal fungal species composition due to elevated 

N deposition (Corkidi et al. 2002, Sigüenza et al. 2006).

4.2.4 Extrapolation

For ecological regions for which there are no data, we 

have cautiously extrapolated our critical load values 

to regions that have similar biota, climate, and soils. 

For example, we have suggested that ectomycorrhizal 

community response to N in an Engelmann spruce 

forest in the Rockies (Northwestern Forested Mountains 

ecoregion), for which there are no data of which we 

are aware, would be similar to that of communities 

in Alaskan white spruce (Marine West Coast Forest 

ecoregion) and northeastern red spruce-balsam fi r 

(Northern Forest ecoregion) forests, for which we 

do have data. Th us we have used, with qualifi cation, 

the critical loads for the latter two as the basis for a 

provisional estimate for the critical load for the former.

4.3 Lichens and Bryophytes

Lichens and bryophytes are among the most sensitive 

bioindicators of N in terrestrial ecosystems (Blett et 

al. 2003, Bobbink et al. 2003, Fenn et al. 2003, Vitt 

et al. 2003). Consequently, the implementation of 

critical loads protective of the most sensitive lichens 

and bryophytes is likely to prevent undesired impacts 

to the broader terrestrial ecosystem by also protecting 

less N-sensitive vegetation. Protecting the integrity of 

lichen and bryophyte communities, moreover, ensures 

preservation of their contribution to biological diversity 

and their varied ecological roles in nutrient cycling, 

hydrological fl ux, and supply of wildlife forage and 

nesting materials (McCune et al. 2007).

Lichens, especially forest epiphytes, have been 

systematically surveyed for air quality monitoring 

purposes by the U.S. Forest Service Forest Inventory 

Analysis (FIA) program and other U.S. Forest Service 

and National Park Service eff orts, and many of these 

data are available electronically: the NPLichen database 

(Bennett and Wetmore 2005a, 2005b; U.S. Geologic 

Survey 2010), the NPElement database (U.S. Geologic 

Survey 2009), the U.S. Forest Service Forestry Inventory 

and Analysis Program (U.S. Forest Service 2005, 2007), 

and the U.S. Forest Service National Lichens and 

Air Quality Database (U.S. Forest Service n.d.). No 

equivalent eff orts exist for bryophytes in the United 

States. Th erefore, apart from the Tundra and Taiga 

chapters (Chapters 5 and 6) (which extrapolate some 

critical loads from European and Canadian studies) and 

the Desert chapter (Chapter 12) (in which bryophytes 

are integral components of cryptogamic crusts), we 

primarily reference lichen studies.

4.3.1 Terminology

Lichens are adapted to diff erent levels of N availability 

and most species can be sorted into one of three N 

indicator groups: oligotrophs are more or less restricted 

to nutrient-poor environments; mesotrophs have an 

intermediate nutrient requirement; and eutrophs

thrive in nutrient-enhanced environments. A few 

lichens have broad ecological amplitudes and tolerate 

a large range of N availability. As anthropogenic N is 

added to background N levels, the excess N favors the 

small, cosmopolitan eutrophs at the expense of native 

oligotrophic and mesotrophic forage and N2-fi xing 

lichens (McCune and Geiser 2009).

To avoid possible confusion from variable terminology 

in the literature, we consistently used the broad terms 

‘oligotroph’, ‘mesotroph’ and ‘eutroph’ in lieu of 

the more classic lichenological terms, ‘acidophyte’, 

‘neutrophyte’, and ‘nitrophyte or nitrophile’, 

respectively. Th is is necessary because most North 

American studies of N deposition eff ects have not 

considered substrate pH, deposition pH, or N source 

when classifying lichens to N indicator groups. Because 

low-fertility environments are frequently dominated by 

acidic conifer substrates, and moderate- to high-fertility 

environments generally include more hardwoods with 

more neutral bark and more nutrient-enriched canopy 
28 Chapter 4—Methods GTR-NRS-80

drip, the classic and broad terms usually correspond to 

similar lichen lists (Jovan 2008, Geiser et al. 2010)

Common names for lichens follow the Natural 

Resources Conservation Service PLANTS database 

(NRCS 2011). Because many lichen common names 

in PLANTS are not unique, the scientifi c name and 

authority are specifi ed when a common name is used for 

the fi rst time within a chapter. Scientifi c nomenclature 

follows Esslinger (2010).

4.3.2 Response Variables

To identify critical loads for the Marine West 

Coast Forests, Northwest Forested Mountains, and 

Mediterranean California ecoregions, we used two 

epiphytic lichen responses that are closely linked to 

ecological impact: N concentration in lichen thalli, 

and composition of lichen communities. Changes in 

thallus N concentrations and physiology usually precede 

measurable changes in lichen community composition. 

Local N deposition is correlated with N concentrations 

of lichen thalli, which passively accumulate N in excess 

of nutritional requirements (Bruteig 1993, Søchting 

1995). Th is response is typically linear at low N 

deposition levels (Glavich and Geiser 2008). Th e upper 

limit of thallus N concentrations among clean sites 

is typically used as the response threshold. Increasing 

N deposition, likewise, causes changes in lichen 

community composition—such as a decrease in the 

number of oligotrophic species and an increase in the 

number of eutrophic species.

For the Taiga, Tundra, and North American 

Desert ecoregions, we included physiological and 

morphological measures of lichen status, such as changes 

in photosynthetic capacity (e.g., chlorophyll pigment 

ratios, carbon dioxide (CO2

) uptake), membrane 

integrity, amounts of characteristic protective chemicals 

(e.g., usnic acid), apportionment of resources to algal vs. 

fungal partners, ultrastructural integrity, or growth rates. 

4.3.3 Acceptable Studies

As with the mycorrhizal fungi, we primarily relied 

on gradient studies as information sources rather 

than short-term fertilizations. In addition, studies of 

epiphytes are considered superior to studies of grounddwelling species due to epiphytes’ greater independence 

from soil nutrients. For the Rocky Mountain, Marine 

West Coast Forests, and Mediterranean California ecoregions, we were fortunate to obtain community data 

from regional, systematic surveys of epiphytic lichens, 

primarily conducted by the FIA program. Most of the 

physiological studies were N-addition studies. Since 

they were short-term additions or added N to already 

elevated background deposition levels, we feel they do 

not adequately predict long-term eff ects. However, they 

do prescribe upper limits for critical loads.

4.3.4 How the Critical Load Value was Chosen

We determined the critical load by identifying the level 

of N deposition which led to specifi c ecological impacts 

(changes in community composition and increased 

thallus N concentration) using the datasets described 

above. Th e ecological impact at any N deposition level 

was quantifi ed by doing a regression on N deposition 

against proportions of oligotrophic, mesotrophic, and 

eutrophic species, or an index called the “air score” 

(Geiser and Neitlich 2007, Jovan 2008). Th e value of 

N deposition at the ecological threshold is the critical 

load. Th e ecological thresholds were community shift 

toward a higher proportion of eutrophic species and 

elevated thallus N concentration. Th e exact value for 

the threshold thallus N concentration was typically set 

based on the 97.5 percent quantile for clean sites. Th e 

97.5 percent quantile is the thallus N concentration 

below which 97.5 percent of clean site thallus N 

concentrations fall. Th e exact threshold for species 

composition varied because the baseline (species 

composition at an unpolluted site) varied in diff erent 

ecosystems (see Table 4.1 and section 4.3.5). Because all 

the lichen response measures were linear with respect to 

deposition, we favor response thresholds high enough 

to allow for natural variability across clean sites, but 

low enough to protect communities from a shift in 

dominance of indicator groups. We used the 95 percent 

confi dence interval as the range for the critical load.

Forest Inventory and Analysis (FIA) Program lichencommunity based “air scores” are a measure of air 

quality generated from lichen community data and 
Chapter 4—Methods GTR-NRS-80 29

Mean ann. 

precip. (cm)

Biological response 

threshold

Ecoregion

Min Max Air score Minimum % 

oligotrophs

Maximum % 

eutrophs

Rationale/Assumptions

Taiga (US) 20 80 0.02-0.21 41 27 Acidic coniferous substrates, cold 

temperatures favor oligotrophs. Therefore, 

community response threshold set with 

comparatively high oligotroph and low 

eutroph composition.

Northern Forests 

(US)

100 240 0.21 30-41 27-34 Climate and forest type similar to Marine 

West Coast Forest. Therefore, use same 

threshold. Synergistic or confounding effects 

of acid deposition of S and N, which is much 

greater here than Marine West Coast Forest, 

is not considered in this analysis.

Northwestern 

Forested 

Mountains: 

Alaska

30 80 0.02-0.21 41 27 Lower thresholds needed for Alaska because 

of cooler temperatures (see Taiga rationale).

Marine West 

Coast Forests

44 451 0.21 30-41 27-34 Model was developed for this region by 

Geiser et al. (2010)

Northwestern 

Forested 

Mountains: 

non-Alaska

30 203 0.21-0.49 25-40 27-47 Lower and upper thresholds calculated by 

Fenn et al. 2008. Low humidity and high 

temperatures favorable to eutrophs and 

unfavorable to oligotrophs except in high 

elevation fog zones.

Eastern 

Temperate 

Forest: eastern 

hardwoods

71 305 0.33 25-30 34-47 Humid climate favors oligotrophs but 

hardwood dominance favors mesotrophs 

and eutrophs, so response threshold allows 

comparatively more eutrophs. Synergistic or 

confounding effects of acid deposition of S 

and N are not considered in this analysis.

Eastern 

Temperate 

Forest: southeast 

Coastal Plain

102 178 0.21-0.33 30 34 Humid climate and acidic conifer substrates 

favor oligotrophs. Warm summer 

temperatures limiting to oligotrophs? 

Synergistic or confounding effects of acid 

deposition of S and N are not considered in 

this analysis.

Mediterranean 

CA: mixed 

Conifer Forest

41 127 0.33-0.49 25 47 Climate and forest type closest to 

Northwestern Forested Mountains--used 

Fenn et al. 2008 upper threshold to account 

for comparatively warmer, drier conditions 

compared to the Sierras.

Mediterranean 

CA: Chaparral 

and Central 

Valley

17 156 0.33-0.49 25 47 Threshold (50% eutrophs) selected from 

overlay of CMAQ data on Jovan and 

McCune 2006 analysis of this ecoregion. Hot 

dry summers, hardwood dominance favors 

mesotrophs and eutrophs. 

Temperate 

Sierras

30 178 0.49 21-25 47-57 Natural contribution of eutrophs to lichen 

community composition presumed relatively 

high due to dry, hot climate.

Table 4.1—The input variables, assumptions, and rationale behind the calculation of lichen critical loads for U.S. 

ecoregions, derived from the Marine West Coast Forest model: Critical load kg N ha-1 yr-1= (0.0918 + 0.0024 [mean 

annual precipitation] + threshold air score)/0.1493. Precipitation is estimated from PRISM 1961-1990 annual average 

precipitation maps; lichen response thresholds are based on expert opinion (Geiser et al. 2010).
30 Chapter 4—Methods GTR-NRS-80

region-specifi c interpretive models separating air 

pollution eff ects on community composition from 

climate, elevation and other site eff ects (Jovan 2008). 

Models are typically developed from a calibration 

subset of regional FIA lichen community data using 

nonmetric multidimensional scaling and Sorenson’s 

distance measure (McCune and Meff ord 1999) to 

ordinate sites in N-dimensional space. With the aid 

of overlays, regression, scatter plots and other tools, 

at least two strong gradients can usually be extracted 

from the ordination: one relating to air quality 

and a second relating to macroclimate. Air quality 

and climate are then scored as the distance of the 

community (site) along the respective gradient. Repeat 

or additional surveys can be fi tted and scored at any 

time to assess spatial and temporal trends. Finished 

models are available for the southeastern United States 

(McCune et al. 1997), California’s greater Sacramento 

Valley (Jovan and McCune 2005) and Sierra Nevada 

(Jovan and McCune 2006), and western Oregon and 

Washington (Geiser and Neitlich 2007). In the western 

United States, nutrient N deposition is considered the 

dominant air pollutant aff ecting lichen communities 

(Jovan 2008, Geiser et al. 2010); in the southeastern 

United States, N- and S-containing acidifying and 

fertilizing compounds are considered most infl uential 

(McCune et al. 1997).

Lichen community and thallus N concentration appear 

to correlate better with N concentration (mg L-1) in 

deposition as opposed to N load (kg ha-1 yr-1) (Geiser 

et al. 2010). Because lichen thalli accumulate and 

leach N in dynamic equilibrium with concentrations 

of pollutants in precipitation, precipitation can have 

a diluting eff ect on the concentrations of pollutants 

to which lichens are exposed. Th us, the same load of 

N deposited in a higher volume of precipitation will 

have a smaller impact. Th erefore, we accounted for 

precipitation volume when developing lichen critical 

loads based on measurements of bulk (NADP-style) or 

modeled N deposition.

In the Marine West Coast Forests, Northwest Forested 

Mountains, and Mediterranean California ecoregions, 

lichen-based critical loads were derived by correlating 

lichen responses (described above) to Interagency 

Monitoring of Protected Visual Environments 

(IMPROVE2

; fi ne particulate ammonium nitrates 

(NH4NO3) and sulfates ((NH4)2SO4) in air) and 

NADP3

 monitoring networks (wet deposition of 

inorganic N), 10 year total annual average deposition of 

16 N-species modeled by CMAQ4

, or in-situ deposition 

measurements of ammonium (NH4

+

) and nitrate (NO3

-

) 

in bulk and canopy throughfall.

4.3.5 Extrapolation

To make initial estimates of lichen-based critical loads 

for N in forested ecological regions where existing 

analyses were lacking, (i.e., Taiga, Northern Forests, 

Eastern Forests, and the Temperate Sierras), and to 

encompass more of the Northwest Forested Mountains 

ecoregion, we applied a model that was developed 

in western Oregon and Washington for the Marine 

West Coast Forests ecoregion (Geiser et al. 2010). 

Th is model has been shown to replicate previously 

published critical loads for wet oak woods of Scotland, 

2

 IMPROVE (Interagency Monitoring of Protected Visual 

Environments) monitors include an aerosol sampler that 

measures total fi ne particulate (<2.5 μm diameter) NO3

-

and sulfate (SO4

-

) concentrations over 24 hours every third 

day. Th ese anions are presumed balanced by NH4

+

 and best 

correlations between fi ne particulate N and lichen-based air 

scores were observed when mean annual averages included 

N from both sources (IMPROVE 2010; University of 

California, Davis 1995).

3

 NADP (National Atmospheric Deposition Program); weekly 

samples of precipitation captured in buckets designed to open 

during rain and snow events are analyzed for concentrations 

(mg L-1) and total wet deposition (kg ha-1 y-1) of NH4

+

 and 

NO3

-

 and total inorganic N (from NH4

+

 and NO3

-

). Data 

that meet quality control criteria can be downloaded from the 

website (NADP 2009).

4

 CMAQ (Community Multiscale Air Quality model). Th e 

CMAQ model forecasts wet and dry deposition of 16 primary 

and secondary N-containing air pollutants: HNO3

, NH3, 

NO2, PAN, NO, RNO3, PAN2, N2O5, HONO, ANH4I, 

MA-PAN, ANO3I, PBZN, ANO3J, and ANH4J. Model 

output accounts for complex interactions among many 

variables (e.g., landscape, vegetation density, atmospheric 

chemistry, and climate). Mean annual deposition, based 

on 1990-1999 emissions data, was calculated on a 36 km 

grid as described in Porter (2007). For model details, see 

the online CMAQ science documentation (US EPA 2009). 

Lichen response was related to the total sum of wet and dry 

deposition of all 16 air pollutants.
Chapter 4—Methods GTR-NRS-80 31

mesic conifer forests of the California Sierras, and 

dry mixed hardwood-conifer forests of California’s 

Greater Sacramento Valley (Geiser et al. 2010). Input 

variables required are minimum and maximum mean 

annual precipitation and a lichen community response 

threshold (e.g., maximum acceptable percent eutrophs). 

Mean annual minimum and maximum precipitation 

within each ecoregion was estimated from Parameterelevation Regressions on Independent Slopes Model 

(PRISM) national map of 1971-2000 normals (Daly 

et al. 2002, PRISM 2010). Expert judgment was used 

to choose reasonable community response thresholds 

(Table 4.1). Th e thresholds are a minimum allowable 

oligotroph or maximum allowable eutroph percentage 

of the community composition. Input variables and 

response threshold justifi cations are presented in Table 

4.1. Basically, forests with hardwood components 

were assigned a higher response threshold than western 

Oregon and Washington coniferous forests because 

the more alkaline bark and richer canopy exudates of 

hardwoods favor eutrophs. Colder forests were assigned 

lower response thresholds than warmer forests because 

native epiphytes are typically oligotrophs adapted to 

cold- and low-nutrient environments—eutrophs are 

better adapted to nutrient-concentrated, heat and 

drought-stressed environments typical of warmer 

climates. Th e most problematic areas were the Eastern 

Forest ecoregion, for which systematically sampled 

lichen community data exist but are confounded by 

current or historic levels of N and sulfur (S) deposition 

and atmospheric concentrations of sulfur dioxide. 

Because the synergistic eff ects of multiple pollutants on 

epiphytic lichens are not well understood, our critical 

load estimate for eastern forests, based on a western 

model that accounts for nutrient N deposition alone, are 

clearly provisional.

4.4 Herbaceous species and shrubs

4.4.1 Response Variables

Responses of herbaceous species to N deposition 

included changes in species abundances and 

composition, particularly increases in nitrophilic and 

invasive species, increases in productivity, and changes 

in foliar chemistry. In some herbaceous-dominated 

communities, changes in N cycling and inorganic 

N leaching below the rooting zone were also used. 

Typically, the critical loads for these variables were 

about 10 kg ha-1 yr-1 above the value estimated from 

changes in plant composition.

In more impacted ecoregions such as the forested areas 

of the eastern United States using herbaceous species 

for critical loads may be problematic, because, as 

previously stated, N deposition (in combination with 

S deposition) over the last half century has probably 

already signifi cantly altered community composition. 

Th e critical loads obtained using herbaceous species in 

these more impacted regions will therefore potentially 

overestimate the level associated with the initial changes 

in plant diversity.

4.4.2 Acceptable Studies

We included studies across N deposition gradients 

and low level N fertilization experiments to evaluate 

herbaceous species and shrub responses to N deposition. 

We excluded fertilization studies when single input 

rates of more than double the current ambient level 

were used. As noted above, an implicit assumption was 

made that the primary factor infl uencing variation in 

herbaceous species composition in gradient studies was 

N deposition. We considered studies that included 

a range of N experimental inputs to be more reliable 

(##) than those which used single fertilization values or 

gradient studies.

4.4.3 How the Critical Load Value was Chosen

Th e critical load value was estimated as the range 

between the input level at which no response was 

observed and the lowest level at which an alteration in 

species abundance or changes in N cycling occurred. 

In some cases, the input level where a response 

occurred could be quantitatively assessed by fi tting 

the response data to simple mathematical functions 

(Bowman et al. 2006). Quantitative changes in species 

composition for both gradient and N fertilization 

studies are best determined using point-intercept or 

frequency estimates. Subtle changes may be missed or 

misinterpreted using more subjective estimates (e.g., 

visual estimates of cover). An increase in nitrophilic 

species cover was used to determine critical loads 
32 Chapter 4—Methods GTR-NRS-80

in tallgrass prairies, Mediterranean shrublands and 

deserts, and alpine tundra. Changes in community 

structure, using ordination analysis, was also used in 

alpine vegetation. In fi re-prone western ecosystems, the 

increased productivity of invasive grasses has promoted 

increased fi re risk, and the productivity threshold under 

N deposition that promotes fi re has been set as the 

critical load (Rao et al. 2010). Decreases in the diversity 

of understory herbaceous species were used in eastern 

forests, again indicating that the initial increase in 

nitrophilic species has already occurred in these more 

impacted ecosystems.

Gradient studies were used for only a few ecoregions 

(e.g., coastal sage scrub). Critical loads were estimated 

as a range below the deposition rate where signifi cant 

increases in nitrophilic species occurred. In general, 

this approach has not been widely used for herbaceous 

species composition due in large part to the multitude 

of factors, other than N deposition, that infl uence 

herbaceous species composition and chemistry.

Critical loads based on changes in N cycling in an alpine 

ecosystem were estimated as below the N inputs where 

signifi cant changes in net N mineralization, nitrifi cation, 

and inorganic N leaching below the rooting zone were 

observed

4.5 Forests

4.5.1 Response Variables

Th e variables considered for forests include changes 

in percent N in trees and soils, alterations in soil N 

cycling, which lead to elevated NO3

-

 losses, increased 

susceptibility to secondary stressors (freezing, drought, 

pests), declines in tree health, changes in tree growth, 

and increased mortality. We elected to include increases 

in tree growth as a response, because it is an indication 

of alteration by N inputs. In Europe, in contrast, growth 

responses were not considered in setting empirical 

critical loads for forest ecosystems (Bobbink et al. 2003). 

In setting the critical load, we only included declines 

in growth and survivorship and increases in mortality. 

Th ere is considerable variation in N concentration in 

tissue by species and site and over time. Th us, setting a 

threshold value for tissue N concentration as was done 

for lichens is diffi cult. Because repeated surveys at the 

same site are not often available, establishing a value for 

acceptable percentage increase typically is not feasible.

Changes in species composition, such as those reported 

for communities of shorter-lived organisms, are 

infrequently observed for trees. We included some 

studies that reported decreased tree health and growth 

and increased mortality, which is observed only rarely. 

Th erefore, even more for trees than other organisms, 

it is important to understand the mechanisms leading 

to catastrophic changes to be able to identify early 

indicators of change.

4.5.2 Acceptable Studies

We included N gradient, fertilization, and long-term 

studies. Results from long-term, low-level fertilization 

and gradient studies are considered the most reliable. 

Th e additions of N in most forest fertilization studies are 

much greater than ambient deposition. While we note 

these high N fertilization studies in our discussion of 

biomass responses to N deposition, they are typically not 

relevant for calculating the critical load. Gradient studies 

have the advantage of demonstrating changes in real 

time. When a single gradient study was used, the results 

could not be considered more than fairly reliable, #. 

When many gradient studies were included (e.g., Aber 

et al. 2003), the results could be considered reliable, ##.

4.5.3 How the Critical Load Value was Chosen

In some cases (Eastern Temperate and Northern Forest 

ecoregions) with extensive datasets, we were able to 

identify an infl ection point on the response curve, 

the point at which the response begins, and thereby 

identify the critical load with high certainty. We used 

this approach for NO3

-

 losses from forest ecosystems 

which have been shown to increase above a threshold 

N deposition (Aber et al. 2003, Wright et al. 2001). 

In regions with sparser NO3

-

 loss data (Northwestern 

Forested Mountains ecoregion), we were able to 

identify only the range within which the critical load 

is most likely to lie. For gradient studies without a 

clear infl ection point, for example, the tree growth and 

mortality studies in eastern and northern forests, we 

set the critical load to be greater than the low end of 

the range of deposition over which the response was 
Chapter 4—Methods GTR-NRS-80 33

observed. For fertilization studies, we identifi ed the 

critical load as lying in the range between the highest N 

input for which we observed no response and the lowest 

N input for which we observed a response.

4.5.4 Extrapolation

For the Tropical and Subtropical Humid Forests 

ecoregion for which there are no data, we have 

cautiously extrapolated our critical load values from 

regions that have similar biota, climate, and soils. Th us 

we have used, with qualifi cation, the critical loads for 

the Southeastern Coastal Plain forests as the basis for a 

provisional estimate for the critical load for the Tropical 

and Subtropical Humid Forests. We did not extrapolate 

critical loads for forests to other ecoregions.

4.6 Other Ecosystems

Methods for the Great Plains, North American Deserts, 

Tropical and Subtropical Humid Forests, Wetlands and 

Inland surface waters are described in those chapters.

LITERATURE CITED

Aber, J.D.; Goodale, C.L.; Ollinger, S.V.; Smith, M.L.; 

Magill, A.H.; Martin, M.E.; Hallett, R.A.; Stoddard, 

J.L. 2003. Is nitrogen deposition altering the 

nitrogen status of northeastern forests? BioScience. 

53: 375-389.

Bennett, J.P.; Wetmore, C.M. 2005a. Lichens of the 

U.S. National Parks. Th e Bryologist. 108(4): 544-

553.

Bennett, J.P.; Wetmore, C.M. 2005b. NPLichen: A 

database of lichens in the US National Parks. 2005 

Final Report. Available at http://www.nbii.gov/fi les/

nplichen/npl_report.pdf (Accessed May 12, 2010).

Blett, T.; Geiser, L.; Porter, E. 2003. Air pollutionrelated lichen monitoring in national parks, 

forests, and refuges; guidelines for studies intended 

for regulatory and management purposes. Tech. 

Rep. NPS-D2292. Denver, CO: U.S. Department 

of the Interior, National Park Service, Air Resources 

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Earth System Sciences. 5: 299-310.
36
Chapter 5—Tundra GTR-NRS-80 37

5 TUNDRA

K. Nadelhoffer and L.H. Geiser

5.2 Ecosystem Responses to N 

Deposition

A growing body of literature shows that plant growth, 

litter and soil organic matter decomposition, primary 

productivity, ecosystem carbon (C) storage, and plant 

community composition in arctic tundra ecosystems are 

at least partly controlled by nitrogen (N) availability and 

N deposition rates. Th is is because low N availability 

limits both microbial decomposition and plant growth 

in arctic tundra (Robinson and Wookey 1997, Shaver 

et al. 1992). For example, high rates of N addition can 

lead to large losses of organic C from tundra peat as 

well as to increases in vascular plant growth (Mack et 

al. 2004). Experimental N additions typically increase 

plant growth in tundra and responses are moderated 

by environmental factors. Graminoids and shrubs 

become dominant in dry heath tundra in well drained 

uplands, shrubs become dominant in moist tundra on 

gentle slopes, graminoid biomass increases in low-lying 

wet tundra, and bryophyte (mosses, hornworts, and 

liverworts) and lichen biomass increases with increasing 

moisture availability (Shaver et al. 1998, van Wijk et 

al. 2004). Even though N deposition may increase 

plant growth and net primary production, increases in 

peat and soil organic matter decomposition resulting 

from N deposition could be large enough to off set C 

accumulation by vegetation, resulting in net C losses 

from tundra ecosystems (Nowinski et al. 2008).

Bryophytes and lichens (referred to collectively as 

cryptogams) are of particular concern in the arctic 

because they regulate tundra water balances and energy 

balances. Cryptogams, especially sphagnum mosses, 

can dominate the top layer of tundra; the accumulated 

necrotic biomass can be a major component of peat, 

which is highly hygroscopic. In addition, acids and 

phenolic compounds in cryptogams inhibit decay. Th e 

growth and abundances of bryophytes and lichens can 

be diminished by the direct eff ects of moderate to high 

N inputs. In addition, increased growth of competing 

graminoid and shrub species in response to N inputs can 

5.1 Ecoregion Description

Th e North American Arctic, comprising the Tundra and 

Arctic Cordillera ecoregions (CEC 1997, Chapter 2), 

covers more than 3 million km2

 (300 million ha), and 

accounts for nearly 14 percent of the North American 

land mass. Th e North American Arctic also constitutes 

about 20 percent of the much larger circumpolar Arctic 

shared by Canada, the United States, and six European 

and Asian countries. Th e ecoregion description is 

adapted from CEC (1997). Th e Tundra ecoregion 

(CEC 1997, Figure 2.2) is a mosaic of alpine meadows, 

foothills, mesas, low-lying coastal plains, river corridors, 

and deltas encompassing northern and western Alaska, 

the arctic islands of Canada, northern portions of the 

Yukon and Northwest Territories, and far northern 

Quebec. Coastal plains are typically wet with high 

densities of shallow lakes which are covered by ice for 

9 months or more. Uplands have fewer, deeper lakes, 

most of which are connected to river networks draining 

into the Arctic Ocean or Bering Sea. A key feature is 

that soils are frozen in spatially contiguous permafrost 

and often contain large stores of organic carbon (or 

“peat”). Soil surface layers thaw in summers, usually to 

a depth of <1 m. Th e climate is characterized by long, 

dark, cold winters and short, cool summers with up to 

several months of daylight. Annual precipitation varies 

from 10 to 50 cm; snow may fall any month and snow 

pack usually persists for at least 9 months. Snowfall 

and its redistribution by winds both infl uence and are 

partly determined by plant community structure. Th e 

vegetation is characterized by combinations of sedges, 

grasses, shrubs, bryophytes, lichens, and forbs. Shrubs 

include dwarf birch (Betula nana), willows (Salix spp.), 

and heath species. Moving from south to north, shrub 

densities typically decrease, and sedges, mosses, and 

lichens increasingly dominate. River valleys are often 

lined with tall birch or willow shrubs, and occasionally 

with small balsam poplar (Populus balsamifera) or 

stunted white and black spruce (Picea glauca and P. 

mariana) trees. Wetlands are common in low-lying areas 

and support sedges and bryophytes.
38 Chapter 5—Tundra GTR-NRS-80

result in shading of low-lying bryophytes and lichens, 

thereby decreasing cryptogam biomass indirectly, even 

when cryptogam growth is N limited (Berendse et al. 

2001, Bubier et al. 2007, van der Wal et al. 2005). 

Shifts in cryptogam cover, biomass, morphology, and 

chemistry; changes in the composition of vegetative 

communities; and shifts in vegetation growth forms to 

greater dominance by vascular plants (graminoids and 

shrubs in particular) are indicators of tundra ecosystem 

responses to N deposition, as well as to other climate 

change drivers. Also, changes in cryptogam communities 

and bryophyte-lichen layer thickness, whether derived 

from N toxicity, shading from vascular plants, or 

other factors, can alter tundra soil moisture balances, 

temperature, and nutrient cycling rates (Gornall et al. 

2007).

Aquatic ecosystems are often “fi rst responders” to 

changes in environmental conditions, as their structure 

and functionality can be altered by changes in exports 

of nutrients and organic materials from terrestrial 

ecosystems within drainage basins. Th is sensitivity is 

particularly true in arctic tundra, where shallow depth 

of soil thaw and short residence times of soil solutions 

can lead to rapid changes in drainage water chemistry 

in response to biogeochemical changes in terrestrial 

ecosystems (Hobbie et al. 1999a, Keller et al. 2007), 

such as could result from elevated N deposition. 

Arctic streams are sensitive to N inputs, whether from 

increased N deposition or from increased decomposition 

in drainage basins due to warming and deeper soil 

thaw (Benstead et al. 2005). Evidence is emerging from 

Canada that diatom communities, nutrient cycling, 

and sediment formation in arctic lakes are being altered 

by elevated N deposition (Wolfe et al. 2006). Such 

studies indicate that N deposition may be altering the 

functioning of terrestrial as well as aquatic ecosystems 

in arctic tundra regions over broad scales, even when 

the eff ects of N deposition on individual terrestrial 

communities and ecosystems are undetectable.

5.3 Range of Responses Observed

Similarities among Northern Hemisphere arctic 

ecosystems in climate, soils, marine infl uences, 

hydrology, and vegetation justify the application of 

Eurasian research to assessments of the North American 

tundra response to N deposition. Th e defi nition of 

tundra employed by Bobbink et al. (2003) and the 

UNECE (2007) for Europe is based on the European 

University Information Systems (EUNIS) system of 

classifi cation (Davies et al. 2004). ‘Tundra’ ecosystems 

share similarities with ‘arctic, alpine and subalpine scrub 

habitats’ ecosystems in that the growing season is short, 

temperatures are low, winds are frequent and strong, 

and the distribution of plant communities depends 

on the distribution of snow during winter and spring. 

Both ecosystem types are infl uenced by frost activity or 

solifl uction, slow rates of organic matter decomposition 

and nutrient cycling, and limitation of primary 

production due to low nutrient availability (Robinson 

and Wookey 1997). As with North American tundra, 

the key feature that distinguishes tundra from other 

ecological regions in both the CEC and EUNIS systems 

is the presence of permafrost, which restricts deep root 

penetration and often keeps the overlying thawed soil 

layer waterlogged in summer. Responses to N deposition 

are summarized in Table 5.1.

5.3.1 Community Level Responses of Tundra 

Vegetation

Long-term N addition experiments typically add N at 

high rates, e.g., 10 to100 kg ha-1 yr-1. Th ese rates are one 

to two orders of magnitude greater than background 

N deposition rates and well in excess of tundra plant 

N uptake requirements. Such experiments provide 

insights into mechanisms of vascular plant responses 

to atmospheric N deposition but are of limited use in 

quantifying ecosystem response levels. Growth rates of 

vascular plant species typically increase in response to 

experimental N additions in dry, moist, and wet tundra 

ecosystems in both low and high latitude arctic regions. 

Because vascular plant species diff er in sensitivity to 

N deposition and N availability, plant community 

composition typically shifts after several years or decades 

of experimental N additions. For example, primary 

productivity increased dramatically in dry heath tundra 

near Toolik Lake, Alaska, after 8 years of fertilization at 

100 kg N ha-1 yr-1 (Gough et al. 2002). Species richness 

decreased in this study as dwarf birch, a low-lying shrub, 

became dominant and various graminoids, shrubs, forbs, 

and cryptogams declined. Species shifts and primary 
Chapter 5—Tundra GTR-NRS-80 39

productivity may increase with N additions in high 

arctic systems as well. For example, Madan et al. (2007) 

reported increases in growth and biomass of shrubs 

(willows and purple mountain saxifrage (Saxifraga 

oppositofolia)) as well as increases in cryptogam cover in 

Greenland polar deserts (ambient deposition 1 kg N ha-1

yr-1) in response to relatively low levels (5 kg N ha-1 yr-1 ) 

of experimental N addition.

A recent 3-year study in northwest Greenland (Arens 

et al. 2008) suggests that N addition rates as low as 

1 to 5 kg ha-1 yr-1 (ambient deposition of 0.1 to 5 kg 

N ha-1 yr-1) can dramatically increase photosynthesis 

and growth of grass species, and signifi cantly alter 

the structure and function of high arctic systems, 

particularly with increases in phosphorus (P) availability. 

Th e target high arctic ecosystem encompassed dwarf 

shrubs, graminoids, forbs, bryophytes, and lichens. 

Measuring carbon dioxide (CO2) fl uxes in a portable 

fi eld chamber, the authors demonstrated a near 

doubling of gross ecosystem photosynthesis at plots with 

0.5 and 1 kg added N ha-1 yr-1 after both 1 and 2 years 

of treatment. Th e photosynthesis response saturated 

between 1 and 5 kg added N ha-1 yr-1, presumably 

because P became limiting. Total plant cover followed 

the same pattern, increasing up to the addition of 1.0 kg 

N ha-1 yr-1, then decreasing or not changing through 5 

kg added N ha-1 yr-1; grasses invaded much of the bare 

ground. Cover of deciduous shrubs increased 41 percent 

in the 0.5 kg N ha-1 yr-1 treatment, graminoids cover 

increased 97 percent with the 1 kg N ha-1 yr-1 treatment, 

and forb cover increased 296 percent and 740 percent 

with 1 and 5 kg N ha-1 yr-1 treatments. Bryophyte and 

lichen cover appeared to increase to 1 kg N ha-1 yr-1 and 

then decrease, but signifi cance was not noted. Th e N+P 

treatment (5 kg N ha-1 yr-1 plus 2.5 kg P ha-1 yr-1)— 

produced the greatest responses, including signifi cant 

increases in cover of alpine fescue (Festuca brachyphylla) 

grass (from 1.3 percent to 43.8 percent), arctic willow 

(Salix arctica) (4.5 percent to 13.8 percent), bryophytes 

(3 percent to 10.7 percent), alpine chickweed (Cerastium 

alpinum) (0.2 percent to 4.8 percent), the establishment 

of arctic bluegrass (Poa arctica) (from 0 percent to 6.5 

percent), signifi cant decreases in total lichen cover 

(from 8.2 percent to 2.3 percent), and the eradication 

of purple mountain saxifrage (from 0.8 to 0 percent). 

Among the three dominant plants (curly sedge [Carex 

rupestris], entireleaf mountain-avens [Dryas integrifolia], 

and arctic willow), foliar N increased linearly through 

5 kg ha-1 yr-1 of added N. Th ese results provide 

striking evidence of ecologically important increases in 

ecosystem productivity, vegetative cover, and altered 

community composition after just 1 to 2 years of very 

low N additions.

A meta-analysis by van Wijk et al. (2004) showed 

that vascular plant growth, primary productivity, and 

biomass of wet and moist tundra ecosystems in midarctic regions of Alaska and Scandinavia increased 

dramatically in experimental nutrient addition plots (N 

alone and in combination with P and other nutrients). 

However, increases in shrub growth in moist tundra 

were greater in Alaska than in Europe and were greater 

in moist than in wet tundra. Shrub and graminoid 

growth increased in both regions, but at the expense of 

forbs and cryptogams. Also, recent work in mid-arctic 

Alaska suggests that N deposition on non-acidic tundra 

(with base-rich soils), unlike the more extensive acidic 

tundra, may not lead to shrub dominance (Gough and 

Hobbie 2003). In non-acidic tundra types, enhanced 

graminoid growth will likely diminish the importance of 

forbs and cryptogam species.

Bryophytes, particularly mosses, can have positive 

growth responses to atmospheric N deposition because 

they lack roots and assimilate N and other nutrients 

almost exclusively via foliar uptake (Salemaa et al. 

2008, Solga and Frahm 2006). Th us, where graminoid 

and shrub species do not overtop mosses and other 

bryophytes, N deposition could increase plant growth, 

primary productivity, and moss mat thickness. 

However, N deposition can also increase tissue turnover 

and mortality in some arctic mosses (Koranda et al. 

2007), and thereby diminish moss cover in some tundra 

ecosystems. Also, competition by more erect vascular 

plants can lead to competitive suppression of moss 

growth and decreases in moss cover even when moss 

growth is N limited (Klanderud 2008). Positive growth 

responses of vascular plants to increased N availability 

can lead to declines in lichen biomass in low- and midarctic regions (van Wijk et al. 2004), but probably not 

in the high arctic (Cornelissen et al. 2001, Madan et 
40 Chapter 5—Tundra GTR-NRS-80

al. 2007), where plant cover is typically less than 100 

percent of soil surface area.

A seminal, long-term experiment in the high arctic of 

the Svalbard Islands, Norway, by Gordon et al. (2001) 

has been used to identify plant responses observed at 

diff erent N input levels. Th e experiment involved three 

common tundra communities dominated by dwarf 

shrubs (mountain-avens, mountain heather (Cassiope), 

and willow species) and also characterized by lichens and 

bryophytes. Background deposition was very low (<1 kg 

N ha-1 yr-1) with N treatments of 0, 10 and 50 kg ha-1

yr-1 as ammonium nitrate (NH4NO3) solution and 0 

and 5 kg P ha-1 yr-1 as potassium dihydrogen phosphate 

(KH2

PO4) in fi ve applications per growing season over 8 

years. Th ey reported the following: 

• Responses were rapid and persistent. Both 

N and P concentrations in a dicranum 

moss (Dicranum scoparium) and the juniper 

polytrichum moss (Polytrichum juniperinum)

increased with each level of nutrient application. 

Nitrate (NO3

-

) reductase activity decreased in 

juniper polytrichum at the low N treatment, 

suggesting N saturation and the potential for 

nitrate to pass through the bryophyte layer to 

become available to microbes and higher plants. 

Composition of the vegetation communities 

shifted signifi cantly with a notable decrease in 

lichen cover. Bryophyte productivity increased 

and persisted 5 years after N additions ceased, 

suggesting that N (mainly NH4

+

) was strictly 

conserved in the bryophyte layer and that any 

reversal of the eff ect would be slow.

• Th e P treatment provided unequivocal evidence 

of partial limitation by P in this ecosystem, 

with greater response to N when P was applied. 

Th erefore, in tundra systems, critical loads are 

expected to be lower in ecosystems that are not 

P-limited. Th is is especially notable because 

global warming is expected to increase both N 

and P availability (currently wet tundra sites are 

more likely to be P-limited than moist sites with 

a thinner peat layer, while dry tundra sites are 

primarily N-limited [Shaver et al. 1998]).

A circumpolar comparison of eff ects of N deposition 

and temperature-warming treatments on lichen cover 

by Cornelissen et al. (2001) included three studies 

from North America: Chapin et al. (1995), Hik (as 

reported in Cornelissen et al. 2001), and Hobbie et al. 

(1999b), and 11 studies from Sweden and Norway. 

Th e Cornelissen meta-analysis demonstrated consistent 

marked decreases in lichen cover with increased N 

availability and increasing temperature; all N application 

rates were relatively high, between 40 and 100 kg ha-1

yr-1 for a minimum of 3 years. In the low arctic and 

subarctic, this response was attributed to increased 

shading of lichens by shrubs, which generally reached a 

taller stature and were more responsive to N fertilization 

at lower latitudes compared to the high arctic. Although 

lichens can achieve a net positive photosynthesis rate 

very rapidly after hydration, they are often light-limited 

in the Arctic due to low incident radiation (Hahn et al. 

1993). Shading, whether from shrubs or fog and cloud 

cover, signifi cantly reduces lichen growth rates. Lichen 

cover responses in the high arctic were more ambiguous, 

with some studies showing slight increases and others 

small decreases. Cornelissen (2001) concluded that 

tundra lichens in climatically milder arctic ecosystems 

may decline with global climate change because higher 

temperatures are associated with greater availability of N 

and P, all of which favor vascular plants.

In addition to the above three North American studies, 

a study was conducted by Weiss et al. (2005) at the 

Toolik Lake Tundra LTER site (Alaska), in the northern 

foothills of the Brooks Range. Th ey documented an 

essential absence of lichens after 10+ years of fertilization 

at 100 kg N ha-1 yr-1. Phosphorus fertilization at 50 and 

100 kg ha-1 yr-1 improved N2

 fi xation rates of the felt 

lichens, Peltigera apthosa and P. polydactylon, but the 

eff ect was unimportant compared to the reduction in 

lichen cover.

Since the Bobbink et al. (2003) assessment, a number 

of studies have provided further insight regarding the 

eff ects of N and P availability on tundra vegetation. 

During a 10-year experiment, Fremstad et al. (2005) 

tested species cover in low alpine and middle alpine 

vegetation communities in the Dover Mountains of 

south-central Norway. Nitrogen was added at 7, 35, and 
Chapter 5—Tundra GTR-NRS-80 41

70 kg ha-1 yr-1; background N deposition was 2 to 4 kg 

ha-1 yr-1. Th e most sensitive species at the warmer, lowalpine site were a witch’s hair lichen (Alectoria nigricans) 

and a cetraria lichen (Cetraria ericetorum). Th ese species 

exhibited signifi cant decreases in cover at the lowest 

N application rate of 7 kg ha-1 yr-1. Cover of other 

witch’s hair (Alectoria ochroleuca), cetraria (Flavocetraria 

cucullata, F. nivalis), and reindeer (Cladonia mitis, 

C. rangiferina, C. stellaris) lichens decreased at the 

application rate of 35 kg N ha-1 yr-1. Th e only lichen 

unaff ected by any level of N application was the N2-

fi xing snow lichen, Stereocaulon paschale. In contrast, 

fertilization had no signifi cant eff ect on the bryophytes 

or vascular plants (dwarf shrubs and graminoids) except 

for sheep fescue (Festuca ovina), a grass, which increased 

slightly. At the colder, mid-alpine site, a cetraria lichen 

(Cetrariella delisei), was the only species in which cover 

changed over 10 years due to N application, and then 

only at the addition rate of 70 kg ha-1 yr-1, implying a 

slower response time in colder climates.

5.3.2 Physiological and Chemical Responses of 

Vegetation

Th e adaptation of arctic vegetation, especially lichens 

and bryophytes, to N-limited ecosystems through 

extreme conservation has been well illustrated recently 

by two studies employing 15N as a tracer. Tye et al. 

(2005) conducted one-time additions of 1 or 5 kg 

15N ha-1, using separate applications of 15N-labeled 

sodium nitrate (NaNO3) and ammonium chloride 

(NH4Cl) at both levels of N addition, in addition to 

the background deposition rate of 1 kg N ha-1 yr-1 at 

two sites in the high arctic tundra of Svalbard, Norway. 

Th ey observed a highly conservative retention by soil, 

litter, and vegetation of ~60 percent of both N ions 

of the initial application across all 3 years. On a unit 

weight basis, the lichens and graminoids were especially 

good scavengers of 15N and exhibited a high capacity 

to immobilize additional N released compared to other 

sinks (microbes, vascular plants, mosses, willows, litter, 

humus, mineral soils), although the bulk of 15N was 

stored in humic and mineral fractions of the soil and in 

bryophytes. Kytöviita and Crittenden (2007), in a study 

of lichen mats dominated by reindeer and snow lichens 

(Cladonia and Stereocaulon) in a dry boreal forest of 

Finland, demonstrated that a signifi cant portion of 15N 

fed to the lower parts of lichens is translocated towards 

growing apices. Th ey concluded that internal recycling 

of N improves N use effi ciency in lichens and is likely 

ecologically important in N-limited environments 

typical of the undisturbed arctic. Th ese and other studies 

have repeatedly shown a direct correlation between 

N application or N availability and concentrations of 

N in lichen thalli (Hahn et al. 1993; Hyvärinen and 

Crittenden 1998a, 1998b; Hyvarinen et al. 2003; 

Walker et al. 2003; Weiss et al. 2005).

In addition to decreases in lichen cover with increasing 

N availability, mostly at relatively high application rates, 

new evidence points to clear but more subtle eff ects 

of low rates of N application (3 to 10 kg ha-1 yr-1) on 

lichen physiology. Makkonen et al. (2007) showed 

the vulnerability of normally N- and P-limited lichens 

to nutrient enrichment. Lichen cushions of the star 

reindeer lichen (Cladonia stellaris) were treated with 

3 and 10 kg N ha-1 yr-1 and 10 kg P ha-1 yr-1, alone 

and in combination, over 14 weeks. Nitrogen, P, and 

N:P ratios infl uenced the proportion of the medullary 

layer volume occupied by the algal cells, the thallus 

volume occupied by the internal lumen, and the algal 

cell-wall area covered by fungal hyphae, indicating 

that ecologically realistic changes in the availability of 

key macronutrients can alter the growth of symbionts. 

Hyvärinen et al. (2003) showed that a moderate N 

treatment (3 kg ha-1 against a background of 0.5 kg 

ha-1 wet deposition in 3 months) lowered usnic acid 

(C18H16O7) concentrations in the star reindeer lichen 

with and without P application; P alone increased usnic 

acid, and higher N (10 kg ha-1) had no eff ect on usnic 

acid. None of the treatments increased growth rate. 

Usnic acid in lichens has well established anti-microbial 

and UV- protection properties (Rikkinen 1995) 

relating to lichen fi tness. By inhibiting catalysis of urea 

to carbon dioxide (CO2

) and ammonia (NH3), usnic 

acid likely moderates metabolic imbalances in lichens 

created by excess C and insuffi cient N (Rikkinen 1995).

Sundberg et al. (2001), in an elegant study of lichen 

transplants, demonstrated the highly regulated, 

species-specifi c balance of C and N for coordinated 

development of lichen thalli. Nitrogen availability 

to transplants of the arctic kidney lichen (Nephroma 
42 Chapter 5—Tundra GTR-NRS-80

arcticum) and a felt lichen (Peltigera aphthosa) was 

manipulated during 4 summer months by removing 

cyanobacterial N2-fi xing structures to cause N starvation, 

or by weekly irrigation with NH4NO3 at a rate of 9 kg 

N ha-1 yr-1 plus background deposition of 5 kg ha-1 yr-1. 

N-starved lichens continued to gain weight, but did 

not increase in cover, demonstrating that C-fi xation 

can proceed under N-limited conditions, although N 

is required for hyphal growth. Cover of lichens with 

added N increased more slowly than nonfertilized 

controls. For all treatments, weight gain was dependent 

on the proportion of N invested in algal photosynthetic 

tissue and there was a positive correlation between light 

use effi ciency and thallus N. Intra-specifi c ratios of 

chlorophyll a:chitin (an N-containing polysaccharide in 

cell walls) and chlorophyll a:ergosterol (an N-containing 

component of cell membranes in fungi and algae) were 

relatively constant across a two- to threefold variation in 

thallus N, suggesting that lichens regulate N distribution 

within the thallus.

5.4 Critical Loads Estimate

We have few experiments assessing the eff ects of N 

additions at our proposed critical load for the North 

American Tundra ecoregion. Th ere is only one 

published study specifi c to the North American Tundra 

ecoregion (Arens et al. 2008) that has used ecologically 

realistic N input rates to suggest vascular plant, 

bryophyte, or lichen-based critical loads (see Table 5.1). 

Critical loads are summarized in Table 5.2.

In 2003, the empirically derived critical load for tundra 

ecosystems in Europe proposed by Bobbink et al. (2003) 

and adopted by the UNECE (2007) was 5 to 10 kg N 

ha-1 yr-1. Th is range was considered reliable, even though 

it is based primarily on one long-term experiment in 

the high arctic of the Svalbard Islands, Norway, by 

Gordon et al. (2001). Th is study is compelling because 

it included the critical load range, involved controlled 

additions of N, and looked at responses over an 

8-year period (Table 5.1). Data from Fremstad et al. 

(2005) and Madan et al. (2007), which demonstrated 

vegetation community responses at 9 to 11 and 6 

kg N ha-1 yr-1, respectively, corroborate the research 

by Gordon et al. (2001) and point to the important 

mediating role of other environmental and genetic 

factors in cryptogam and vascular plant responses to N 

deposition. In particular, harsher climates and intrinsic 

properties of individual species may slow the rate of, or 

override, responses.

European analyses suggest critical loads not exceed 

5 to 10 kg N ha-1 yr-1, but the long-term eff ects of 

deposition rates lower than 5 kg N ha-1 yr-1 on tundra 

ecosystems have not yet been adequately assessed. 

Recent work, however, suggests arctic systems may 

respond to rates of N deposition between 1 and 5 kg 

ha-1 yr-1 with altered gross photosynthetic productivity, 

structure, and function of plant communities (Arens et 

al. 2008). Background N deposition rates in European 

tundra are higher than N deposition in relatively less 

polluted regions of the North American Arctic. For 

example, precipitation chemistry measurements at the 

Toolik Lake Arctic Tundra LTER site in northern 

Alaska (Hobara et al. 2006, Shaver5

) and at NADP 

sites elsewhere in Alaska (See Chapter 6, and data from 

Ambler, Alaska, at http://nadp.sws.uiuc.edu/ads/2004/

AK99.pdf) show that inorganic N inputs from the 

atmosphere are <0.5 kg N ha-1 yr-1.

Given the multiple lines of evidence from Arens et al. 

(2008) that deposition as low as 1 kg N ha-1 yr-1 can 

impact tundra ecosystems, setting critical loads at 10 

to 20 times current N deposition rates measured at 

relatively pristine tundra sites would be unwise. Aquatic 

systems in arctic drainages (Wolfe et al. 2006) and 

indicator species among the shrubs, grasses, lichens, 

and mosses appear to be responding to N deposition at 

low levels of ambient atmospheric or experimental N 

inputs. Cellular ultrastructure and pigment production 

are sensitive to deposition loading of ≤3 kg N ha-1

yr-1 (Hyvärinen et al. 2003, Makkonen et al. 2007). 

Moreover, vegetation in tundra ecosystems is highly 

effi cient with respect to N cycling (Chapin et al. 1980), 

with plant N uptake from soil ranging between 1 to 6 kg 

ha-1 yr-1, or 20 to 25 percent of the 5 to 20 kg N used to 

supply annual primary productivity (Shaver et al. 1991, 

5

Shaver, G.R. Unpublished data. Senior scientist of Marine 

Biological Laboratory, Ecosystems Center, Woods Hole, MA 

02543.
Chapter 5—Tundra GTR-NRS-80 43

Table 5.1—Responses to N deposition for the Tundra ecoregion. Reliability rating: ## reliable; # fairly reliable; (#) expert judgmentSite N input

kg ha-1 yr-1

Reliability Response Comments Study

NW Greenland 1-5 ## Increased net CO2 uptake by 

combined vegetation saturating at 1.5 kg N ha-1 yr-1; changes in shrub, 

graminoid, forb, bryophyte, and lichen diversity and cover; increased grass cover, increased foliar N

Response is greater if P availability increases

Arens et al. 2008

Hailuota Island, northern Finland, 65o N

<3 (#) Decreased usnic acid concentrations 

in lichens at 3 kg; no effect at 10 kg (concave response)

3 kg ha-1 on background of 0.5 kg ha-1

for 3 months. P increased usnic acid

Hyvärinen et al. 2003

Hailuota Island, northern Finland, 65o N

3 to 10 (#) Increase in lichen N concentrations, 

reduction in medullary volume, increase in internal lumen volume

14 week addition of 3 and 10 Kg N (total) and 10 kg P to lichen cushion transplants; P moderated N response

Makkonen et al. 2007

Greenland polar deserts

6 ## Increases in shrub and cryptogam 

growth and biomass

N addition of 5 kg ha-1 yr-1, plus 1 kg 

ha-1 yr-1 deposition

Madan et al. 2007

Dover Mtns, southcentral Norway 60 o N <9-11 ## Decreased lichen cover and increased grass cover at warmer but 

not colder site

10 yr experimental addition of N at 7, 35 and 70 kg ha-1 yr-1 to low and mid alpine 

vegetation, plus 2 to 4 kg N ha-1 yr-1

deposition

Fremstad et al. 2005

Svalbard Islands, Norway 78 o N

<10 ## Increased N and P concentrations 

in cryptogams, reduced nitrate reductase activity in bryophytes, N saturation of cryptogam layer, decrease in lichen cover, and increase in bryophyte cover 

8 yr addition of 0, 10 and 50 kg N ha-1

yr-1; 0 and 5 kg P ha-1 yr-1 to three 

high arctic vegetation types. Adding P exacerbated adverse N effects. Wet tundra (P limited) may have higher critical loads than dry tundra (N limited)

Gordon et al. 2001

Västerbotten, Sweden 66.2o N (collected); 

Ulterviken, Sweden 63.5o N (transplanted)

<14 # Increased lichen N concentrations; 

slightly slowed growth in cover area; increase in chlorophyll a, ergosterol, and chitin

Comparison of N starved (cephalodia removed), control and 10 kg ha-1 yr-1

N to N2-fi xing lichens. After 4 months 

fertilization appeared to mostly benefi t lichens, which were N-limited

Sundberg et al. 2001

Circumpolar arctic <40 # Increased growth of shrubs, 

decrease in lichen cover (attributed to shading by shrubs)

Meta-analysis of 14 arctic studies. Responses diminish with latitude due to colder temps and poorer N and P mobilization

Cornelissen et al. 2001

Toolik Lake tundra LTER, Brooks Range, Alaska 66o N

<50 ## Decreased cover and species 

diversity of lichens

Lichens disappeared after 10 years fertilization at 100 kg ha-1 yr-1

Weiss et al. 2005
44 Chapter 5—Tundra GTR-NRS-80

Shaver and Chapin 1991). Nitrogen deposition rates 

equal to or in excess of annual plant N uptake would 

likely alter competitive interactions among plant species, 

dominant growth forms, biogeochemical exports, and 

energy balances of tundra ecosystems. Given the high 

degree of N limitation of plant growth in the North 

American Tundra ecoregion, low background levels of 

N deposition in historic and geological time, low levels 

of N cycling between tundra plants and soils, and the 

apparent sensitivity of freshwater ecosystems in the 

Arctic to changes in landscape biogeochemistry, the 

North American Tundra ecoregion critical load for N 

deposition should considerably lower than the European 

tundra critical load. We recommend that the critical 

load for the North American Tundra ecoregion be 1 to 3 

kg N ha-1 yr-1; the lower value is two times estimates for 

inorganic N deposition in Alaskan tundra monitoring 

sites. We consider this critical load to be reliable, but 

note that it is primarily supported by work in a single 

high arctic location in North America (Arens 2008).

5.5 Future Research Directions and 

Gaps in Data

A more extensive network of atmospheric deposition 

monitoring stations is clearly needed in the North 

American Tundra ecoregion. Th ese monitoring stations 

should focus on acquiring high quality, year-round 

measurements of inorganic, particulate, and organic 

N forms as well as other biologically active elements 

including S, mercury (Hg), calcium (Ca), magnesium 

(Mg), and potassium (K). Further research is needed, 

specifi cally in North American tundra locations 

outside of Greenland, in both high and low arctic 

environments. Th is research should compare ecologically 

realistic applied, measured, or modeled N deposition 

with biogeochemical responses to N deposition, with 

physiological/morphological/distribution data for 

various indicator vascular plant, bryophyte and lichen 

species, and with foliar or thallus N concentrations. 

Until then, critical loads based on responses from 

Nordic countries will remain our best estimates.

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48
Chapter 6—Taiga GTR-NRS-80 49

6 TAIGA

L.H. Geiser and K. Nadelhoffer

6.2 Ecosystem Responses

Responses to increased nitrogen (N) deposition in 

boreal regions (which include Taiga and Northern 

Forests ecoregions) include increased productivity, 

foliar N concentration, and N leaching from soils, 

plant community changes (including vascular plants, 

bryophytes, lichens, and algae), and physiological changes. 

In boreal ecosystems and other largely oligotrophic 

environments, N is growth-limiting, and most plant 

species are adapted to low available N. As N availability 

increases, faster growing, but less N-use effi cient species 

(more N taken up per unit growth) typically outcompete slower growing species (Aerts and Chapin 

2000). Increasing N deposition alters plant community 

structure, often leading to a short-term “positive” response 

of increased productivity and vigor, followed by long-term 

changes in species composition and richness (Gough et 

al. 2000). Such community-level responses occur at low 

deposition rates and have been used to mark the low 

end of ecosystem N response (Bobbink et al. 2003, de 

Vries et al. 2007). In boreal forests, overall ground fl ora 

species number may not be aff ected by N enrichment, 

despite a drastic change in species composition, due to 

reciprocal increases in nitrophilous species with declines 

in typical species (Bobbink 2004). However, in peatland 

bogs, where high water levels and low pH prevent 

nitrophyte invasions (Allen 2004), decline in species 

richness is a more characteristic response.

Increased N supply is often accompanied by increased 

foliar N concentration, which in turn may increase 

susceptibility of vegetation to frost or diseases. Th is eff ect 

is documented for boreal forest trees (e.g., Aronsson 

1980, Balsberg-Påhlsson 1992, Kallio et al. 1985, 

Schaberg et al. 2002), for their common understory 

ericaceous species (e.g., Strengbom et al. 2002, 2003), 

and for bogs (e.g., Wiedermann et al. 2007). At higher 

deposition levels, as N is no longer completely taken up 

by vegetation or immobilized in soils, N saturation and 

leaching may occur (Lamontagne 1998; Lamontagne 

and Schiff 1999, 2000; Tamm et al. 1999).

6.1 Ecoregion Description

Th e Taiga ecoregion (CEC 1997; Figure 2.1) includes 

most of interior Alaska and much of Canada’s northern 

boreal forest. Th e ecoregion description is adapted 

from CEC (1997). Th e Alaskan portion is underlain by 

horizontal limestone, shale, and sandstones, creating 

a fl at to gently rolling plain covered with organic 

deposits, hummocky moraines, and lacustrine deposits; 

the eastern portion is underlain by the Canadian 

Shield. Lowlands are mostly peatlands and permafrost 

is widespread. Nutrient-poor soils dominate southern 

portions and permafrost soils occur in the north. Th e 

climate is subarctic: the short summers have long 

daylight and cool temperatures; winters are long and 

very cold. Snow and freshwater ice persist for 6 to 8 

months annually. Mean annual temperatures are -10 

to 0 o

C; mean annual precipitation is 200 to 500 mm. 

Innumerable lakes, bogs, other wetlands, and forests 

are interspersed with tundra-like shrublands and sedge 

meadows. In the north, forests transition to woodlands 

with a lichen groundcover and fi nally merge into 

tundra at the climatic limits of tree growth. Mid-zone, 

dwarf birch (Betula nana), Labrador tea (Ledum spp.), 

willow (Salix spp.), bearberry (Arctostaphylos alpina), 

mosses, and sedges dominate. In the south, open stands 

of stunted black spruce (Picea mariana) and jack pine 

(Pinus banksiana) are accompanied by alder (Alnus 

spp.), willow, and tamarack (Larix larcinia) in fens 

and bogs. White spruce (Picea glauca),black spruce, 

lodgepole pine (Pinus contorta), quaking aspen (Populus 

tremuloides), balsam polar (Populus balsamifera), and 

paper birch (Betula papyrifera) grow on well drained, 

warm upland sites, rivers, and streams. Mat-forming 

lichens can constitute in excess of 60 percent of the 

winter food intake of caribou and reindeer (Longton 

1997). Th e abundant wetlands attract hundreds of 

thousands of birds (e.g., ducks, geese, loons, and swans) 

which come to nest, or rest and feed on their way to 

arctic breeding grounds.
50 Chapter 6—Taiga GTR-NRS-80

Th e severity of N deposition impacts depends on: 

duration and amount of N deposition, the form(s) 

of atmospheric N, the sensitivity of the ecosystem 

components; other environmental conditions; and 

management history (Bobbink et al. 2003). As a result, 

ecosystem response to N loading varies temporally, 

spatially, and by habitat type. Change or loss of habitats 

can have direct implications on the animals using these 

habitats.

6.3 Range of Responses Observed

Like the Tundra ecoregion, the paucity of North 

American studies and the many similarities in climate, 

topography, and vegetation communities across the 

circumboreal environment support consideration of 

European fi ndings to predict ecosystem N responses 

for North America. Th erefore, we include European 

data in this discussion of the range of observed 

responses. European analyses considered Taiga and 

Northern Forest ecoregions (CEC 1997) as a single 

boreal biogeographic region under the European 

University Information System (EUNIS) classifi cation 

system; hence the frequent use of the term ‘boreal’ 

in the following sections. Responses to N inputs are 

summarized in Table 6.1.

6.3.1 Deposition

Our knowledge of current N deposition and deposition 

eff ects in the North American taiga is limited. Wet 

deposition of inorganic N at the National Atmospheric 

Deposition Program (NADP) monitors at Denali 

National Park and Preserve and Fairbanks, Alaska, has 

been low and stable, averaging 0.23 kg ha-1 yr-1 (std. 

dev. = 0.129) from 1981 to 2007 (NADP 2008). Total 

nitrate (NO3

-

) deposition in snow at two interior Alaska 

sites north of Fairbanks in 1988 was consistent with this 

range and averaged 0.32 kg N ha-1 (Jaff e and Zukowski 

1993). Recent measurements of N from ammonium 

(NH4

+

) and NO3

-

 ions in throughfall deposition at 

remote sites in northeastern Alberta, Canada (Berryman 

and Straker 2008)ranged from 0.6 to 2.0 kgN ha-1 yr-1.

6.3.2 Forests and Woodlands

Plant community composition changes. A large number 

of studies, summarized by Bobbink et al. (2003), have 

demonstrated increases in abundance of nitrophilous 

species with increased N deposition over time and 

along N gradients. In Swedish boreal forests, changes 

in ground vegetation, e.g., decreasing whortleberry 

(Vaccinium myrtillus), occurred at deposition rates 

≥ 6 kg N ha-1 yr-1 (Nordin et al. 2005, Strengbom 

et al. 2003) and increased growth of wavy hairgrass 

(Deschampsia fl exuosa) at ≥ 5 kg N ha-1 yr-1 (Kellner and 

Redbo-Torstensson 1995, Nordin et al. 2005).

In an N fertilization experiment with 0, 12.5, and 50 

kg N ha-1 yr-1 (background atmospheric deposition 2 

to 3 kg N ha-1 yr-1), the abundance of wavy hairgrass 

increased signifi cantly after 3 years with 12.5 kg N ha-1

yr-1 while the abundance of whortleberry decreased 

(Strengbom et al. 2002). Bryophyte species are also 

responsive; the biomass of Schreber’s big red stem moss 

(Pleurozium schreberi) and a dicranum moss (Dicranum 

polysetum) fertilized with 25 and 30 kg N ha-1 yr-1 was 

reduced by 60 percent and 78 percent, respectively, after 

4 years (Mäkipää 1998).

Pest and disease resistance. In a large-scale fi eld study 

of 557 coniferous stands in Sweden, the occurrences 

of whortleberry, lingonberry (Vaccinium vitis-idaea) 

and wavy hairgrass, were investigated (Strengbom et 

al. 2003). Where N deposition was ≥ 6 kg ha-1 yr-1, 

whortleberry was less frequent and more susceptible to 

the fungal leaf pathogen Valdensia heterodoxa. Frequency 

of lingonberry was also strongly negatively correlated 

with increasing N deposition (Strengbom et al. 2003).

Whortleberry showed increased parasite burdens at 

N fertilization ≥ 12.5 kg ha-1 yr-1 (Nordin et al. 1998, 

2005; Strengbom et al. 2002). Disease incidence by 

the fungus Valdensia heterodoxa was more than twice 

as high in plots receiving 12.5 kg N ha-1 yr-1 and more 

than three times as high in plots receiving 50 kg N ha-1

yr-1 compared to controls. Th e abundance of the fungus 

V. heterodoxa on whortleberry was increased tenfold by 

25 kg N ha-1 yr-1 compared to 0.5 kg N ha-1 yr-1. As a 

consequence, whortleberry density decreased, and wavy 

hairgrass cover increased (Strengbom et al. 2002). In 

addition, shoots of whortleberry were signifi cantly more 

damaged by moth larvae such as the rusty tussock moth 

(Orgyia antique) after addition of 12.5 kg N ha-1 yr-1 in 
Chapter 6—Taiga GTR-NRS-80 51

the fi rst year of treatment (Nordin et al. 1998). Cover 

of wavy hairgrass continued to increase over 5 years 

of N additions between 8 and 12 kg ha-1 yr-1, induced 

by increased light penetration resulting from disease 

damage to whortleberry (Nordin et al. 2005).

Cold tolerance. Frost hardiness and plant tissue cold 

hardiness can be reduced by N additions. For example, 

after 12 years of fertilization with ammonium chloride 

(NH4Cl) at 15 kg N ha-1 yr-1, fi rst-year foliage in red 

spruce (Picea rubens) at a high elevation forest in New 

England (with ambient wet + dry deposition of ~10 kg 

N ha-1 yr-1) showed diminished cold tolerance, greater 

electrolyte leakage, and increased susceptibility to frost 

damage (Schaberg et al. 2002). Nutrient imbalances 

created by N and sulfur (S) fertilization and low 

potassium (K), phosphorus (P), and magnesium (Mg) 

status in southern Sweden and Denmark have been 

shown to increase the risk for development of cold 

temperature-related bark lesions in beech (Fagus spp.); 

advance bud burst of needles, putting young needles 

at risk for frost damage by a temperature backlash; 

and increase the frost-sensitivity of the inner bark and 

needles of Norway spruce (Picea abies) (Jönsson et al. 

2000, 2001, 2004a, 2004b).

N storage in cryptogam mats. Woodlands or tree-islands 

of pine with ground vegetation of mat-forming lichens 

are a common taiga vegetation type. In general, when 

inorganic N is received in low concentrations, either 

from natural rainfall (Hyvärinen and Crittenden 1998b) 

or applied in solution (Crittenden 1998), the estimated 

uptake effi ciency of (non-N2

-fi xing) mat-forming lichens 

has ranged from 90 to 100 percent. Internal recycling of 

N and P in mat-forming lichens improves nutrient-use 

effi ciency and is likely ecologically important in N- and 

P-limited environments typical of the taiga (Crittenden 

et al. 1994, Kytöviita and Crittenden 2007). Much of 

the eastern portion of the Taiga ecoregion is underlain 

by the Canadian Shield and soils can be shallow and 

acidic. During a short-term experimental addition of 

40 kg N ha-1 yr-1 to a small catchment of boreal pinereindeer lichen (Pinus-Cladonia) forest in Ontario, 

Canada, the higher export of mineral N from lichendominated bedrock surfaces compared to treed soil 

islands was attributed to the lower retention of N 

deposition in combination with leaching of mineralized 

N from lichen and moss patches (Lamontagne 1998). 

Net nitrifi cation increased or remained similar to 

reference sites in lichen patches, while N-amended forest 

islands had a strong tendency to consume NO3

-

 and 

produce NH4

+

 (Lamontagne and Schiff 2000). By the 

second year, lichen-covered bedrock surfaces no longer 

retained N additions. In contrast, N-amended and 

reference forest islands retained a similar proportion of 

N inputs, indicating that forest islands did not become 

N saturated in this time frame. However, because 

the components of the boreal shield landscape are 

hydrologically connected, there is concern that longterm N deposition to such a heterogeneous landscape 

will ultimately lead to N saturation of habitats with 

relatively high N retention capacities (Lamontagne and 

Schiff 1999). Finally, organic forms of N (e.g., free 

amino acids) serve as important plant N sources, and 

N deposition can disrupt the ratio between organic and 

mineral N supply in boreal soils (Näsholm et al. 1998, 

Nordin et al. 2001).

Epiphytic lichens and algae. Jack pine and black spruce

boreal forests of northeastern Alberta were sampled 

at 5, 10, 15, 20, 25, 30, 50, 60, 80, 100, and 120 km 

along cardinal directions from the Athabasca Oil Sands 

operations, a major source of nitrogen oxide (NOx) and 

sulfur dioxide (SO2) emissions for Canada. Nitrogen 

and S in the mat-forming gray-green reindeer lichen 

(Cladina rangiferina) and in two epiphytes, a tube 

(Hypogymnia physodes) and ring (Evernia mesomorpha) 

lichen, decreased with distance and were elevated 

within 30 km of the operations. Lichen community 

composition was weakly correlated with distance: 

sensitive shrubby beard and horsehair lichens were less 

abundant and showed dwarfi ng, hyper-growth of asexual 

structures, discoloration of the lichen, and increased 

parasitism by fungi at sites close to the Athabasca Oil 

Sands (Berryman et al. 2004). Preliminary data indicate 

that N deposition in canopy throughfall under jack pine 

forests at clean sites ranges from 0.6 to 2.0 kg ha-1 yr-1, 

increasing to 3.0 kg ha-1 yr-1 at sites near the mines. A 

new study is under way to relate deposition measures 

and modeled estimates of N and S to lichen response 

and to elucidate the relative roles of SO2

, NOx, metals, 

and alkaline dust from mining operations in observed 
52 Chapter 6—Taiga GTR-NRS-80

community responses (Berryman and Straker 2008). 

Th e correlation of N accumulated by lichen thalli with 

measures of deposition or distance from a point source is 

consistent with other studies in boreal forests in Eurasia 

(Bruteig 1993, Crittenden 2000, Crittenden et al.

1994, Hyvärinen and Crittenden 1998a, Walker et al.

2003). Information on lichen research methodology and 

terminology can be found in Chapter 4 of this report.

During a 10-year monitoring period in Sweden, 

patterns for sensitive epiphytic lichens were opposite to 

those observed for algae (Bråkenhielm and Quinghong 

1995). Th e number of individuals, total cover, and 

proportion of sensitive lichens decreased with growing 

season length and with increasing N and S deposition; 

the fi rst marked decrease occurred as N deposition 

exceeded 5 to 8 kg N ha-1 yr-1. In contrast, colonization 

rate and colony thickness of epiphytic green algae 

(mainly Protococcus viridis) increased with increases in 

deposition and growing season length, especially as N 

deposition exceeded 12 kg ha-1 yr-1. Poikolainen et al. 

(1998) observed increases in the abundance of green 

algae on conifers at lower deposition levels than did the 

Swedish researchers. Based on observations from 3009 

permanent Finnish forest plots between 1985 and 1995, 

increased algal cover was associated with N deposition 

levels as low as 3 kg ha-1 yr-1. Nitrogen concentration 

in the epiphytic tube lichen, Hypogymnia physodes, 

the splendid feather moss (Hylocomium splendens), 

Schreber’s big red stem moss, and bark of Scots pine 

(Pinus sylvestris) were also correlated with N deposition. 

Temperature was confounded with N and S deposition 

in both studies, as both temperature and deposition 

decrease in more northerly latitudes. 

Additions of N can cause changes in physiology and 

ultrastructure in common lichens of taiga ecosystems. In 

an electron microscopy study of the epiphytic horsehair 

(Bryoria capillaris) and tube (Hypogymnia physodes)

lichens, treatments of 560 μg m-3 nitrogen dioxide 

(NO2

) or 10 mM sodium nitrate (NaNO3), ammonium 

chloride (NH4

Cl), and ammonium nitrate (NH4NO3) 

induced accumulation of electron-opaque substances 

in the vacuoles of both the algal and fungal cells and 

general degeneration of the fungal cells (Holopainen 

and Karenlampi 1985). Th is demonstrates that N 

additions can infl uence cellular structure and metabolic 

processes in lichens—an important functional group in 

both boreal and arctic regions. See Chapter 5 (Tundra) 

for descriptions of eff ects on mat-forming and other 

terricolous lichens.

Ectomycorrhizal fungi. Th ere is no information on 

ectomycorrhizal community responses to N in the 

Taiga ecoregion. We postulate that, given the climate, 

species composition and soils, white and black spruce 

dominated ecosystems may have similar deposition 

thresholds to those seen in spruce forests of the same 

species in coastal Alaska (see Chapter 9, West Coast 

Marine Forests) and to those of the spruce-fi r forests of 

the northeastern United States (see Chapter 7, Northern 

Forests).

6.3.3 Ericaceous Shrublands (Heaths)

Heathlands are dwarf shrub communities that 

form on shallow peat or drained soils. Nitrogenenrichment responses of heathland plant communities 

are characterized by increased growth of grasses 

and decreased growth of dwarf shrubs, lichens, and 

bryophytes (Allen 2004). Initially, shrubs respond 

to N additions with increased growth and tissue 

N concentrations, and the fi rst adverse eff ects are 

shade-induced declines in lichen and bryophyte cover 

(Cornelissen et al. 2001). However, long-term N 

enrichment decreases root biomass and increases the 

susceptibility of shrubs to insect attack and frost. Th ese 

eff ects, together with increased competition from 

grasses, cause a deterioration of the dominant shrub 

community. During a 5-year fertilization study (10.8 

kg ha-1 yr-1 modeled ambient total deposition plus a 

10 kg N ha-1 yr-1 application) of low-alpine heather 

(Calluna vulgaris) heaths in the Cairngorm Mountains 

of Scotland, species richness was reduced by exposure 

to added N deposition (Britton and Fisher 2007). 

Th e lichen component of the vegetation was most 

sensitive to N additions, although vascular plants were 

also aff ected via interactions with climate (Britton and 

Fisher 2007). Curtis et al. (2005) used isotope tracers in 

other European heaths and moorlands to demonstrate 

that bryophytes and lichens were important sinks for 

N deposition, reducing the amount of inorganic N 

available to both higher plants and soil microbes in 
Chapter 6—Taiga GTR-NRS-80 53

grassland and ericoid shrub-dominated catchments. 

However, as N deposition increased from 2 to 30 kg 

ha-1 yr-1, these sinks became saturated and cryptogam 

biomass decline was accompanied by increased NO3

-

leaching (see also Emmett 2007).

6.3.4 Peatlands

Peatlands are moss-dominated bogs and fens that are 

prevalent in the Taiga ecoregion. About 43 percent 

of Alaska can be classifi ed as wetlands, much of that 

peatlands, contrasting strongly with the contiguous 

48 states where wetlands comprise barely 5 percent of 

the total land area (Hall et al. 1994). Th e sensitivity of 

peatlands to N deposition decreases along a poor-to-rich 

productivity gradient from ombrotrophic bogs, which 

receive nearly all nutrient inputs from the atmosphere, 

to poor and rich fens, which receive additional nutrients 

from slow moving ground water and mineral sources 

(Vitt et al. 2003). Th e general progression of N 

deposition eff ects is: 

1) At 0.7 to 8.1 kg N ha-1 yr-1 peat accumulation 

increases with N deposition due to increases in 

net photosynthesis and growth of previously 

N-limited mosses (Moore et al. 2004, Vitt et al. 

2003). With average bulk deposition of 0.81 kg 

N ha-1 yr-1 and 1.14 kg S ha-1 yr-1 from 2005 to 

2008, a dominant sphagnum moss (Sphagnum 

fuscum) in 10 remote bogs across northeastern 

Alberta was N-limited, and no diff erences in 

moss growth or net primary productivity were 

observed compared to previously published 

values for this area (Wieder et al. 2010).Th ere 

are some indications that net photosynthesis 

in this sphagnum species may peak as early as 

3 kg N ha-1 yr-1 in some Canadian peatlands 

(Vitt et al. 2003). (Oligotrophic mosses are the 

predominant N-sink and biomass in peatlands).

2) At 12 to 18 kg N ha-1 yr-1, growth rates do not 

increase further and mosses begin to accumulate 

foliar N (Lamers et al. 2000, Moore et al. 2004, 

Vitt et al. 2003).

3) At >18 kg N ha-1 yr-1, the natural moss N fi lter 

fails and N begins to leach from the saturated 

moss layer, altering competitive relationships 

(Lamers et al. 2000). Vegetation changes 

are characterized by decreases in cover of 

oligotrophic mosses and carnivorous plants and 

increases in graminoids, especially cotton grass 

and certain ericaceous shrubs (Allen 2004).

Chapter 17 (Wetlands) contains a more detailed 

discussion of these and interrelated responses, especially 

in relation to P availability, water availability, and 

climate.

6.3.5 Tundra

Areas of tundra are included in the Taiga ecoregion, 

notably at the northern boundaries. See Chapter 5 

(Tundra) for descriptions of the range of responses.

6.4 Critical Loads Estimates

Th ere are many indications that low levels of N 

deposition can aff ect community composition, 

abundance, net photosynthesis, N accumulation, 

physiology, and ultrastructure of lichens, mosses, and 

algae (see Table 6.1). Th ese taxa, especially peatland 

sphagnum mosses and reindeer lichens, are dominant 

components of taiga ecosystems. Application here 

of a model relating epiphytic lichen community 

composition of Oregon and Washington coniferous 

forests to N deposition (Chapter 4, Geiser et al. 2010) 

yields a critical load estimate of 1 to 3 kg ha-1 yr-1 for 

Taiga ecoregion forests and woodlands. Th is value was 

calculated by substituting a realistic precipitation range 

of 20 to 80 cm and applying a conservative communitycomposition response threshold allowing no less than 41 

percent oligotrophs or more than 27 percent eutrophs. 

For perspective, about 85 percent of sites in the western 

Oregon and Washington study area did not exceed this 

response threshold (Geiser et al. 2010). Considering 

this result and those of Moore et al. (2004), Vitt et al. 

(2003), Strengbom et al. (2003), Berryman et al. (2004), 

Berryman and Straker (2008), and Poikolainen et al. 

(1998), we recommend that the critical load for lichen, 

moss, and algae of the North American Taiga ecoregion 

be no more than 1 to 3 kg ha-1 yr-1. Th is estimate can 

be considered fairly reliable, pending results from more 

defi nitive work currently in progress in North America. 

Critical loads are summarized in Table 6.2.
54 Chapter 6—Taiga GTR-NRS-80 Site Critical load for N dep. kg N ha-1 yr-1 Reliability Response Comments Study Epiphytic lichens 1-3 (#) Community composition shift Application of model developed for Marine West Coast Forests ecoregion to the Taiga ecoregion Geiser et al. 2010, Chapter 4 Peatlands, NE Canada 2.7-8.1 # Increase in peat accumulation N addition and gradient studies Moore et al. 2004 Boreal forest, Sweden <3 (#) Decreased frequency of the shrub lingonberry, especially in spruce (as opposed to pine) stands Possible interactions with other biotic and abiotic factors Strengbom et al. 2003 Black spruce and jack pine forests, NE Alberta 3 (#) Increased lichen N concentrations, morphological damage to and decreased abundance of sensitive epiphytic lichens N effects likely confounded and/or amplifi ed by elevated SO2, alkaline dust and/or metal deposition Berryman et al. 2004; Berryman and Straker 2008 Boreal forest, Finland 3 (#) Increased cover of green algae on conifers Increasing temperature, decreasing S may also be contributing to increased algal cover Poikolainen et al. 1998 Peatlands, NE Alberta >3 # Gradual decrease in net primary photosynthesis of Sphagnum fuscum Combined multiple N- addition studies Vitt et al. 2003 Boreal forest, 

Sweden

5-8 (#) Decreases in the proportion of 

N-sensitive epiphytic lichens

Confounding temperature effect

Bråkenhielm and Quinghong 1995

Boreal forest, Sweden

6 ## Increase in grass cover, esp. wavy 

hairgrass; decreased growth of shrubs whortleberry and lingonberry

Long term experiment in area with low background deposition using low N addition rates

Nordin et al. 2005

Boreal forest, Sweden

6 # Decreased frequency of the shrub, 

whortleberry, concurrent with rates of parasitism by Valdensia heterodoxa

Possible interactions with other biotic and abiotic factors

Strengbom et al. 2003

Low-alpine heather heaths, Cairngorm Mtns, Scotland

5-15 # Reduced species richness of lichens; 

also higher plants via interactions with climate

Britton and Fisher 2007 

Northern Europe 10-15; 10-20 # Increased mineralization, nitrifi cation 

and N leaching of soils

Bobbink et al. 2003, de Vries et al. 2007

Table 6.1—Responses to N input relevant to the North American taiga. Reliability rating: ## reliable; # fairly reliable; (#) expert judgment
Chapter 6—Taiga GTR-NRS-80 55

Northern Europe 10-15 # Changes in composition and cover 

of understory vascular plants, bryophytes, lichens, and free-living algae

Bobbink et al. 2003, de Vries et al. 2007

Northern Europe 10-20 (#) Reduced sporocarp production, 

changed/reduced below-ground mycorrhizal species composition

Bobbink 2003, de Vries et al. 2007

Boreal forest, Sweden

<12 # Increased fungal parasitism 

of whortleberry by Valdensia heterodoxa; enhanced growth of wavy hairgrass from defoliation of parasitized whortleberry and from N addition

Strengbom et al. 2002 

Boreal forest, Sweden

12 (#) Increased abundance of the green 

alga, Scoliciosporum chlorococcum, on conifers 

Confounding temperature effect

Bråkenhielm and Quinghong 1995

Red spruceHigh elevation New Hampshire spruce fi r forest 

<16 # First year foliage showed lower 

membrane instability, increased electrolyte losses, greater susceptibility to frost damage

Similar effects were observed after 12 years of fertilization at 15.7 and 31.4 kg N ha-1 yr-1, 

suggesting that the CL is well below 15.7

Schaberg et al. 2002

Norway spruce stand, southern Finland

25 (#) Decreased biomass of dominant 

oligotrophic mosses Pleurozium schreberi and Dicranum polysetum

4 yr study. 30 kg S also added therefore effect not due to N alone 

Mäkipää 1998 

Pinus-Cladina boreal forest, Ontario, Canada

<40 ## Increase in nitrifi cation (lichendominated bedrock) leading to N 

leaching; increased production of NH4

+

in tree islands with deeper soils

Hydrological connectivity in taiga ecosystems may permit cascading N saturation in boreal shield landscapes

Lamontagne 1998, Lamontagne and Schiff 1999
56 Chapter 6—Taiga GTR-NRS-80

Provisional N critical loads using ectomycorrhizal fungal 

abundance and diversity as indicators ranged from 5 to 

7 kg ha-1 yr-1, based on critical loads determined using 

the same indicators in similar forest types in the Marine 

West Coast Forests and Northern Forests ecoregions 

(Chapters 9 and 7; Lilleskov 1999; Lilleskov et al. 2001, 

2002, 2008).

Based on reliable estimates from European work 

described previously and in section 6.5, critical 

loads for alterations in shrub and grass community 

composition in the Taiga ecoregion could be set at 6 

kg ha-1 yr-1 (Nordin et al. 2005, Strengbom et al. 2003; 

Table 6.1). However, the higher European critical 

loads estimates correspond to the levels of N added 

in experimental manipulations, which may far exceed 

the level of N input that would eventually induce a 

response. As deposition levels exceed 5 to 15 kg N 

ha-1 yr-1, the following become evident: changes in 

woodland, heath, and peatland communities of vascular 

plant (especially shrubs and grasses) and mycorrhizal 

communities; increased foliar nutrient imbalances; and 

increased susceptibility of vegetation to frost, drought, 

pathogens and pests. At deposition levels exceeding 15 

kg N ha-1, changes in nitrifi cation rates, soil saturation 

and N leaching can be documented, especially when 

the N-holding capacity of moss and lichen mats are 

exceeded.

6.5 Critical Load Data from Europe or 

from Other Ecoregions

European critical loads for boreal forest ecosystems 

(Bobbink et al. 2003) as updated by the Alterra Rapport 

1382 (de Vries et al. 2007) are 10 to 15 kg N ha-1 yr-1

for soil processes (increased mineralization, nitrifi cation 

and N leaching); 10 to 20 kg N ha-1 yr-1 for increased 

risk of foliar nutrient imbalances (decreased P, K 

and Mg to N ratios in foliar tissue) in coniferous and 

deciduous trees and changes in community composition 

and of ground vegetation, lichens, and mycorrhizae; 

and 15 to 25 kg N ha-1 yr-1 for increased susceptibility 

to frost, drought, pathogens, and pests, and increases 

in free algae. Based largely on biodiversity changes 

documented by Nordin et al. 2005 (see section 6.3), de 

Vries et al. (2007) recommended decreasing the overall 

critical load for boreal forests from 10 to 20 (Bobbink et 

al. 2003) to 5 to 10 kg N ha-1 yr-1.

European critical loads (Bobbink et al. 2 003) for other 

taiga ecosystem components are 5 to 10 kg ha-1 yr-1 for 

tundra with permafrost, 10 to 15 kg ha-1 yr-1 for alpine 

and subalpine scrub without permafrost, and 10 to 20 

kg ha-1 yr-1 for wet and dry heaths. Critical loads for 

Tundra and Northern Forests ecoregions are especially 

relevant to tundra-covered areas and large river valley 

areas of the Taiga ecoregion, respectively. All three 

ecoregions share some fl oral and faunal species.

Table 6.2—Empirical critical loads of nutrient N for the Taiga ecoregion. Reliability rating: ## reliable; # fairly reliable; 

(#) expert judgment

Ecosystem 

component

Critical 

load for N 

deposition

kg ha-1 yr-1

Reliability Response Comments Study

Lichen, moss, 

and algae

1-3 # Changes in alga, 

bryophyte, and lichen 

community composition, 

cover, tissue N or 

growth rates

Berryman et al. 2004, 

Berryman and Straker 2008, 

Geiser et al. 2010, Moore 

et al. 2004, Poikolainen et 

al. 1998, Strengbom et al. 

2003, Vitt et al. 2003

Mycorrhizal 

fungi, sprucefi r forests

5-7 (#) Change in 

ectomycorrhizal fungi 

community structure

Expert judgment

extrapolated from Marine 

West Coast spruce and 

northern spruce-fi r forest

Lilleskov 1999; Lilleskov et 

al. 2001, 2002, 2008

Shrublands 6 ## Alterations in shrub and 

grass cover; increased 

parasitism of shrubs

Long term, low N addition 

study: shrub cover 

decreased, grass cover 

increased

Nordin et al. 2005, 

Strengbom et al. 2003 
Chapter 6—Taiga GTR-NRS-80 57

6.6 Future Research Directions and 

Gaps in Data

Research is needed to confi rm that European critical 

loads and study results are applicable to North America. 

Long-term studies at low levels of experimentally applied 

N (i.e., 1 to 5 kg ha-1 yr-1) are most needed. Although 

local Alaskan population and emissions sources are 

relatively small, regional and trans-Pacifi c sources of N 

are increasing with massive energy development projects, 

industrial expansion, and high latitude population growth. 

Fire frequency and intensity in boreal ecosystems are 

increasing with global climate change; fi re will continue to 

be a major regional source of NOx

. Because temperature 

increases are predicted to be greater in northern latitudes, 

North American studies are needed to elucidate the 

interplay between climate change, N volatilization, and 

N deposition on dry- and wet-land biota, communities, 

and ecosystems. Indeed, climate change may be the most 

important driver of vegetation changes in most of the 

Alaskan portion of the ecoregion, considering the low 

current background levels of N deposition.

Finally, a greater understanding is needed regarding 

the underlying mechanisms responsible for N-aff ected 

changes, in addition to quantifi cation of the functional 

relationships between N deposition and specifi c 

ecological responses. Such a mechanistic understanding 

of how N addition alters low N ecosystems could be 

used to inform modeled predictions.

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Chapter 7—Northern Forest GTR-NRS-80 61

7 NORTHERN FORESTS

L.H. Pardo, C.L. Goodale, E.A. Lilleskov, L.H. Geiser

7.2 Ecosystem Responses to N 

Deposition

Most northern forests are naturally nitrogen (N) limited; 

cold temperatures limit rates of decomposition and slow 

rates of release of N from decomposing organic matter. 

Th ese low rates of N mineralization limit the supply of 

N to plants and microbes. Th ere is strong competition 

for N between plants and microbes and most of the N 

entering the ecosystem is tightly retained in soils and in 

vegetation, with little loss of inorganic N to drainage 

waters. When air pollution increases the supply of N 

to these ecosystems through N deposition, a range of 

ecosystem responses, described below, are anticipated. 

Some of these responses occur as gradual changes, others 

as cumulative threshold responses in progressive stages 

of N saturation (cf. Aber et al. 1998). See Chapter 1 for 

a more detailed explanation of the stages. Briefl y, early 

stages (especially stage 1) are marked by increased supply 

of N to plants and to microbes, indicated by increasing 

plant foliar N concentration, growth rate, and soil N 

mineralization rate. Cumulative N inputs may decrease 

the carbon:nitrogen (C:N) ratio of soils and change rates 

of decomposition. Responding to increased N supply 

and low soil C:N ratios, net nitrifi cation marks a next 

stage of ecosystem response (stage 2). Nitrifi cation can 

acidify the soil and produce high levels of nitrate (NO3

-

), 

a form of N that readily leaches with runoff . Th is phase 

of response is marked by increased NO3

-

 concentrations 

in surface water and leaching of N from catchments. 

When NO3

-

 leaches, it removes base cations from the 

soil and leads to soil and drainage water acidifi cation. 

Collectively, acidifi cation and nutrient imbalances 

may cause growth declines or increased mortality 

rates (stage 3). Th ese direct eff ects of N deposition 

may also lead to increased susceptibility to secondary 

stresses such as winter injury, drought damage, pest 

outbreak, and diseases, and may reduce cold tolerance 

(Schaberg et al. 2002). Declining vegetation is likely 

to take up less N than healthy vegetation, and so losses 

of NO3

-

 are expected to accelerate as N saturation 

7.1 Ecoregion Description

Th e Northern Forests ecological region spans much 

of Canada, from Saskatchewan to Newfoundland; its 

southern portion extends into the northern United 

States (CEC 1997). Th e U.S. component includes the 

northern hardwood and spruce-fi r forest types and 

encompasses parts of the Northeast (mountainous 

regions in Pennsylvania, New York, New Jersey, 

Connecticut, Massachusetts, Maine, and most of 

Vermont and New Hampshire) and the northernmost 

parts of the upper Midwest (Minnesota, Wisconsin, 

Michigan; CEC 1997; see Chapter 2, Figure 2.2). Th e 

ecoregion description is adapted from CEC (1997).

In Canada, the ecoregion is underlain by ancient 

Canadian Shield bedrock interspersed with glacial 

moraine deposits, creating a hilly terrain dotted with 

numerous lakes. Soils are generally coarse-textured and 

nutrient-poor, characterized typically by thick organic 

accumulations in the forest fl oor and leached layers 

in the mineral soil below. Shallow soils and exposed 

bedrock are covered with various plant communities 

dominated by shrubs, forbs, and lichens. Peatlands 

are extensive in northern Minnesota. Th e climate is 

characterized by long, cold winters (means of -20.5 

°C in the west to -1 °C in the east) and short, warm 

summers (means of 11 to 18 °C). Annual precipitation 

varies from 400 to 1,000 mm. Th e ecoregion is more 

than 80 percent forested, and dominated by conifers, 

mostly white and black spruce (Picea glauca and P. 

mariana), jack pine (Pinus banksiana), balsam fi r (Abies 

balsamea), and tamarack (Larix laricina) (CEC 1997). 

Toward its southern and eastern extremes in the upper 

midwest and northeastern United States, the ecoregion 

includes more paper birch (Betula papyrifera), yellow 

birch (Betula alleghaniensis), quaking aspen (Populus 

tremuloides), balsam poplar (Populus balsamifera), eastern 

white pine (Pinus strobus), red pine (Pinus resinosa), 

sugar maple (Acer saccharum), red maple (Acer rubrum), 

beech (Fagus spp.), red spruce (Picea rubens), and 

various species of oak (Quercus spp.).
62 Chapter 7—Northern Forest GTR-NRS-80

proceeds. In addition to these ecosystem-level responses, 

N deposition can aff ect the competitive balance of 

diff erent species, leading to changes in composition of 

mycorrhizal fungi, lichen, bryophyte, understory shrub 

and herb, and tree species composition. In this chapter, 

we review the evidence for impacts of N deposition 

on overstory trees, herbaceous vegetation, mycorrhizal 

fungi, and lichens in the Northern Forests ecoregion.

7.3 Range of Responses Observed

Th ree types of studies provide empirical information on 

the eff ects of N on ecosystem properties: (1) long-term 

observations at sites receiving increasing or chronically 

elevated N deposition; (2) long-term, low- to moderatelevel N fertilization studies; and (3) observational studies 

along gradients of N deposition. All three study types 

are discussed below, and each approach has its strengths 

and limitations. Long-term observations can reveal 

trends through time, but also can be infl uenced by a 

great many factors in addition to N deposition, factors 

which are diffi cult to disentangle. Fertilization studies 

may provide some additional insight about the timing 

and nature of responses of forests to elevated N inputs. 

Often, fertilization studies utilize such high N inputs 

that they do not help defi ne the actual threshold when a 

response is induced. Gradient studies examine ecosystem 

responses along gradients of N deposition, but can also 

suff er from diffi culty in separating out eff ects of N from 

eff ects of other covariates.

7.3.1 Foliar N, Growth, and Mortality

Foliar N concentrations increased in most N fertilization 

experiments in the Northern Forests ecoregion 

(Table 7.1). Th is increase in foliar N might be expected 

to increase plant growth, since photosynthesis typically 

increases with foliar N concentration within and 

across most natural ecosystems. Increased foliar N was 

accompanied by increased tree growth in fertilized sugar 

maple stands in Michigan (Pregitzer et al. 2008) and 

Maine (Elvir et al. 2003). However, fertilized red spruce 

trees in Maine (Elvir et al. 2003) and black spruce and 

balsam fi r stands in northern Quebec (Houle and Moore 

2008) showed no change in growth rates. At a subalpine 

spruce-fi r forest at Mt. Ascutney, Vermont, long-term 

fertilization with ammonium chloride (NH4Cl) at 15.7 

kg N ha-1 yr -1 (in addition to ambient wet plus dry N 

deposition of approximately 10 kg ha-1 yr -1) increased 

foliar N concentration and decreased growth and 

increased mortality, leading to an approximately 20 

percent decline in live basal area after 14 years (McNulty 

et al. 2005). Other eff ects observed at this fertilization 

rate were reductions in foliar membrane-associated Ca, 

reduced cold tolerance, and increased winter freezing 

injury (Schaberg et al. 2002). Th e responses observed at 

the high treatment plot, where 31 kg N ha-1 yr-1 above 

ambient deposition were added, are far more striking. 

In these plots, mortality is high (40 percent decline in 

live basal area) and species composition changes were 

induced after only 14 years of N addition (McNulty et 

al. 2005).

Overall, these fertilization experiments show modest 

growth enhancement by some, but not all, species. 

Sometimes the initial growth enhancement is followed 

by growth declines or increased mortality in the longterm studies, especially at the highest rates of N inputs 

or sites with soils containing few base cations. It is 

unclear precisely where this threshold for detrimental 

eff ects occurs, although the experimental work at Mt. 

Ascutney demonstrated substantial mortality by red 

spruce in response to just 15.7 kg N ha-1 yr - 1 above 

the ambient deposition of approximately 10 kg N ha-1

yr-1. While Mt. Ascutney clearly represents the sensitive 

end-member of forest ecosystems in the Northeast, it 

is instructive because the nonlinear path that the high 

treatment plots followed may also occur over time at the 

low treatment plot. It is possible, then, that inputs even 

lower than 26 kg N ha-1 yr-1 would lead to N saturation 

over an extended period of time (e.g., 50 to 100 years).

Gradient studies have shown mixed evidence for foliar 

N response to atmospheric deposition to northern 

forest ecosystems. In a broad synthesis of data from 

354 plots in upland forests across the Northeast, 

Aber et al. (2003) did not fi nd patterns of increasing 

foliar N concentration with deposition, even when 

controlling for tree species; variation in foliar N was 

most strongly associated with elevation and climatic 

variables. In red spruce, foliar lignin:N ratio was related 

to N deposition, but deposition covaried with elevation 
Chapter 7—Northern Forest GTR-NRS-80 63

Table 7.1—Response of Northern Forest ecosystems to experimental N fertilization (≤ 50 kg N ha-1 yr -)

Site Location Michigan 

(4 sites)

Mt. Ascutney, 

Vermont

---- Bear Brook, ---- 

Maine

Adirondacks, 

NY (3 sites)

Northwest 

Ontario

---- Northern ----

Quebec 

Forest type Sugar maple Montane 

spruce-fi r

Northern 

hardwoods

Red 

spruce

Northern 

hardwoods

Rock, Jack 

pine / spruce

Black 

spruce

Balsam 

fi r

Type of N deposition Wet + dry Wet + dry Wet + dry Throughfall Bulk Wet only

N Deposition 

(kg ha-1 yr-1)

6.8-11.8 10 8.4 4.6-12.0 3.4-6.5 3 5.7

N fertilization rate 

(kg ha-1 yr-1)

+30 +15.7, +31.4 +25 +14, +28 +36-40 +9, +30 +18, +60

Form of N fertilizer NaNO3 NH4Cl (NH4)2SO4 HNO3, 

(NH4)2SO4

NaNO3 NH4NO3

Start year 1994 1988 1989 1991 1995 2001

Duration (years) 10 14 13 3 2 3 3 - Response relative to Controla

-

Tree growth + – + 0 ND ND 0 0

Mortality + + ND ND ND ND ND ND

Foliar %N + + + + ND ND + 0

Foliar %Ca ND – 0, – – ND ND + 0

NO3

-

 leaching + ND + 0, 0, + + + +

Cation loss ND ND + ND ND ND ND

Soil C/N 0 0 0 – ND ND ND ND

N mineralization 0 0, – + + ND ND ND ND

Nitrifi cation + + + + ND ND ND ND

Soil respiration – – 0 ND ND ND ND ND

Microbial biomass 0 – 0 ND ND ND ND ND

References 1-5 6-7 7-11 12 13 14 14

a

Responses are indicated as positive (+), negative (–) or no response (0) in fertilized relative to control plots, or no data (ND) in indicated 

references.

References: 1) Burton et al. 2004; 2) Pregitzer et al. 2004; 3) Pregitzer et al. 2008; 4) Zak et al. 2004; 5) Zak et al. 2006; 6) McNulty et al. 2005; 

7) Wallenstein et al. 2006; 8) Elvir et al. 2003; 9) Elvir et al. 2006; 10) Fernandez et al. 2003; 11) Jefts et al. 2004; 12) Mitchell et al. 2001; 13) 

Lamontagne and Schiff 1999; 14) Houle and Moore 2008.

and climatic variables within this dataset (Aber et al. 

2003). However, other studies have shown that foliar N 

concentration increases with increasing N deposition. 

McNeil et al. (2007) saw an increase in foliar N 

concentration for eight of nine canopy tree species across 

73 plots spanning an N deposition gradient (3.5 to >7 

kg ha-1 yr-1 of wet-only N deposition) in the Adirondack 

Park, New York. American beech (Fagus grandifolia) did 

not respond, but yellow birch (Betula alleghaniensis), 

paper birch, red maple (Acer rubrum), sugar maple,

eastern white pine, eastern hemlock (Tsuga canadensis), 

balsam fi r, and red spruce did. Another regional foliar 

analysis, which included 2702 foliar samples from 434 

plots in the Northeast, found signifi cant correlations 

between N deposition (5 to 12 kg ha-1 yr-1 of wet plus 

dry deposition) and foliar N concentration for both 

hardwoods and conifers (Pardo et al. 2006). Th is study 

also examined foliar 15N, another measure used to 

assess the N status of ecosystems because it provides 

an integrated measure of past N cycling and often 

increases in response to disturbance (Emmett et al. 

1998, Nadelhoff er and Fry 1994). Foliar 15N increased 

with deposition, although correlations were stronger 

for conifers than for hardwoods (Pardo et al. 2006); 

in all cases the correlation between N deposition and 

foliar N concentration was stronger than with foliar 
64 Chapter 7—Northern Forest GTR-NRS-80

15N. In another gradient study evaluating N cycling in 

red spruce stands from Maine to New York, sampling 

in 1987 and 1988 (McNulty et al. 1991) and again in 

1999 (Boggs et al. 2007) showed an increase in foliar 

N concentration both over time and with deposition 

along the gradient from 5 to 11 kg N ha-1 yr-1 (Pardo 

et al. 2007). Hardwood foliar N concentration was 

measured only in 1999 and increased with deposition 

along the gradient. Th ese relationships were weaker 

than for spruce, but were statistically signifi cant (Pardo 

et al. 2007). Th ese studies broadly demonstrate that 

foliar N concentration in Northern Forests tends to 

increase with increasing N inputs. Nonetheless, because 

of the variability in response and the lack of a threshold 

above which the rate of change is altered, it is diffi cult 

to use foliar N concentration alone as an indicator of N 

saturation.

Until recently, there has been little evidence outside of 

fertilization experiments for either enhanced growth 

or mortality in U.S. forests attributable to elevated 

N deposition. Th is lack of evidence is largely due to 

the diffi culty of detecting such a signal against the 

large background variation due to factors that aff ect 

growth in natural ecosystems and the relative paucity 

of such studies. Caspersen et al. (2000) discerned a 2.0 

percent enhancement of forest biomass in several states 

(Minnesota, Michigan, Virginia, North Carolina, and 

Florida) over the last century, but could not identify 

which mechanism (e.g., rising atmospheric CO2

concentrations, climate changes, N deposition) might 

be responsible for this growth increase. Similarly, Hicke 

et al. (2002) reported changes in forest growth across 

North America over the last two decades derived from 

satellite observations of canopy properties, but saw no 

trends that were easily attributable to N deposition. 

In Europe, regional patterns of forest growth are 

largely driven by forest management decisions, but 

growth enhancement from N deposition has long 

been speculated (e.g., Kauppi et al. 1992, Nabuurs et 

al. 2003). Recently, Magnani et al. (2007) reported 

a strong correlation between wet N deposition and 

estimates of net ecosystem production derived from 

chronosequences and eddy fl ux towers at 20 temperate 

forest sites. Th is correlation implied a very large 

stimulation of C sequestration with increasing N 

deposition of approximately 180 to 480 kg C ha-1 yr -1

per 1 kg ha-1 yr -1 received in N deposition (Magnani 

et al. 2007). Several researchers have questioned the 

plausibility of so large an eff ect (de Vries et al. 2008, 

Högberg 2007, Sutton et al. 2008).

In a recent gradient analysis, Th omas et al. (2010) 

used forest inventory data from plots spanning the 

northern and eastern forests in the northeastern 

United States to discern a N deposition-induced treegrowth enhancement in plot-level biomass increment 

amounting to 5.5 percent per kg N ha-1 yr-1, or 61 kg 

C ha-1 yr -1 per 1 kg ha-1 yr-1 received in N deposition. 

Overall, this plot-level growth enhancement did not 

saturate or decline within the range of wet + dry 

deposition experienced in the region (3 to 11 kg N 

ha-1 yr-1). However, diff erent tree species responded 

in diff erent ways: some species had no response, some 

showed steady growth increases across the range of N 

deposition (e.g., yellow-poplar [Liriodendron tulipifera], 

black cherry [Prunus serotina], white ash [Fraxinus 

americana]), some showed growth rate peaks at midlevels of N deposition (quaking aspen, scarlet oak 

[Quercus coccinea]), and one showed growth declines 

with increasing deposition (red pine). Tree survivorship 

increased with N deposition for some species (black 

cherry, red maple) and decreased for others (e.g., 

quaking and bigtooth aspen [Populus grandidentata], 

yellow birch, basswood [Tilia spp.], chestnut oak 

[Quercus prinus], and scarlet oak). Overall, this analysis 

indicates that current rates of N deposition enhance 

carbon sequestration in northeastern forests, but that 

N deposition can have negative impacts on growth and 

survivorship for some sensitive species.

7.3.2 Nitrate Leaching and Soil N

 Conceptual models of N saturation indicate that 

NO3

-

 leaching should increase in response to long-term 

increases in or chronic exposure to atmospheric N 

deposition (Aber et al. 1998, Stoddard 1994). In fact, 

long-term NO3

-

 leaching trends in the northeastern 

United States have been inconsistent. Some sites have 

shown increases in NO3

-

 loss over the last 20 to 30 years, 

particularly in regions with the highest deposition (10 

to 15 kg N ha-1 yr-1) such as the Catskill Mountains, 

New York (Burns et al. 2006) or Fernow Experimental 
Chapter 7—Northern Forest GTR-NRS-80 65

Forest, West Virginia (Peterjohn et al. 1996, Adams 

et al. 2006); however, there is considerable temporal 

variability within these records. At other sites in the 

Adirondacks (Driscoll et al. 2003, 2007) and at the 

Hubbard Brook Experimental Forest (Dittman et al. 

2007) and other sites in the White Mountains, New 

Hampshire (Goodale et al. 2003), surface water NO3

-

concentrations have decreased over the last 20 to 30 

years despite near-constant deposition.

Th e absence of a straightforward relationship between 

N deposition and surface water NO3

-

 leaching over time 

highlights the value of other approaches for assessing 

N eff ects on leaching, such as gradient studies and 

fertilization experiments. In a synthesis of NO3

-

 data 

across a northeastern N deposition gradient, Aber et al. 

(2003) report that surface water NO3

-

 concentrations in 

catchments receiving less than ~8 kg N ha-1 yr-1 rarely 

exceed 1 μmol L-1 (Fig. 7.1). Nitrate concentrations 

increase on average above this threshold, though with 

substantial site-to-site variation. Th e wide range of 

NO3

-

 concentrations observed likely refl ect the many 

factors other than N deposition that impact NO3

-

leaching, such as tree species composition, soil and 

stand characteristics, land-use history, and hydrology. 

For example, a high abundance of sugar maple has 

been associated with elevated nitrifi cation and NO3

-

leaching (Lovett and Rueth 1999, Lovett et al. 2004). 

Shallow soils at Mt. Ascutney in southern Vermont 

may be implicated in the site’s sensitivity to N inputs 

and rapid increases in nitrifi cation rates (McNulty and 

Aber 1993, McNulty et al. 1996, 2005). Old-growth 

stands have been shown to have higher nitrifi cation 

rates and NO3

-

 leaching than second-growth historically 

logged or burned stands in the White Mountains of 

New Hampshire (Goodale and Aber 2001). However, 

elevated NO3

-

 losses at some sites, including the oldgrowth Bowl Research Natural Area, New Hampshire, 

may refl ect hydrologic fl ow paths more than biological 

controls on N saturation (Martin et al. 2000, Pardo et 

al. 2004). Because so many factors can aff ect stream 

NO3

-

 losses, determining thresholds for NO3

-

 leaching 

in response to N deposition requires information from a 

very large number of catchments to discern the leaching 

signal over variation driven by other factors.

Th e Aber et al. (2003) N gradient study included a 

synthesis of soil data from 250 plots in the northeastern 

United States, yet detected only weak relationships 

between N deposition and soil N cycling measures 

(e.g., forest fl oor C:N, nitrifi cation), due to the great 

number of additional factors that aff ect N cycling 

Spring (n=212)

Estimated N Deposition (kg N ha-1 yr-1)

NO3

- (Pmol L-1)

Summer (n=350)

6 8 10 12

0

10

20

30

40

50

60

4 6 8 10 12

Lakes

Streams

Figure 7.1—Surface water NO3

-

 concentrations increase with estimates of N deposition at the base of study 

watersheds in both summer and spring. N deposition to the entire watershed may be approximately 1 kg ha-1

yr-1 greater than at the base (From Aber et al. 2003).
66 Chapter 7—Northern Forest GTR-NRS-80

beyond just N deposition. Stronger relationships were 

found between forest fl oor C:N ratio and nitrifi cation, 

which presumably should lead to increased N availability 

for NO3

-

 leaching. Th e study evaluating N cycling 

in red spruce stands across an N deposition gradient 

from Maine to New York found signifi cant increases 

in net nitrifi cation potential with increasing deposition 

in 1987 (McNulty et al. 1991). When the same sites 

were evaluated in 1999, the relationship between N 

deposition and net nitrifi cation potential was still 

observed, although the rate of nitrifi cation had decreased 

signifi cantly over time (Boggs et al. 2007). Decreasing 

patterns in stream water NO3

-

 and nitrifi cation rate 

contradict the expectation that an ecosystem exposed to 

elevated N deposition over an extended period would 

move toward N saturation. Instead, these data suggest 

that the other factors beyond N deposition, such as 

climate, contribute to the observed patterns of ecosystem 

N cycling and loss.

7.3.3 Herbaceous Layer

Th ere are few studies of N deposition eff ects on the 

herb layer in northern forests. In a low-level fertilization 

study, Hurd et al. (1998) added N in two forms at 

varying rates at three sites in the Adirondack Mountains 

of New York: (1) biweekly as dissolved nitric acid 

(HNO3) spray at the rate of 14 kg N ha-1 yr-1 (two times 

the ambient of 7 kg N ha-1 yr-1 at the lowest deposition 

site); and (2) single applications of ammonium sulfate at 

14 and 28 kg N ha-1 yr -1 (two and four times ambient). 

Th ey found that cover of prominent herbaceous 

species declined signifi cantly after only 3 years of 

treatment, particularly in response to ammonium sulfate 

((NH4)2SO4). Th is decline resulted from increased 

shading by ferns, which can substantially reduce light 

levels to the forest fl oor. Th e response to N additions 

was more pronounced at sites experiencing lower 

ambient inputs of atmospheric N.

7.3.4 Mycorrhizal Fungi

Ectomycorrhizal fungi. Lilleskov et al. (2008) found 

that community structure of ectomycorrhizal fungi 

changed over a regional N deposition gradient from 

northeastern New York to Maine. Frequency of the 

diff erent morphotypes varied continuously with 

increasing root N, and root N was positively related to 

N deposition. Wet deposition at the low N end of this 

gradient was estimated at approximately 3 kg N ha-1

yr-1; total N deposition would be approximately 4 kg 

N ha-1 yr-1 (Ollinger et al. 1993). Th e estimated level 

of N deposition resulting in community change is 5 to 

7 kg N ha-1 yr-1. Th is is quite similar to patterns found 

in white spruce forests in Alaska (Lilleskov et al. 2001, 

2002) and pitch pine forests in the Pine Barrens of New 

Jersey (Dighton et al. 2004).

If elevated soil NO3

-

 availability is a good indicator 

of ectomycorrhizal fungal community change in 

oligotrophic N-limited conifer forests (Lilleskov et al. 

2002, Lilleskov 2005), then N deposition thresholds 

that minimize excess NO3

- availability should be 

suffi cient to protect ectomycorrhizal fungal diversity in 

these forest types.

Arbuscular mycorrhizal fungi. We have less empirical data 

on arbuscular mycorrhizal community response to N in 

northern forests. One study in sugar-maple-dominated 

forests in Michigan has found declines of arbuscular 

mycorrhizal fungal biomass (van Diepen et al. 2007, 

2010) and changes in community structure (van Diepen 

2008) in response to N addition of 30 kg ha-1 yr-1 above 

background deposition of approximately 5 to 12 kg ha-1

yr-1 wet deposition, but the lower threshold of response 

to N is not yet known. Th ere is some indication 

that arbuscular mycorrhizal biomass has declined 

and community composition has shifted over the N 

deposition gradient (van Diepen 2008), suggesting a 

threshold of <12 kg ha-1 yr-1 of wet deposition, but it is 

likely that there are interactions with within-site factors, 

and defi nitive determination of a response threshold 

awaits further analysis.

7.3.5 Lichens and Bryophytes

Lichen and bryophyte responses documented in the 

Taiga and Eastern Temperate Forests ecoregions sections 

can also be expected in the U.S. portion of the Northern 

Forests ecoregion (see Chapters 6 and 10). Nearly all 

northern forest lichens belong either to the boreal or 

to the eastern fl ora (Brodo et al. 2001). However, air 

quality in this ecoregion is much more similar to the 

densely populated Eastern Forests ecoregion than it is 

to the relatively pristine Taiga ecoregion (see Chapter 3, 
Chapter 7—Northern Forest GTR-NRS-80 67

air quality for the Taiga ecoregion). Responses to 

combined sulfur dioxide (SO2

-

), acidity, and enhanced 

N deposition in western Pennsylvania (Showman and 

Long 1992) and New York (Smiley and George 1974) 

have included dramatic declines in species diversity and 

a shifting of community composition favoring acidand N-tolerant species. Information on lichen research 

methodology and terminology can be found in Chapter 

4 of this report.

7.4 Critical Loads Estimates

Empirical critical loads for both hardwood and 

coniferous ecosystems within the Northern Forest 

ecoregion can be set at 8 kg ha-1 yr-1 (Table 7.2), the 

observed threshold for surface water NO3

-

 leaching 

from these forests (Aber et al. 2003). Th is estimate is 

very reliable because it is based on 354 data points from 

sites around the northeastern United States (Aber et 

al. 2003). Rather than threshold responses, foliar N 

concentration and forest growth and survivorship all 

tend to show gradual responses with variation in N 

deposition received in the northern United States, with 

responses that often vary by species. Some sensitive 

tree species show declines in growth (red pine) or 

survivorship (yellow birch, scarlet and chestnut oak, 

quaking aspen, and basswood) as deposition increases 

across the Northeast from 3 to 11 kg N ha-1 yr-1

(Th omas et al. 2010); the critical load for negative tree 

growth response to N deposition is >3 kg ha-1 yr-1. 

Th is value is fairly reliable, based on measurements 

from thousands of trees of each species. Based on the 

N-addition study in montane spruce-fi r systems at Mt. 

Ascutney, adverse growth and mortality responses are 

likely to occur at deposition > 10 but <26 kg N ha-1 yr-1

Table 7.2—Critical loads of nutrient N for the Northern Forests ecoregion. Reliability rating: ## reliable; 

# fairly reliable; (#) expert judgment

Ecosystem 

component

Critical 

load for N 

deposition

kg ha-1 yr-1

Reliability Response Comments Study

Trees >3 # Decreased growth of 

red pine, and decreased 

survivorship of yellow 

birch, scarlet and chestnut 

oak, quaking aspen, and 

basswood

Thomas et al. 2010

Lichens 4-6 (#) Community composition 

shift

Application of model 

developed for Marine 

West Coast Forests to 

Northern Forests

Geiser et al. 2010

Ectomycorrhizal 

fungi

5-7 # Change in fungal 

community structure

Lilleskov et al. 2008

Herbaceous 

species cover

>7 and <21 # Loss of prominent species Response observed in 

low-level fertilization 

experiment

Hurd et al. 1998

Northern 

Hardwood and 

Coniferous Forests

8 ## increased surface water 

nitrate leaching

Aber et al. 2003

Tree growth and 

mortality

>10 and <26 # Decreased growth and/or 

induced mortality

Response observed in 

low-level fertilization 

experiment in old-growth 

montane red spruce

McNulty et al. 2005

Arbuscular 

mycorrhizal fungi

<12 (#) biomass decline and 

community composition 

change

Observed along a 

Michigan N gradient

van Diepen 2008
68 Chapter 7—Northern Forest GTR-NRS-80

(McNulty et al. 2005). Although it only occurred at a 

single site, this striking and clear response of induced 

mortality is considered fairly reliable. Similarly, based 

on an N addition study in the Adirondacks, understory 

herbaceous composition is likely to be altered at 

deposition >7 kg N ha-1 yr-1 but < 21 kg N ha-1 yr-1

(Hurd et al. 1998). Because this response was observed 

at a single site, this value is considered fairly reliable. 

Ectomycorrhizal fungal-based critical loads appear to 

be about 5 to 7 kg N ha-1 yr-1 (Lilleskov et al. 2008); 

the upper end of this range converges with the critical 

load based on surface water NO3

-

 leaching reported 

above. Arbuscular mycorrhizal fungal critical loads are 

tentatively set at <12 kg N ha-1 yr-1 (van Diepen 2008). 

Th e value for ectomycorrhizal fungi can be considered 

fairly reliable, as it is based on the convergence of 

several studies in this and ecologically similar regions. 

Th e value for arbuscular mycorrhizal fungi is considered 

expert judgment as it is based on one gradient with only 

four sites.

Application of a model relating epiphytic lichen 

community composition of Oregon and Washington 

coniferous forests to N deposition, accounting for 

precipitation (Chapter 4; Geiser et al. 2010), to the 

Northern Forests ecoregion yields a critical load 

estimate of 4 to 6 kg ha-1 yr-1. Th is value was calculated 

using a precipitation range of 100 to 240 cm modeled 

by PRISM for the U.S. portion of northern forests, 

and a lichen community-composition response 

threshold that allows as few as 30 percent oligotrophs 

and no more than 34 percent eutrophs. Th is is the 

same threshold that was used for western Oregon and 

Washington coniferous forests by Geiser et al. (2010). 

Because it relies on data from other ecoregions, the 

value for lichens is considered expert judgment.

7.5 Comparison to European Critical 

Loads

Critical loads for comparable forest ecosystems in 

Europe are reported to range from 10 to 15 kg N ha-1

yr-1 for boreal and temperate forest, based largely on 

changes in soil processes and NO3

-

 leaching (Bobbink 

et al. 2003). Critical loads for expected changes in 

foliar N:P and N:Mg ratios were set at 15 to 20 kg N 

ha-1 yr-1; no critical load was set for growth (Bobbink 

et al. 2003). For herbaceous species in boreal forests, 

the critical load may be as low as 7 kg N ha-1 yr-1. For 

ectomycorrhizal fungi, Bobbink et al. (2003) stated an 

expert judgment for critical loads of 10 to 20 kg N ha-1

yr-1. For lichens and epiphytic algae, the critical load is 

likely not higher than 10 to 15 kg N ha-1 yr-1 (Bobbink 

et al. 2003).

7.6 Future Research Directions and 

Gaps in Data

Fundamental gaps in understanding the regulation 

of N availability in forest ecosystems remain. Hence, 

simple relationships between inputs and ecosystem 

outputs and damage have been diffi cult to determine. 

Other factors that may be important include climate 

events (disturbance and variation), prior land use, 

species composition, site characteristics, and hydrology. 

Interactions between diff erent stressors add complexity 

to the input/output relationship. Low-level fertilization 

induced substantial mortality in a montane sprucefi r ecosystem at Mt. Ascutney (McNulty et al. 2005), 

but it is not clear to what extent that research can 

be extrapolated to the rest of the region, especially 

to lowland forests. Determining whether this site is 

unique would be useful to understanding the potential 

consequences of elevated N deposition in the Northern 

Forest ecoregion.

Far less research exists for herbaceous vegetation than 

for trees. Future research might include additional 

low-level fertilizations with an emphasis on examining 

the responses of herbaceous species, mycorrhizal 

fungi, bryophytes (especially in bog ecosystems), and 

lichens. In addition, diversity and community structure 

could be included in the parameters measured. Better 

quantifi cation of N deposition at high elevations would 

also help improve understanding of the susceptibility of 

the ecosystems and the thresholds when responses occur. 

Finally, the results of this analysis may be useful in 

improving understanding of the value chosen for critical 

N concentrations or leaching values in the steady-state 

mass balance method for calculating critical loads.
Chapter 7—Northern Forest GTR-NRS-80 69

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Chapter 7—Northern Forest GTR-NRS-80 73

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74
Chapter 8—Northwestern Forested Mountains GTR-NRS-80 75

8 NORTHWESTERN FORESTED MOUNTAINS

W.D. Bowman, J.S. Baron, L.H. Geiser, M.E. Fenn, E.A. Lilleskov

8.2 Ecosystem Responses to N 

Deposition

Th e wide diversity of communities and climates, along 

with diverse bedrock geologies, results in substantial 

variation in ecological responses to nitrogen (N) 

deposition. Unfortunately, little information exists on 

the response of many areas in the region. Documented 

responses to elevated N deposition within this bioregion 

include alteration of soils (carbon:nitrogen (C:N) ratios, 

base cation composition, and N cycling rates, including 

mineralization and nitrifi cation), plant, lichen, and 

algal chemistry (N concentration and nitrogen:calcium 

(N:Ca), nitrogen:magnesium (N:Mg), nitrogen: 

phosphorus (N:P), and C:N ratios; lichen thallus N 

concentrations), surface water N concentration and 

acid-neutralizing capacity, catchment N leaching rate, 

and changes in the community composition of plants, 

lichens, and phytoplankton. Th e region provides several 

important ecosystem services that may be compromised 

by N deposition, including water supply for much of 

western North America, timber, forage for livestock, and 

recreation.

8.3 Range of Responses Observed

8.3.1 Forest

Much of the research on ecological responses to N 

deposition in this ecoregion has been conducted in 

the Colorado Front Range of the Rocky Mountains, 

where a strong connection between regional urban 

and agricultural emissions and adjacent wildlands 

responses has been documented (e.g., Rocky Mountain 

National Park; Baron et al. 2000, Gebhart et al. 2011). 

Responses to N deposition are shown in Table 8.1. 

Rueth and Baron (2002) evaluated characteristics 

of subalpine Engelmann spruce stands spanning the 

Continental Divide across a N deposition gradient 

from 1-2 to 3-5 kg ha-1 yr-1. Deposition reported is 

total (wet + dry) at the eastern end of the deposition 

8.1 Ecoregion Description

Th e Northwestern Forested Mountains are ecologically 

diverse and geographically widespread, encompassing 

the mountain ecosystems of central and northwestern 

North America (CEC 1997; Figure 2.2). Th e 

ecoregion description is adapted from CEC (1997).

Geographically, they extend from the Rocky Mountains 

and the Sierra Nevada north through the Siskiyous, 

the east side of the Cascade Range, and then east of the 

Coast Ranges to interior Alaska. Climatically, the region 

is characterized by a transition from a moist, maritime 

climate in the northwest, to a continental and drier 

climate in the Rockies in the southeast. Orographically 

generated rainfall creates both rain shadows and wet 

belts, often in close proximity.

Th e vegetation of the ecoregion is extremely diverse, 

with distinct community zonation occurring along 

elevation gradients. Alpine communities at the highest 

elevations contain various forb, lichen, and shrub 

associations. Subalpine communities include lodgepole 

pine (Pinus contorta), subalpine fi r (Abies lasiocarpa), 

Pacifi c silver fi r (Abies amabilis), grand fi r (Abies 

grandis), and Engelmann spruce (Picea engelmannii). 

Mid-elevation forests are characterized by ponderosa 

pine (Pinus ponderosa), Rocky Mountain Douglas-fi r 

(Pseudotsuga menziesii var. glauca), lodgepole pine, and 

quaking aspen (Populus tremuloides) in the east, and by 

western hemlock (Tsuga heterophylla), western red cedar 

(Th uja plicata), Douglas-fi r (Pseudotsuga menziesii), 

and western white pine (Pinus monticola) in the west 

and southwest. White and black spruce (Picea glauca 

and P. mariana) dominate the Alaskan portion of the 

ecoregion. Vegetation of the interior valleys in the 

southern portion of the region includes big sagebrush 

(Artemisia tridentata), rabbitbrush (Chrysothamnus spp.), 

and antelope bitterbrush (Purshia tridentata).
76 Chapter 8—Northwestern Forested Mountains GTR-NRS-80 Table 8.1—Observed and modeled responses to N inputs for the Northwestern Forested Mountains ecoregion. Reliability rating: ## reliable; # fairly reliable; (#) expert judgment Ecosystem Site N deposition kg ha-1 yr-1 Reliability Response Comments Study Coniferous forest Alaska 1.2-3.7 (#) Lichen community composition Application of western Oregon and Washington model Geiser et al. 2010 Alpine lake S. Rockies/Loch Vale RMNP 1.5 ## Diatom assemblages Wet deposition estimate Baron 2006 Coniferous forest Northwestern Forested Mountains, non-Alaska 2.5-7.1 ## Lichen community composition Application of western Oregon and Washington model, Sierra thresholds Geiser et al. 2010 Alpine S. Rockies/ Rocky Mountain National Park 3-4 (#) Soil chemistry Baron et al. 1994 Coniferous forest Sierra Nevada 3.1-5.2 ## Lichen community composition N measured as canopy throughfall Fenn et al. 2008 Subalpine Forest S.Rockies/12 stands of paired high and low N deposition <4 ## Foliar chemistry, mineralization, nitrifi cation, initial increases in N leaching below the organic layer Effects are cumulative over time Rueth and Baron 2002 Subalpine Forest S. Rockies/fertilized plots in Fraser and Loch Vale <4 ## Foliar chemistry, mineralization, nitrifi cation Effects are cumulative over time Rueth et al. 2003 Alpine vegetation S. Rockies/ Niwot Ridge 4 ## Species composition ; individual species responses Bowman et al. 2006 Alpine lake S. Rockies/Niwot 4.0 # Episodic freshwater acidifi cation Williams and Tonnesson 2000

Hardwood and coniferous forests

Columbia River Gorge 4-5 # Lichen community composition N measured as canopy 

throughfall

Geiser et al. in 2010

Alpine stream S. Rockies/Loch Vale 

RMNP

5 (#) Episodic freshwater 

acidifi cation

Baron and Hartmana, 

model results

Alpine and subalpine lakes and streams

Sierra Nevada > 5 ## NO3- leaching Increased NO3-leaching 

not observed at the highest deposition (5 kg ha-1 yr-1)

Sickman et al. 2001
Chapter 8—Northwestern Forested Mountains GTR-NRS-80 77

Alpine stream S. Rockies/Loch Vale 

RMNP

7.3 (#) Chronic freshwater acidifi cation Baron and Hartmana, 

model results

Alpine stream S. Rockies/Loch Vale 

RMNP

8 (#) Freshwater acidifi cation ANC=20 μeq/L Sullivan et al. 2005 

(modeled)

Alpine vegetation

S. Rockies/ Niwot Ridge 10 ## Species composition; whole 

community

Bowman et al. 2006

Subalpine stream/lake

S. Rockies/Loch Vale RMNP

12 (#) Freshwater acidifi cation ANC=20 μeq/L Sullivan et al. 2005 

(modeled)

Alpine stream S. Rockies/Loch Vale 

RMNP

14 (#) Freshwater acidifi cation ANC=0 μ/L Sullivan et al. 2005 

(modeled)

Mixed conifer forest

Sierra Nevada 17 ## NO3

- leaching; reduced fi ne 

root biomass

Critical load not exceeded in NW forests; critical load based on SW Sierra Nevada and San Bernardino Mts

Fenn et al. 2008

Alpine terrestrial

S. Rockies / Niwot Ridge >20 # NO3

- leaching, soil nitrogen 

fl uxes

Bowman et al. 2006

Subalpine stream/lake

S. Rockies/Loch Vale RMNP

21 (#) Freshwater acidifi cation ANC=0 μeq/L Sullivan et al. 2005 

(modeled)

aBaron, J.S.; Hartman, M.D. Unpublished data. On fi le with Jill Baron, Research ecologist, U.S. Geological Survey, Natural Resource Ecology Laboratory, Colorado State University, 

Fort Collins CO 80523-1499
78 Chapter 8—Northwestern Forested Mountains GTR-NRS-80

gradient; wet deposition only is reported in the west 

where dry deposition was assumed to be insignifi cant 

(Rueth and Baron 2002). At the eastern, high end of 

the deposition gradient, forest stands had signifi cantly 

lower organic soil horizon C:N, higher N mineralization 

rates, higher potential net nitrifi cation rates and foliar N 

concentration, and higher N:P, N:Ca, and N:Mg ratios. 

Nitrate (NO3

-

) leaching was elevated at the eastern, 

high end of the deposition gradient (Baron6

, Rueth and 

Baron 2002). An earlier study reported elevated foliar 

N:P ratios in high elevation bristlecone pine (Pinus

aristata) stands receiving N deposition of 6 kg N ha-1

yr-1 (Williams et al. 1996).

Four years of low-level (25 kg ha-1 yr-1) ammonium 

nitrate (NH4NO3) fertilization to old-growth coniferous 

forests with diff erent initial levels of atmospheric N 

deposition showed diff erent responses based on initial 

conditions (Rueth et al. 2003). Th e low deposition 

site with initial organic horizon C:N of 36 and N pool 

of 605 kg N ha-1 showed no signifi cant increase in N 

mineralization rates. At this low deposition site, foliar 

and soil organic horizon C:N decreased signifi cantly 

with fertilization. In contrast, N mineralization rates and 

extractable inorganic N in soil increased signifi cantly 

at the higher N deposition site with greater initial 

soil N (C:N of 24, N pool of 991 kg N ha-1). At this 

higher N deposition site, foliar and soil organic horizon 

percentages of N did not change (Rueth et al. 2003).

In mixed conifer forests of the southwestern Sierra 

Nevada (See Chapter 13, Mediterranean California), 

NO3

-

 leaching increases at a deposition of 17 kg N 

ha-1 yr-1, as determined from empirical data and from 

biogeochemical modeling (Fenn et al. 2008). At this 

level of N deposition, fi ne root biomass was reduced 

by 26 percent. Th e highest throughfall N deposition 

inputs reported for the northern and eastern regions 

of the Sierra Nevada range are 7 kg ha-1 yr-1 (Fenn et 

al. 2008). Mixed conifer forests in California with 

these levels of N deposition are highly conservative of 

N, with no appreciable hydrologic NO3

-

 leaching or 

gaseous N losses from soil (Fenn et al. 2008). Eff ects 

on soils and vegetation could be occurring in the most 

polluted northern and eastern regions, but are likely to 

be subtle and may not have been observed for lack of 

in-depth study. However, Community Multi-scale Air 

Quality (CMAQ) simulations of N deposition indicate 

that N inputs may be as high as 11 to 17 kg ha-1 yr-1

along portions of the western edge of the central and 

northern Sierra Nevada range (Fenn et al. 2010). If 

deposition actually occurs at these levels in some sites, 

N enrichment of soil and plants are expected. Increased 

NO3

-

 leaching and gaseous N loss may also occur at the 

upper end of this deposition range (Fenn et al. 2008). 

Eff ects of N deposition on biodiversity of understory or 

shrub communities in the Sierra Nevada range have not 

been studied, although plant community composition 

of many nutrient-poor low biomass ecosystems in 

California are likely impacted by deposition levels 

≤10 kg ha-1 yr-1 (Weiss 2006). Weiss (2006) estimated 

that 10,000 km2

 in California are vulnerable to plant 

community impacts from N deposition >10 kg ha-1 yr-1.

Th e most notable eff ect of N deposition in the low 

to moderately polluted forested regions of the Sierra 

Nevada range (N deposition of 3 to 7 kg ha-1 yr-1) is 

the highly signifi cant alteration of lichen communities 

(Fenn et al. 2008).

Ectomycorrhizal fungi. Th ere is no information on 

ectomycorrhizal community responses to N in the 

Northwestern Forested Mountains ecoregion. We 

postulate that, given the climate, species composition 

and soils, white and black spruce dominated ecosystems 

may have similar deposition thresholds to those seen 

in spruce forests of the same species in coastal Alaska 

(see Chapter 9, West Coast Marine Forests). Similarly, 

deposition thresholds for Engelmann spruce-subalpine 

fi r dominated ecosystems may be similar to those of the 

spruce-fi r forests of the northeastern United States (see 

Chapter 7, Northern Forests).

Forest Lichen Communities. In the California Sierra 

Nevada range, a strong ammonia (NH3

) deposition 

gradient was positively correlated with NH3 indicator 

lichen species, or eutrophs (r = 0.93). Th e greatest lichen 

6

Baron, J.S. Unpublished data. Research ecologist, U.S. 

Geological Survey, Natural Resource Ecology Laboratory, 

Colorado State University, Fort Collins CO 80523-1499
Chapter 8—Northwestern Forested Mountains GTR-NRS-80 79

community impacts were observed in the portions of 

the Sierra Nevada range in the southwest and near the 

Tahoe Basin, and the Modoc Plateau, near urban areas 

and at popular recreation areas in Sequoia and Yosemite 

National Parks (Jovan and McCune 2006). Increased 

N content of the lichen epiphyte Letharia vulpina, as 

defi ned by a threshold concentration of 1.0 percent, was 

also observed in these areas (Jovan and Carlberg 2006). 

Th allus N concentrations above this threshold are 

correlated with decreasing representation of oligotrophic 

species relative to mesotrophic and eutrophic species 

within epiphytic macrolichen communities of the Sierra 

Nevada (Fenn et al. 2008). Th allus N concentrations 

below the threshold represent the clean site range; 

at these sites no community shifts are detected. 

Information on lichen research methodology and 

terminology can be found in Chapter 4 of this volume.

Within the mixed conifer forests in California, 

which extend from the west side of the California 

Sierra Nevada from Tahoe National Forest south 

through Sequoia-Kings Canyon National Park, simple 

indices of three N-indicator groups—the oligotrophs 

(adapted to low nutrient availability), the mesotrophs 

(requiring moderate nutrient levels and tolerating some 

anthropogenic enhancement but not extreme levels of N 

deposition) and the eutrophs (nitrophilous species that 

respond favorably to enhanced N deposition)—were 

used to identify N loads that correspond with major 

shifts in Sierra Nevada lichen communities. At 3.1 

kg N ha-1 yr-1 (95 percent confi dence interval: 0.70 to 

5.5 kg ha-1 yr-1; Fenn et al. 2008; see also Table 8.2), 

the N concentration threshold in the lichen Letharia 

vulpina was exceeded, and the lichen community 

composition was already shifting from oligotrophic to 

more N-tolerant mesotrophic and eutrophic species. 

At an estimated N deposition of 5.7 kg ha-1 yr-1 the 

lichen community had shifted from the natural state 

of oligotroph dominance. Th is change is of particular 

concern because oligotrophic species are important 

components of winter food webs, hydrologic and 

nutrient cycles, and wildlife habitat (McCune et al. 

2007, McCune and Geiser 2009). Th e data from this 

study predict a complete extirpation of oligotrophs 

from the lichen community at an N load of about 10.2 

kg N ha-1 yr-1. Th is work demonstrates that known 

biological impacts are occurring at N deposition levels 

as low as 3 to 5 kg ha-1 yr-1, levels which are exceeded 

over large areas of the forests of California: most of the 

forested region of the San Bernardino and San Gabriel 

Mountains, the western portions of the Sierra Nevada, 

and large portions of other mountain ranges in southern 

California (Fenn et al. 2003a, b; Fenn et al. 2008, Fenn 

et al. 2010).

Geiser et al. (2010) applied a model developed for 

Oregon and Washington forests to Sierra Nevada 

lichen communities, using the same response thresholds 

as Fenn et al. (2008) and median site precipitation 

(111 cm). At 2.5 to 3.8 kg N ha-1 yr-1, some sensitive 

oligotrophic species were modeled to be absent; at 4.6 

to 5.7 kg N ha-1 yr-1, the lichen community would shift 

away from oligotroph dominance. Th ese results are 

very close to the results obtained by Fenn et al. (2008). 

Fenn et al. (2008) did not account for precipitation in 

the Sierra Nevada study, as recommended by Geiser 

et al. (2010) for Oregon and Washington forests. 

Because mean annual precipitation across Sierra Nevada 

study sites varied only twofold (79 to 165 cm), it was 

probably less infl uential there than in Oregon and 

Washington, where precipitation varied tenfold. If 

the entire precipitation range in the ecoregion were 

considered, the Geiser et al. 2010 model predicts a 

lichen response threshold of 2.5 to 7.1 kg N ha-1 yr-1 for 

the Northwestern Forested Mountains ecoregion in the 

contiguous United States(30 to 203 cm), and 1.2 to 3.7 

kg N ha-1 yr-1 (30 to 80 cm) for Alaska (See Table 4.1 

for details).

Along the Snake River and its tributaries in Idaho and 

Oregon, within Hells Canyon National Recreation 

Area, comparatively high NH4

NO3 in fi ne particulates 

originating from the agriculturally intensive Snake River 

basin accounted for increased lichen N throughout the 

Recreation Area relative to other remote sites in Oregon, 

Washington, Wyoming, and Colorado. Average seasonal 

NH3 concentrations between 4.2 and 7.7 parts per 

billion (ppb), with 2-week averages peaking at 9 to 18 

ppb, were associated with signifi cantly higher cover of 

the nitrophilous lichens Xanthomendoza and Xanthoria, 
80 Chapter 8—Northwestern Forested Mountains GTR-NRS-80 Table 8.2—Responses to N input for lichens of the Northwestern Forested Mountains ecoregion. Ecosystem Site N levels generating observed response Measured response Comments Study Coniferous forest Central-southern Sierra Nevada 3.1 kg N ha-1 yr-1 Some sensitive lichen species absent Critical load based on exceedance of a N concentration threshold in the lichen Letharia vulpina Fenn et al. 2008 Coniferous forest Central-southern Sierra Nevada 2.5-3.8 kg N ha-1 yr-1 Some sensitive lichen species absent Application of Oregon/ Washington model using Fenn et al. 2008 response threshold and median precipitation Geiser et al. 2010 Coniferous forests Sierra Nevada (Tahoe) 4-5 kg N ha-1 yr-1 Lichen N >1.03% 

(Letharia vulpina).

In the SW Sierras, Tahoe and Modoc Plateau

Jovan and Carlberg 2006, Jovan and McCune 2006 

Coniferous forest

Central-southern Sierra Nevada

5.7 kg N ha-1 yr-1 50 % sensitive lichen 

species absent

Critical load based on exceedance of a N concentration threshold in the lichen Letharia vulpina

Fenn et al. 2008

Coniferous forest

Central-southern Sierra Nevada

4.6-5.7 kg N ha-1 yr-1 50 % sensitive lichen 

species absent

Application of Oregon/Washington model using Fenn et al. 2008 response threshold and median precipitation

Geiser et al. 2010

Coniferousforests

Sierra Nevada (Sequoia NP)

6-15 kg N ha-1 yr-1 Lichen N > 1.03% 

(Letharia vulpina).

N was mostly NH3; N content was elevated in the SW Sierras, Tahoe and Modoc Plateau

Jovan and Carlberg 2006, Jovan and McCune 2006 

Coniferous forest

Central-southern Sierra Nevada

10.2 kg N ha-1 yr-1 Extirpation of all sensitive 

lichen species

Critical load based on exceedance of a N concentration threshold in the lichen Letharia vulpina

Fenn et al. 2008

Coniferous forest

Columbia River Gorge OR/WA

<11.5 kg ha-1 in 4.5 

months

Loss of sensitive species; lichen N >1.5% (Xanthoparmelia cumberlandia). 

Lichen N and measured N in throughfall were correlated. Lichen N concentration exceeded the threshold at all sites

Fenn et al. 2007

Deciduous riparian forests

Hells Canyon NRA,Snake River OR/ID

4.2-7.7 μg NH3 m-3

annual mean; 

5-20 μg NH3 m-3 summer 

biweekly peaks

Increased cover of eutrophs; lichen N >1.5% (Xanthoparmelia cumberlandia)

4.2-7.7 μg NH3 m-3 annual 

mean; 5-20 μg NH3 m-3 summer 

biweekly peaks

Geiser et al. 2008
Chapter 8—Northwestern Forested Mountains GTR-NRS-80 81

and higher lichen N concentrations in Xanthoparmelia 

cumberlandia (Geiser et al. 2008). Similarly, abundance 

of the lichen epiphyte, Xanthomendoza fallax, in 

northern Utah and southern Idaho, increased with 

proximity to peak urban (Logan, Utah) and agricultural 

NH3 emission centers, where summertime NH3

ranged from 7.3 to 92.2 μg m-3. Much of the eff ect was 

attributed to long distance transport and deposition of 

NH4

+

 (Rogers et al. 2009).

Topographic and meteorological conditions make the 

Columbia River Gorge a conduit for N-containing 

pollutants generated by the Portland, Oregon to 

Vancouver, Washington metropolis to the west and 

the agriculturally intensive Columbia Basin to the east. 

Nitrogen pollution contributes to decreased visibility 

and highly acidic fog events and potentially impacts 

valuable historic petroglyphs (Fenn et al. 2007, Geiser 

et al. 2008). Lichen N and measured N in throughfall 

were correlated throughout the Columbia River 

Gorge, and elevated levels were measured, especially 

at eastern- and western-most sites. Lichens indicating 

N-enriched environments were more abundant and 

lichen N concentrations were two to three times higher 

in the Columbia River Gorge than surrounding national 

forests; throughfall N deposition ranged from 11.5 to 

25.4 kg ha-1 over 4.5 months (Fenn et al. 2007). Lichen 

N concentrations exceeded regional background ranges.

8.3.2 Alpine Terrestrial

Alpine ecosystems are particularly sensitive to increased 

availability of N due to inherently low rates of N 

cycling, low rates of primary production, and thin, 

poorly weathered soils (Fenn et al. 1998). Most of the 

studies examining the responses of alpine ecosystems 

come from the Colorado Front Range of the Rocky 

Mountains. Baron et al. (1994) used the CENTURY 

model to estimate N uptake by plants and soils. Th ey 

estimated that increases in the export of NO3

-

 began 

at lower inputs (3 to 4 kg ha-1 yr-1) in the alpine than 

adjacent subalpine ecosystems, which had larger 

biological sinks for N.

Bowman et al. (2006) used an experimental approach 

to examine N impacts on alpine vegetation. Noting 

that previous studies using high N inputs resulted in 

modest increases in production and large changes in 

species composition, they established experimental plots 

in dry meadow tundra with an ambient control of 6 

kg ha-1 yr-1 and experimental N additions of 20, 40, 

and 60 kg ha-1 yr-1. Diversity increased at all levels of N 

treatment relative to the control, while changes in net 

N mineralization, nitrifi cation, and leaching occurred 

at inputs greater than 20 kg N ha-1 yr-1 (Bowman et 

al. 2006). Nearly identical results were obtained from 

a similar study in Rocky Mountain National Park 

(Bowman and Murgel7

). Changes in plant species 

composition associated with N deposition have the 

potential to enhance rates of N cycling, leading to a 

nonlinear increase in N leaching and acidifi cation of 

soils (Bowman and Steltzer 1998).

8.3.3 Alpine Freshwater Lakes and Streams

Because many alpine freshwaters are strongly N-limited, 

they are highly responsive to even slight changes in 

nutrient availability (Elser et al. 2009a, b). Nutrient data 

collected from regional lake surveys and analyzed as part 

of the Surface Water Chapter (18) of this monograph 

indicate that 45 percent of Rocky Mountain lakes 

and 35 percent of Sierra Nevada/Cascade lakes were 

N limited in 1985 (see Table 18.1). Slight amounts 

of additional N from deposition or experiments have 

elicited an increase in algal biomass and changes in the 

composition of Rocky Mountain and Sierra Nevada 

lake species assemblages (Bergström and Jansson 2006; 

Goldman 1988; Interlandi and Kilham 1998; Lafrancois 

et al. 2003a, 2003b; Michel et al. 2006; Morris and 

Lewis 1988; Nydick et al. 2004). Lake sediment records, 

including algal diatoms, organic compounds, and stable 

isotopes, indicated that the onset of change due to N 

fertilization from atmospheric deposition occurred in 

the decade 1950-1960 (Das et al. 2005; Enders et al. 

2008; Wolfe et al. 2001, 2003).

Two species of diatom that are favored by high 

N availability, Asterionella formosa and Fragilaria 

crotonensis, now dominate the fl ora of at least several 

7

Bowman, W.D.; Murgel, J. Unpublished data on fi le with 

William Bowman, Professor, Department of Ecology and 

Evolutionary Biology, University of Colorado, Boulder, CO 

80309-0334
82 Chapter 8—Northwestern Forested Mountains GTR-NRS-80

alpine and montane Rocky Mountain lakes (Baron 

et al. 2000, Interlandi and Kilham 1998, Saros et al. 

2003, Saros et al. 2005, Wolfe et al. 2001, Wolfe et 

al. 2003). Other species have shown negative growth 

responses to increased N; two of these, Tetracyclus 

glans and Staurosirella pinnata, were reported by 

Michel et al. (2006). In studies of remains of diatoms 

in lake sediment, typical oligotrophic species such as 

Aulacoseria perglabra, Cyclotella steligera, and Achnanthes 

spp. declined coincident with the rise in dominance of 

A. formosa and F. crotonensis (Wolfe et al. 2001, 2003).

Several studies have noted higher surface water NO3

-

concentrations in the Colorado Front Range compared 

with other lakes of the Southern Rocky Mountains, 

especially lakes east of the Continental Divide (Baron 

et al. 2000, Elser et al. 2009b, Musselman et al. 2004). 

Another study reported acid neutralizing capacity 

(ANC) in surface water < 0 μmol/L as a result of 

acidifying inputs of N deposition, suggesting that 

current deposition levels are having an observable 

impact on catchments in the Front Range (Williams 

and Tonnesson 2000).

In contrast to the elevated NO3

-

 leaching observed with 

relatively low N deposition inputs in the Colorado 

Front Range, similar N deposition levels (2 to 5 kg 

ha-1 yr-1) in Sierra Nevada high elevation catchments 

have not been reported to increase NO3

-

 leaching or 

NO3

-

 levels in lakes. Even when atmospheric deposition 

is low, high elevation watersheds in the Sierra Nevada 

appear to naturally export low levels of NO3

-

 during 

early snowmelt, particularly in catchments with low 

amounts of soil cover (Fenn et al. 2003b, Sickman et 

al. 2001). Th e greater response of Rocky Mountain 

catchments to atmospheric deposition may be due to 

climatic diff erences. For example, the greater extent 

of soil freezing in the Rockies likely reduces the N 

retention capacity of these watersheds (Sickman et al. 

2002). However, the combination of a snowmelt NO3

-

pulse, coupled with a short growing season and limited 

soils and vegetation, results in temporal asynchrony 

between N availability and N demand, suggesting that 

high elevation ecosystems in the Sierra Nevada will 

respond quickly to increased N deposition (Fenn et al. 

2003b).

8.4 Critical Load Estimates

8.4.1 Forest

Th e empirical N critical load for subalpine forests 

of the Rocky Mountains is ≤ 4 kg N ha-1 yr-1, based 

on the incipient eff ects of N saturation (organic 

horizon and foliar N enrichment, higher potential 

net N mineralization rates, and NO3

-

 leaching; more 

information about N saturation can be found in 

Chapter 7) reported by Rueth and Baron (2002) 

and Baron.6

 We expect that more productive forest 

types with deeper soils in the Pacifi c Northwest, e.g., 

Douglas-fi r, would have higher critical load thresholds. 

A modeling study conducted for subalpine forests of 

the Colorado Front Range projected an increase in 

forest soil N leachate to streams at N deposition values 

> 4 kg ha-1 yr-1 (Baron et al. 1994). In mixed conifer 

forests of the Sierra Nevada, the empirical critical load 

for incipient NO3

-

 leaching is 17 kg ha-1 yr-1 (Fenn et al. 

2008). Critical loads are summarized in Table 8.3.

Critical loads for lichens ranged from 1.2 to 7.1 kg N 

ha-1 yr-1. Th e western Sierra Nevada critical loads range, 

3.1 to 5.2 kg N ha-1 yr-1 (Fenn et al. 2008), is considered 

reliable. Th e most protective response threshold was 

based on N concentrations above the threshold value 

in the lichen, Letharia vulpina, in mixed conifer forests 

of California (Fenn et al. 2008). At this level of N 

deposition, community composition of epiphytic 

macrolichens was already shifting from oligotrophic to 

more N-tolerant species. Th e higher threshold marked 

a shift to domination by eutrophic species. Using the 

Sierra Nevada (Fenn et al. 2008) response thresholds, 

a more conservative response threshold for Alaska 

(justifi ed in Table 4.1), and relevant precipitation 

ranges, the western Oregon and Washington model 

(Geiser et al. 2010) predicted critical loads of 2.5 to 7.1 

for the Northwestern Forested Mountains ecoregion in 

the contiguous United States and 1.2 to 3.7 kg N ha-1

yr-1 for Alaska.

Provisional N critical loads using ectomycorrhizal fungal 

abundance and diversity as indicators ranged from 5 

to 10 kg ha-1 yr-1, based on critical loads determined 

using the same indicators in similar forest types in the 

marine west coast forests and northern forests ecoregions 
Chapter 8—Northwestern Forested Mountains GTR-NRS-80 83

(Chapters 9 and 7; Lilleskov 1999; Lilleskov et al. 2001, 

2002, 2008).

8.4.2 Alpine Terrestrial

Using the rates of changes in species abundances, 

Bowman et al. (2006) estimated individual plant species 

responded to inputs at 4 kg ha-1 yr-1, while whole 

community responses were detectable at 10 kg N ha-1

yr-1. Estimates of critical load based on changes in net 

N mineralization, nitrifi cation, and NO3

-

 leaching were 

above 20 kg N ha-1 yr-1 (Bowman et al. 2006), higher 

than the 3 to 4 kg N ha-1 yr-1 predicted by Baron et al. 

(1994) using a modeling approach.

8.4.3 Alpine Freshwater Lakes and Streams

A hindcasting technique suggested the tipping point for 

change for aquatic diatoms occurred at about 1.5 kg N 

ha-1 yr-1 as wet deposition (Baron 2006). A meta-analysis 

of lakes across the northern hemisphere, including 

alpine lakes of the western United States, suggests lake 

productivity increases at deposition values of 2.5 kg N 

ha-1 yr-1 (Bergström and Jansson 2006). Williams and 

Tonnessen (2000) compared catchment yields of N with 

deposition in the Green Lakes Valley, and estimated 

a N critical load for acidifi cation of 4 kg ha-1 yr-1 wet 

deposition. Aquatic critical loads for alpine and other 

lakes are discussed in greater detail in Chapter 18.

8.5 Comparison to European Critical 

Loads

8.5.1 Forest

Northwestern forests can most easily be compared to the 

Scandinavian forests, which were the subject of a recent 

workshop (Nordin 2007). Using empirical results from 

Table 8.3—Empirical critical loads of nutrient N for the Northwestern Forested Mountains ecoregion. 

Reliability rating: ## reliable; # fairly reliable; (#) expert judgment

Ecosystem 

component

Critical load for 

N deposition 

kg ha-1 yr-1

Reliability Response Comments Study

Alpine lakes 1.5 ## Changes in diatom 

assemblages

Wet deposition estimate Baron 2006

Lichens 1.2-3.7 (#) Community change in mixedconifer forests in AlaskaApplication of western 

Oregon and Washington 

model

Geiser et al. 2010

2.5-7.1 ## Community change in mixedconifer forestsFenn et al. 2008, 

Geiser et al. 2010

Subalpine forest 4 ## Increase in organic horizon 

N, foliar N, potential net N 

mineralization, and soil solution 

N, initial increases in N leaching 

below the organic layer

Baron et al. 1994,

Rueth and Baron 

2002 

Alpine lakes 4.0 # Episodic freshwater acidifi cation Williams and 

Tonnesson 2000

Alpine vegetation 4-10 ## Plant species composition Bowman et al. 

2006

Ectomycorrhizal 

fungi

5-10 (#) Ectomycorrhizal fungi 

community structure in white, 

black, and Engelmann spruce 

forests

Expert judgment 

extrapolated from Marine 

West coast spruce and 

northern spruce-fi r forest

Lilleskov 1999; 

Lilleskov et al. 

2001, 2002, 2008

Mixed conifer 

forest

17 ## NO3

-

 leaching Critical load based on 

SW Sierra Nevada and 

San Bernardino Mts

Fenn et al. 2008

17 # Reduced fi ne root biomass Critical load based on 

SW Sierra Nevada and 

San Bernardino Mts

Fenn et al. 2008
84 Chapter 8—Northwestern Forested Mountains GTR-NRS-80

N fertilization experiments, a critical load for changes in 

understory plant species composition of less than 10 kg 

N ha-1 yr-1 was suggested for forest ecosystems, while for 

more sensitive parts of the ecosystem, such as bogs and 

poor minerotrophic mires, a critical load of less than 8 

kg ha-1 yr-1 was suggested (Nordin 2007). Bobbink et al. 

(2003) recommended a critical load of 10 to 15 kg N 

ha-1 yr-1 for forests based on increased N mineralization 

and nitrifi cation.

8.5.2 Alpine Terrestrial

Th e synthesis by Bobbink et al. (2003) included 

estimates of critical loads for alpine ecosystems based 

on vegetation responses in fertilization experiments. 

For ecosystems on soils derived from granitic parent 

material, an estimated critical load of 10 kg N ha-1 yr-1

was reported, while for soils on basic parent material, 

a critical load of 15 to 20 kg N ha-1 yr-1 was estimated. 

Th ese estimates are 2 to 5 times higher than those 

estimated for alpine sites in the southern Rockies 

discussed above.

8.6 Future Research Directions and 

Gaps in Data

Improved methods for quantifying N deposition 

and increased eff orts to monitor trends over time are 

critically important for the Northwestern Forested 

Mountains ecoregion (Fenn et al. 2009), especially 

considering the projected increase in atmospheric N 

deposition to the region. Given the highly variable 

terrain, orographic eff ects, diffi culties of measuring 

winter deposition in alpine sites, potentially high inputs 

of dry deposition (particularly to coniferous stands), 

and isolated or concentrated emission source areas, 

deposition rates are often underestimated. As a result, 

it is diffi cult to extrapolate deposition rates, even to 

areas that are physically close to each other but diff erent 

in climate, elevation, or vegetation. Using wet-only 

estimates for deposition, and the problems associated 

with underestimating current deposition rates, may 

result in setting critical loads too low. In contrast, 

empirical critical loads are often based on linking 

observed responses with shorter-term N additions and 

the time lag in ecological responses to N deposition 

could lead us to set critical loads too high. Th e results 

of long-term fertilization studies would improve 

understanding of the rate of change possible in these 

high elevation ecosystems. Eff ects of N deposition 

on plant biodiversity and mycorrhizae have not been 

studied over most of the Northwestern Forested 

Mountains ecoregion. Plant biodiversity often responds 

to low to moderate levels of chronic N deposition. Such 

sensitive indicators will be most valuable, considering 

the low to moderate N deposition levels that occur over 

most of this region.

Far more is known for the more heavily impacted 

Colorado Front Range than for other parts of the 

Rocky Mountains. Additional studies are needed to 

better understand the sensitivity of alpine ecosystems 

in other parts of the west, particularly in the Teton 

and Wind River Ranges, which may have elevated 

N deposition due to agricultural activities and oil 

and gas extraction. Existing plant and lichen data 

collected systematically by the U.S. Forest Service 

Forest Inventory and Analysis program in California, 

Oregon, Washington, Idaho, Montana, Wyoming, 

Utah, Colorado, and New Mexico can potentially 

provide further insights into regional critical loads if 

community composition responses can be correlated 

with N deposition estimates.

Fire is an essential component of western forests, 

and many stands have suff ered severe dieback due 

to bark beetle irruptions. More research is needed 

on the interactions between fi re, bark beetle attack, 

and atmospheric deposition (Grulke et al. 2009). 

For freshwaters, surveys and experiments are needed 

to determine the eff ects of N deposition on algal 

biodiversity in N-limited lakes. No studies to date 

have adequately addressed this question. Finally, 

climate change will aff ect every aspect of the N cycle, 

from emissions and agricultural practices to climatemediated ecological and biogeochemical N response 

rates. Modeling and empirical studies should be used 

to address whether the already low estimates of critical 

loads for the Northwestern Forested Mountains 

ecoregion will need to be adjusted to changes in 

temperature and precipitation regimes.
Chapter 8—Northwestern Forested Mountains GTR-NRS-80 85

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Canadian Journal of Fisheries and Aquatic Science. 

62: 1681-1689.

Sickman, J.O.; Melack, J.M.; Stoddard, J.L. 2002. 

Regional analysis of nitrogen yield and retention 

in high-elevation ecosystems of the Sierra Nevada 

and Rocky Mountains. Biogeochemistry. 57: 341-

374.

Sickman, J.O.; Leydecker, A.; Melack, J.M. 2001. 

Nitrogen mass balances and abiotic controls on N 

retention and yield in high-elevation catchments 

of the Sierra Nevada, California, United States.

Water Resources Research. 337: 1445-1461.

Sullivan, T.J.; Cosby, B.J.; Tonnessen, K.A; Clow, 

D.W. 2005. Surface water acidifi cation responses 

and critical loads of sulfur and nitrogen deposition 

in Loch Vale watershed, Colorado. Water Resources 

Research. 41: WO1021.

Weiss, S.B. 2006. Impacts of nitrogen deposition on 

California ecosystems and biodiversity. California 

Energy Commission, PIER Energy-Related 

Environmental Research. CEC-500-2005-165. Santa 

Barbara, CA: University of California. 91 p.

Williams, M.W.; Baron, J.S.; Caine, N.; Sommerfeld, 

R.; Sanford, R., Jr. 1996. Nitrogen saturation in 

the Rocky Mountains. Environmental Science and 

Technology. 30(2): 640-646.

Williams, M. W.; Tonnessen, K.A. 2000. Critical loads 

for inorganic nitrogen deposition in the Colorado 

Front Range, USA. Ecological Applications. 10: 

1648-1665.

Wolfe, A.P.; Baron, J.S.; Cornett, R.J. 2001. 

Anthropogenic nitrogen deposition induces rapid 

ecological changes in alpine lakes of the Colorado 

Front Range (USA). Journal of Paleolimnology. 25: 

1-7.

Wolfe, A.P.; Van Gorp, A.C.; Baron, J.S. 2003. 

Recent ecological and biogeochemical changes in 

alpine lakes of Rocky Mountain National Park 

(Colorado, USA): a response to anthropogenic 

nitrogen deposition. Geobiology. 1(2): 153-168.
Chapter 9—Marine West Coast Forest GTR-NRS-80 89

9 MARINE WEST COAST FORESTS

S.S. Perakis, L.H. Geiser, E.A. Lilleskov

9.2 Ecosystem Responses to N 

Deposition

Data are sparse on Marine West Coast Forests ecosystem 

responses to N deposition, with information available 

on lichen responses in Oregon, Washington, and 

southeastern Alaska, and for plant, soil and mycorrhizal 

responses in white spruce dominated ecosystems in 

south-central Alaska. Two of the best understood lichen 

responses to N deposition across the Marine West Coast 

Forests ecoregion are changes in lichen community 

composition and accumulation of N in lichen thalli. 

Declines in the growth and photosynthesis of rare 

aquatic lichens have also been associated with elevated 

streamwater nitrate (NO3

-

) concentrations.

Responses to N emissions from an industrial fertilizer 

facility by coastal white spruce forests in south-central 

Alaska include declines in tree health (thinning 

crowns, chlorotic foliage), changes in forest understory 

composition, foliar nutritional imbalances (elevated 

nitrogen:magnesium and nitrogen:calcium), elevated 

NO3

-

 in forest fl oor and mineral soil, and declines in 

ectomycorrhizal fungal diversity and ectomycorrhizal 

fungal community change (sporocarp and root-tip level; 

Lilleskov et al. 2001, 2002; Whytemare et al. 1997). A 

suite of other responses typical of forested regions (see 

Chapters 7, 8, and 10) are also likely to be important 

in the Marine West Coast Forests ecoregion, including: 

increases in plant foliar N, soil N cycling and NO3

-

leaching; decreases in soil carbon:nitrogen; changes in 

species composition; and increased likelihood of pest 

outbreak (Brozek 1990, Compton et al. 2003, Perakis et 

al. 2006, Perakis and Sinkhorn in press, Prescott et al. 

1993, 1995).

9.3 Range of Responses Observed

9.3.1 Forest Lichens

Lichen community composition and chemistry are 

currently the best indicators for assessing ecosystem 

9.1 Ecoregion Description

Th e Marine West Coast Forests ecoregion encompasses 

the Pacifi c Coast from central Alaska and the Aleutian 

Islands south through northern California (CEC 1997; 

Fig. 2.2). Th e ecoregion description is adapted from 

CEC (1997). Topography is mountainous bordered 

by coastal plains. Soils range from very nutrient rich 

to nutrient poor; nitrogen (N) is the nutrient most 

commonly limiting to plant productivity, although 

some areas are naturally very rich in N. Th e maritime 

infl uence of the Pacifi c Ocean results in high 

precipitation (600 to 5000 mm yr-1), a long growing 

season, and moderate mean annual temperatures (5 to 9 

o

C). Within North America, all of the wettest climates, 

many of the most productive forests, and all temperate 

rain forests occur in this ecoregion. Many lowland 

aquatic ecosystems in this region sustain a diversity 

of anadromous salmonids that can be a signifi cant 

source of marine-derived N to rivers and lakes, whereas 

upland aquatic ecosystems are often nutrient poor. 

Terrestrial ecosystems range from mild, humid coastal 

rain forest to boreal forests, to cool sub-alpine forests 

and alpine conditions at higher elevations. Tree species 

composition of the temperate coastal forests varies by 

latitude, and from north to south, and includes white 

spruce (Picea glauca), Sitka spruce (Picea sitchensis), 

western hemlock (Tsuga heterophylla), Nootka cypress 

(Alaska yellow cedar; Callitropsis nootkatensis), red 

alder (Alnus rubra), western red cedar (Th uja plicata), 

Douglas-fi r (Pseudotsuga menziesii), and California 

redwood (Sequoia sempervirens). Many of these trees can 

reach large size and live to great age. In the drier rainshadow areas, Oregon white oak (Garry oak; Quercus 

garryana) and Pacifi c madrone (Arbutus menziesii) 

occur with Douglas-fi r. Mountain hemlock (Tsuga 

mertensiana) and Pacifi c silver fi r (Abies amabilis) 

dominate subalpine forests; alpine tundra is dominated 

by shrubs, herbs, mosses, and lichens.
90 Chapter 9—Marine West Coast Forest GTR-NRS-80

responses to N deposition across the Marine West Coast 

Forests ecoregion (Fenn et al. 2003, 2007; Geiser and 

Neitlich 2007; Glavich and Geiser 2008; Geiser et al. 

2010; Jovan 2008); information on lichen research 

methodology and terminology can be found in Chapter 

4 of this report.

A recent analysis by Geiser and Neitlich (2007) of 

1470 epiphytic lichen surveys in western Oregon and 

Washington shows that lichen community composition 

and thallus N concentrations shift continuously across 

the regional N deposition gradient of 0.8 to 8 kg N 

ha-1 yr-1 (Porter 2007). Oligotrophs and mesotrophs, 

encompassing nearly all of the larger bodied, endemic, 

and ecologically important species, comprised more than 

90 percent of species richness at 865 of the sites (Geiser 

et al. 2010). Oligotroph detections peak at N deposition 

ranges up to 2.5 kg N ha-1 yr-1 (McCune and Geiser 

2009). As N deposition increases, the relative proportion 

of mesotrophic species increases, peaking at deposition 

rates between 2.5 and 4.0 kg N ha-1 yr-1. At deposition 

rates of 4.5 kg N ha-1 yr-1, smaller, eutrophic species 

dominate and oligotrophs and mesotrophs decline 

to less than 12 percent of species. Oligotrophs play 

important ecological roles as forage, nesting material, 

habitat for mollusks and invertebrates, and contribute 

to N2-fi xation, nutrient cycling, and moderation 

of humidity; these roles are not fi lled by eutrophs 

(McCune and Geiser 2009). Th erefore, shifts in lichen 

community composition, particularly in old-growth 

forests where lichen biomass is highest, can have broader 

eff ects on forest ecosystems and food webs. Responses to 

N inputs are summarized in Table 9.1.

Lichen community-based air quality scores have 

been linearly correlated with N deposition estimates 

from Interagency Monitoring of Protected Visual 

Environments (IMPROVE), National Atmospheric 

Deposition Program (NADP), and Community 

Multiscale Air Quality (CMAQ; see Chapter 4 for 

descriptions). Geiser and Neitlich (2007) used a 

nonmetric multi-dimensional scaling ordination 

technique to separate climate from pollution eff ects on 

lichen communities and score air quality at each of the 

western Oregon and Washington survey sites. In this 

analysis, air score increased as air quality worsened. Air 

scores were directly and positively correlated with N 

concentrations in ammonium ion (NH4

+

) wet deposition 

at eight NADP sites (mg L-1 yr-1; r2

 = 0.55, p < 0.0001; 

Geiser and Neitlich 2007), and with ammonium sulfate 

((NH4

)2SO4) and ammonium nitrate (NH4NO3) 

concentrations in fi ne particulates at 12 IMPROVE sites 

in the study area (μg m-3 yr-1; r2

 = 0.93, p < 0.0001; Geiser 

et al. 2010). After accounting for precipitation, lichen air 

scores were also positively correlated with total kg N ha-1

yr-1 in wet deposition measured by NADP (r2

 = 0.64, p 

= 0.006) and modeled by CMAQ (r2

 = 0.23, p < 0.005). 

CMAQ deposition estimates were matched to each of 

1470 air scores by overlaying lichen survey coordinates 

on 36 x 36 km CMAQ grid cell modeled output. When 

dry deposition was included to estimate total deposition, 

still accounting for precipitation volume, the strength 

of correlations between air scores and CMAQ output 

improved (r2

 = 0.35, p<0.005).8 Responses of lichens to 

N concentrations are summarized in Table 9.2.

Field exposure of the air-pollution tolerant epiphytic 

lichens Parmelia sulcata and Platismatia glauca to 2 kg 

wet N ha-1 yr-1 in the western Oregon and Washington 

Cascades resulted in mainly benefi cial eff ects (shifts 

in carbon balance between cellular compartments, 

stimulation of carbon uptake, and improved protection 

from chlorophyll degradation), though some sensitive 

lichens were sparse at these sites (Ra et al. 2004, 2005). 

Th is deposition rate, while relatively low in comparison 

to polluted areas, is more than double the estimated 

background deposition of 0.8 kg N ha-1 yr-1 for Oregon 

and Washington (Porter 2007).

Th e sensitive and conspicuous epiphytic macrolichen, 

Menzies’ cartilage (Ramalina menziesii), is distributed 

throughout the Coast Ranges and inland valleys from 

southern California to southeastern Alaska. Fumigation 

with nitric acid (HNO3

) at levels from 7 to 25 μg 

8

Th e 36 x 36 km resolution for deposition estimates is low. 

Point measurements, especially in grid cells that include 

urban areas or steep deposition gradients, can be very diff erent 

than the grid cell mean. Th erefore correlations of 0.35 seem 

reasonable, and, because of statistical power generated by 

1470 measurements, the equation for the linear relationship 

is expected to be reliable. Uncertainty lies primarily in the 

accuracy of the model output.
Chapter 9—Marine West Coast Forest GTR-NRS-80 91

Table 9.1—Responses to N input for Marine West Coast Forests ecoregion.Ecosystem Site N input

kg N ha-1 yr-1

Indicator Response Study

Coniferous forest

Western OR; WA

~1-8 Lichens Defi ned regional indicator oligotroph, mesotroph, and 

eutroph species response to N deposition range

Jovan 2008

Coniferous forest

Western OR; WA

<1.5->4.5 Lichens Defi ned regional indicator oligotroph, mesotroph, and 

eutroph species response to N deposition range

McCune and Geiser 2009

Coniferous forest

Western OR; WA

2.7-9.2

CMAQ

0-7.0

NADP (wet)

0.4-4.4

CMAQ (wet)

Lichens Regional scale lichen community composition shifts 

from oligotroph dominance to mesotroph and eutroph dominance with increasing lichen thallus N; N input increases with precipitation

Geiser et al. 2010

Coniferous forest

Western OR; WA Cascades

2 (wet) Lichens N tolerant lichens Parmelia sulcata and Platismatia 

glauca from moderately polluted sites had increased chlorophyll and lipid content, cytoplasmic lipid droplets, and algal cell wall thickness, indicating altered storage allocation among cellular compartments 

Ra et al. 2004

Coniferous forest

Western OR; WA Cascades

2 (wet) Lichens Positive physiological effect on tolerant lichens but 

some sensitive lichens were sparse. Platismatia glaucaexposed to moderate levels of fertilizing air pollutants had increased N, S, K, Na; increased chlorophyll and carotenoids; increased OD435/415 ratios; and, in summer, increased CO2 uptake and decreased thallus 

density

Ra et al. 2005

Coniferous forest

S. Central Alaska

5 Tree Crown thinning, chlorotic foliage Lilleskov et al. 2001, 2002; 

Whytemare et al. 1997

Coniferous forest

S. Central Alaska

5 Tree foliar 

chemistry

Elevated N:Mg and N:Ca ratios Lilleskov et al. 2002, 

Whytemare et al. 1997

Coniferous forest

S. Central Alaska

5 Understory and 

ground cover

Compositional shifts Lilleskov et al. 2001, 

Whytemare et al. 1997

Coniferous forest

S. Central Alaska

5 Ectomycorrhizal 

fungi

Loss of diversity Lilleskov et al. 2001, 2002

Coniferous forest

S. Central Alaska

5 Ectomycorrhizal 

fungi

Compositional shifts towards nitrophilic taxa Lilleskov et al. 2001, 2002

Coniferous forest

S. Central Alaska

5 Soils Elevated available NO3

- Lilleskov et al. 2001, 2002; 

Whytemare et al. 1997
92 Chapter 9—Marine West Coast Forest GTR-NRS-80

m-3 (equivalent to concentrations observed in the 

Los Angeles basin) were unequivocally damaging and 

associated with increased membrane permeability, 

decomposition of photosynthetic pigments, and 

decreased carbon exchange capacity (Riddell et al. 

2008).

In southeastern Alaska, response thresholds for N 

concentration in the epiphytic macrolichens witch’s 

hair (Alectoria sarmentosa, 0.56 percent), tube lichen 

(Hypogymnia enteromorpha, 0.75 to 0.89 percent), 

inactive tube lichen (Hypogymnia inactiva, 0.76 

percent), and ragged lichen (Platismatia glauca, 0.59 to 

0.80 percent) have been used to detect hotspots of N 

deposition in forests near cruise ship docks in Juneau 

and Skagway’s Klondike Goldrush National Historic 

Park, near popular anchorages for marine fi shing vessels 

at Warren and Coronation Island Wildernesses, and at 

Greens Creek Gold mine (Dillman et al. 2007, Furbish 

et al. 2000). Increasing N concentrations in lichens 

during the past 15 years were attributed to additional 

N deposition from transpacifi c sources and increased 

regional marine traffi c (Dillman et al. 2007).

9.3.2 Other Forest Responses

General information on the response of the Marine 

West Coast Forests to N deposition, beyond 

information derived from lichens, is relatively sparse. 

Adverse eff ects of excess N resulting from industrial 

NH3 emissions have been observed in plants, soils, and 

mycorrhizae along a gradient of N deposition (range: 

0.7 to 21.0 kg N ha-1 yr-1) in coastal white spruce 

forests of south-central Alaska. Although estimates of N 

deposition along this gradient are not well constrained, 

and likely varied over ~30 years of industrial emissions, 

strong gradients in forest response to N were observed. 

Bulk N deposition of approximately 5 kg ha-1 yr-1 at 

the time of sampling was associated with declines in 

tree health (thinning crowns, chlorotic foliage), changes 

in forest understory composition, foliar nutritional 

imbalances (elevated N:Mg and N:Ca), elevated 

NO3

-

 in forest fl oor and mineral soil, and declines in 

ectomycorrhizal fungal diversity and ectomycorrhizal 

fungal community change (sporocarp and root-tip level; 

Lilleskov et al. 2001, 2002; Whytemare et al. 1997). 

Most changes were correlated with increased soil NO3

-

Table 9.2—Range of responses to N concentrations for lichen in Marine West Coast Forests ecoregion.

Ecosystem 

component

Site N input Indicator Response Study

Coniferous 

forest

Western OR; 

WA

>0.06 mg wet NH4

+

 L-1

from NADP

Lichens Declines in sensitive and increases in 

tolerant and eutrophic epiphytic lichens, 

increases in lichen thallus N. Lichen N 

thresholds for Platismatia glauca (0.59% 

N dry weight)

Geiser and 

Neitlich 2007

Mountain 

streams

Coast, 

Cascade, 

and Siskiyou 

Ranges of w OR 

WA and n CA

Up to 0.1 to 0.33 mg 

L-1 during summer fi eld 

visits

Lichens Presence of the aquatic lichens Peltigera 

hydrotheria and Leptogium rivale

Glavich 2009

Mountain 

streams

Western 

Cascades, OR

6.3-12.6 mg L-1 annual 

mean as NO3

-

Lichens Declines in rate of photosynthesis 

of Peltigera hydrotheria after 1 year 

exposure

Davis et al. 

2000

Coniferous 

forest

Western OR 

and WA

0.51 μg m-3 or 0.26-

0.33 kg N ha-1 yr-1in 

fi ne particulate N from 

(NH4)2SO4 and NH4NO3

(IMPROVE)

Lichens Shifts in lichen community composition 

dominance from oligotrophs to eutrophs

Geiser et al. 

2010;

Glavich and 

Geiser 2008

Deciduous 

forest

CA Coast 

Range

7-25 μg HNO3 m-3

fumigation

Lichens Increased membrane permeability, 

decomposition of photosynthetic 

pigments, and decreased carbon 

exchange capacity in Ramalina menziesii

Riddell et al. 

2008
Chapter 9—Marine West Coast Forest GTR-NRS-80 93

availability, which appears to be a useful indicator of 

long-term response to N inputs in this ecosystem.

Other pathways of N input to forests of the Marine 

West Coast Forests ecoregion provide insight into 

forest responses to added N. High rates of biological 

N2-fi xation by alder and ceanothus (Ceanothus spp.) are 

a distinguishing feature of many forested ecosystems 

in the region, and commercial forest fertilization with 

N is also widely practiced. While such N inputs diff er 

from atmospheric deposition in their mode, timing, 

and rates, they can nevertheless be useful for developing 

a qualitative understanding of the impacts of elevated 

N. Forest N fertilization usually occurs as a large dose 

of urea exceeding 200 kg N ha-1, typically applied 

once or twice over 40 to 80 years of plantation growth 

(Stegemoeller et al. 1990). Pure stands of red alder can 

sustain N2-fi xation rates of 100 to 300 kg N ha-1 yr-1over 

50 to 75 years of growth, adding fi xed N to ecosystems 

primarily in organic forms via above- and belowground 

tissue turnover, as well as some leakage of NH4

+

 from 

nodules (Binkley et al. 1994). Many of the observed 

eff ects of N addition by fertilization and biological 

fi xation qualitatively resemble those associated with 

atmospheric N deposition, including: elevated tissue N 

content (Brozek 1990), increased nitrifi cation and NO3

-

leaching to soils and streams (Compton et al. 2003), 

loss of divalent base cations from soil and increased 

soil acidifi cation (Van Miegroet and Cole 1984), and 

changes in forest understory composition (Prescott et 

al. 1993, 1995). However, it remains diffi cult to assess 

how high rates of N input from biological N2-fi xation 

and N fertilization translate quantitatively into response 

thresholds under chronic N deposition.

Responses to atmospheric N inputs in Marine West 

Coast Forests are likely to be shaped by wide variation 

in N status across the region. Site N status in the region 

ranges from ultra N-poor soils developed on young 

volcanic ash and glacial outwash to extremely N-rich 

soils that escaped the last glaciation and which rank 

among the most N-rich areas on Earth. Sites with high 

N status, particularly in coastal regions, often show 

negative or no response to N fertilization (Hopmans 

and Chappell 1994, Peterson and Hazard 1990). 

Responses to N fertilization can also vary depending 

on the dominant tree species considered, for example, 

with Douglas-fi r being more N responsive than western 

hemlock (Edmonds and Chappell 1993). Variation in 

N status infl uences ecosystem response to N2-fi xation 

and forest fertilization (Binkley et al. 1992, Edmonds 

and Hsiang 1987), and is therefore likely to shape 

response to atmospheric N deposition as well. Previous 

site occupancy by woody N2

-fi xers, fi re history, and 

climate all contribute to variations in site N status 

across the region. Legacy eff ects of N2-fi xing red alder 

are particularly signifi cant and can result in addition of 

>10,000 kg N ha-1 during the 75-year lifespan of a pure 

stand of red alder. In the moist Oregon Coast Range, 

historic N2

-fi xation by red alder is a proximate cause of 

exceptionally high soil N accumulation (~30 Mg ha-1), 

and ultimately yields coniferous forests that exhibit 

elevated nitrifi cation (80 kg N ha-1 yr-1), NO3

-

 leaching 

(25 kg N ha-1 yr-1), soil acidifi cation, base cation 

depletion, and aluminum (Al) mobilization (Perakis et 

al. 2006, Perakis and Sinkhorn in press). High foliar 

N also predisposes large areas of Douglas-fi r plantation 

forests to outbreaks of pathogenic Swiss needle cast 

fungi (Phaeocryptopus gaeumannii) (El Hajj et al. 

2004). Further N additions to such areas may have the 

potential to rapidly intensify symptoms of N saturation; 

the stages of N saturation in forested ecosystems are 

described in Chapter 7. Disturbances such as fi re and 

forest harvest can remove ~1,000 kg N ha-1 from a site 

and may ameliorate some eff ects of excess N, although 

N losses may be reversed within several decades where 

alder or ceanothus dominate early succession.

9.3.3 Aquatic Systems

Currently there is limited information on how 

atmospheric N deposition shapes aquatic ecosystem N 

status across the Marine West Coast Forests ecoregion. 

Transpacifi c transport of pollutants from Asia to western 

North America is an issue of increasing concern in 

the region (Jaff e et al. 1999, Wilkening et al. 2000). 

Short-lived spikes in N deposition have been associated 

with increased streamwater NO3

-

 in old-growth 

forested watersheds of Olympic National Park, raising 

concentrations from baseline values < 5 μg N L-1 up to 

40 μg N L-1, and illustrating the potential sensitivity of 
94 Chapter 9—Marine West Coast Forest GTR-NRS-80

old-growth forests in the region to added N (Edmonds 

et al. 1998). Th e infrequent and highly pulsed nature of 

these inputs, however, makes it diffi cult to understand 

long-term ecological eff ects and develop watershed N 

input-output relationships.

Variation in the cover of N2-fi xing red alder across 

watersheds of the Oregon coast range is related 

positively to streamwater fl uxes of NO3

-

 and major 

base cations. Assuming an N2

-fi xation rate of 150 kg 

ha-1 yr-1 for pure alder stands (Binkley et al. 1994), 

approximately 25 to 30 percent of fi xed N is lost 

as streamwater N export (Compton et al. 2003). 

Th ese losses correspond to maximum fl ow-weighted 

streamwater N concentrations of 2.5 mg L-1, mostly 

(>90 percent) as NO3

-

. Studies of N2-fi xing red alder 

forests in Oregon and Washington (Binkley et al. 1992) 

and of conifer forests in Oregon (Perakis and Sinkhorn 

in press) further suggest that the susceptibility of sites 

to NO3

-

 leaching increases with soil N status and 

nitrifi cation rates. Collectively, these results suggest a 

high potential for atmospheric N deposition to increase 

NO3

-

 leaching across some N-rich coastal areas of the 

Marine West Coast Forests ecoregion, and that some 

streams are already exposed to naturally elevated NO3

-

. 

In colder and more inland areas that support less active 

N2-fi xation, however, streamwater fl uxes of NO3

-

 remain 

low, with unknown sensitivity to atmospheric N inputs.

A lichen-indicated critical level for stream-water NO3

-

would appear to be between 6.3 to 12.6 mg N L-1. 

Th e critical level is the concentration above which 

adverse eff ects to sensitive vegetation may occur (UBA 

2004). Th is estimate is based on a year-long laboratory 

study of the growth and photosynthetic responses of 

the submerged aquatic hydrothyria lichen (Peltigera 

hydrothyria) collected from the Oregon Cascades and 

exposed to NO3

-

 concentrations between 1.6 and 

94.5 mg N L-1 (Davis et al. 2000). Th ese levels greatly 

exceed typical summer NO3

-

 levels (< 0.05 mg L-1) in 

the native habitat of this lichen; however stream N 

could be higher during fall fl ushes. During a 2-year 

fi eld study of randomly selected watersheds in western 

Oregon, western Washington, and northern California, 

the highest summer stream total N concentrations 

associated with occurrence of the aquatic skin 

(Leptogium rivale) and hydrothyria lichens were 0.1 and 

0.33 mg N L-1, respectively (Glavich 2009). Gilbert 

et al. (1997) observed a lower biological diversity in 

aquatic lichen communities in waters of England where 

total inorganic N concentrations were just above 0.6 

mg L-1.

9.4 Critical Loads Estimates

Th e onset of elevated soil nitrifi cation and NO3

-

availability, conifer nutrient imbalance and decline, 

declines in ectomycorrhizal fungal diversity, and 

ectomycorrhizal fungal community change are expected 

at bulk N deposition of approximately 5 kg ha-1 yr-1 in 

the northern forests of this region (Lilleskov et al. 2001, 

2002; Whytemare et al. 1997). Th is critical load must be 

considered expert opinion because historic unrecorded 

N inputs were likely to have been higher than inputs 

recorded at the time of study. In addition, it is based on 

only one study system in southeast Alaska that is almost 

certainly not representative of all forest types, soils, and 

climates across the region.

Geiser et al. (2010) created regression models relating 

lichen-based air scores to NADP and CMAQ measures 

of N deposition, accounting for precipitation. At an air 

score response threshold of 0.21, oligotrophs comprised 

as little as 30 percent, and eutrophs up to 34 percent, 

of species richness (see Table 4.1). Th e resulting critical 

loads were 2.7 to 9.2 kg N ha-1 yr-1 for total deposition 

(and 0.7 to 4.4 kg N ha-1 yr-1 for wet-only deposition). 

Because hardwood forests growing on valley fl oors favor 

eutrophs by providing a more nutrient-rich canopy 

drip and higher pH bark substrates compared to the 

dominant coniferous forests, the response threshold was 

selected to allow for a natural range in N-availability 

among clean sites. Critical loads increased with 

precipitation, which varied approximately tenfold across 

the study area, because high precipitation volumes 

dilute the impact of N loading on lichens. Th e 95 

percent confi dence interval for the 2.7 to 9.2 kg N ha-1

yr-1 critical load range was 0 to 13.5 kg N ha-1 yr-1. Th e 

critical load range can be considered reliable because 

the systematic sampling design yielded high quality, 

spatially extensive data. However, while the data for 
Chapter 9—Marine West Coast Forest GTR-NRS-80 95

total deposition are also extensive, they are based on 

a fairly coarse grid (36 km) with some uncertainties 

regarding accuracy of the modeled data. Critical loads 

are summarized in Table 9.3.

9.5 Comparison to Critical Loads from 

Europe and Other Ecoregions

Many vascular plants and cryptogams of the Tundra, 

Taiga, and Northern Forests ecoregions also occur in 

the Marine West Coast Forests ecoregion. For example, 

extensive peatlands in the Alaskan portion of the Marine 

West Coast Forests ecoregion support mat-forming 

lichens and bryophytes and ericaceous shrubs (heaths) 

typical of boreal ecoregions; alpine areas, especially in 

Alaska, support tundra species. Th erefore, references to 

Eurasian vegetation responses in these three preceding 

chapters are potentially relevant to parts of the Marine 

West Coast Forests. Th e estimated critical load range 

for Marine West Coast Forests (2.7 to 9.2 kg N ha-1

yr-1) is consistent with all other U.S. lichen-based 

estimates reported in this document. Th e upper-end 

estimate is consistent with European values, and the 

lower-end estimate is more than threefold greater than 

background deposition in the region. For mycorrhizal 

fungi, the results for this region correspond well with 

those expected for spruce-fi r forests for the northeastern 

United States (5 to 7 kg N ha-1 yr-1) and are lower than 

those reported for European forests (10 to 20 kg N 

ha-1 yr-1).

9.6 Future Research Directions and 

Gaps in Data

Th ere is a need to understand how the response of 

forest soils, streams, and their biota will be mediated 

by underlying variations in site N status in the region. 

Generally, lowland coastal Oregon and Washington 

forests and streams possess the highest N status in the 

region, yet may escape signifi cant elevated N inputs in 

the near-term due to the lack of signifi cant population 

and industrial activity along upwind coastal areas. 

Potential increases in N transport from Asia, however, 

could make these areas susceptible. On the other hand, 

naturally N-poor areas of the Oregon and Washington 

Cascades and glaciated montane forests of the Olympic 

Peninsula may be sensitive to added N originating 

from anthropogenic sources in the Willamette Valley 

and Puget Sound. Th e wide range in forest ages and 

in natural and anthropogenic disturbances (e.g., 

fi re, logging) across the region is likely to interact 

with N status in unknown ways to shape ecosystem 

responses to N deposition. A better understanding of 

the relative sensitivity of lichens and other ecosystem 

response indicators to diff erent forms of atmospheric N 

deposition would allow more accurate apportionment 

and prediction of the eff ects from diff erent emissions 

sources. Analysis of existing Forest Inventory and 

Analysis and other U.S. Forest Service lichen community 

data from southeastern and south-central Alaska will test 

their application to the northern parts of the ecoregion.

Table 9.3—Empirical critical loads of nutrient N for the Marine West Coast Forests ecoregion. 

Reliability rating: ## reliable; # fairly reliable; (#) expert judgment

Ecosystem 

component

Critical 

load for N 

deposition 

kg ha-1 yr-1

Reliability Response Comments Study

Western 

OR and WA 

forests

2.7-9.2 ## Lichen community 

composition

Critical load increases 

with regional range in 

mean annual precipitation 

from 45-450 cm

Geiser et al. 

2010

SE Alaska

forests

5 (#) Fungal community 

change; declines 

in ectomycorrhizal 

fungal diversity

Lilleskov 1999; 

Lilleskov et al. 

2001, 2002; 

Whytemare et 

al. 1997
96 Chapter 9—Marine West Coast Forest GTR-NRS-80

LITERATURE CITED

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Nitrogen fi xation by red alder: biology, rates, and 

controls. In: Hibbs, D.E.; DeBell, D.S.; Tarrant, 

R.F., eds. Th e biology and management of red alder. 

Corvallis, OR: Oregon State University Press: 57-72.

Binkley, D.; Sollins, P.; Bell, R.; Sachs, D.; Myrold,D. 

1992. Biogeochemistry of adjacent conifer and 

alder-conifer stands. Ecology. 73: 2022-2033.

Brozek, S. 1990. Eff ect of soil changes caused by red 

alder (Alnus rubra) on biomass and nutrient status 

of Douglas-fi r (Pseudotsuga menziesii) seedlings.

Canadian Journal of Forest Research. 20: 1320-1325.

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1997. Ecological regions of North America: 

Toward a common perspective. Montreal, Canada: 

Commission for Environmental Cooperation. 

71 p. Available at http://www.cec.org/fi les/pdf/ 

BIODIVERSITY/eco-eng_EN.pdf. (Accessed May 

24, 2010).

Compton, J.E.; Church, M.R.; Larned, S.T.; Hogsett, 

W.E. 2003. Nitrogen saturation in forested 

watersheds of the Oregon Coast Range: Th e 

landscape role of N2-fi xing red alder. Ecosystems. 

6: 773-785.

Davis, W.C.; Gries, C.; Nash, T.H. III. 2000. Th e 

ecophysiological response of the aquatic lichen, 

Hydrothyria venosa to nitrates in terms of weight 

and photosynthesis over long periods of time.

Bibliotheca Lichenologica. 75: 201-208.

Dillman, K.; Geiser, L.; Brenner, G. 2007. Air quality 

biomonitoring with lichens on the Tongass 

National Forest. Final Report. Petersburg, AK: 

Tongass National Forest, Air Resource Management 

Program. Available at http://gis.nacse.org/

lichenair/?page=reports#R10 (Accessed August 4, 

2010).

Edmonds, R.L.; Chappell, H.N. 1993. Relationships 

between soil organic matter and forest productivity 

in western Oregon and Washington. Canadian 

Journal of Forest Research. 24: 1101-1106.

Edmonds, R.L.; Hsiang, T. 1987. Infl uence of forest 

fl oor and soil properties on response of Douglas-fi r 

to urea. Soil Science Society of America Journal. 57: 

1332-1337.

Edmonds, R.L.; Blew, R.D.; Marra, J.L.; Blew, J.; Barg, 

A.K.; Murray, G.; Th omas, T.B. 1998. Vegetation 

patterns, hydrology, and water chemistry in small 

watersheds in the Hoh River Valley, Olympic 

National Park. NPS Scientifi c Monograph NPSD/

NRUSGS/NRSM-98/02. Washington, D.C.: U.S. 

Department of Interior, National Park Service. 131 p.

El-Hajj, Z.; Kavanaugh, K.; Rose, C.; Kanaan-Atallah, 

Z. 2004. Nitrogen and carbon dynamics of a foliar 

biotrophic fungal parasite in fertilized Douglas-fi r.

New Phytologist. 163: 139-147.

Fenn, M.E.; Baron, J.S.; Allen, E.B.; Rueth, H.M.; 

Nydick, K.R.; Geiser, L.; Bowman, W.D.; Sickman, 

J.O.; Meixner, T.; Johnson, D.W.; Neitlich, P. 2003. 

Ecological eff ects of nitrogen deposition in the 

Western United States. BioScience. 53(4): 404-420.

Fenn, M.E.; Geiser, L.; Bachman, R.; Blubaugh, T.J.; 

Bytnerowicz, A. 2007. Atmospheric deposition 

inputs and eff ects on lichen chemistry and 

indicator species in the Columbia River Gorge, 

USA. Environmental Pollution. 146: 77-91.

Furbish, C.E.; Geiser, L.; Rector, C. 2000. Lichenair quality pilot study for Klondike Gold Rush 

National Historical Park and the city of Skagway, 

Alaska. Final Report. Skagway, AK: U.S. Department 

of Interior: National Park Service, Klondike Goldrush 

National Historic Park. 47 p. Available at http://gis.

nacse.org/lichenair/?page=reports#R10 (Accessed May 

12, 2010)
Chapter 9—Marine West Coast Forest GTR-NRS-80 97

Geiser, L.H.; Jovan, S.E.; Glavich, D.A.; Porter, M. 

2010. Lichen-based critical loads for atmospheric 

nitrogen deposition in Western Oregon and 

Washington Forests, USA. Environmental 

Pollution. 158: 2412-2421.

Geiser, L.H.; Neitlich, P.N. 2007. Air pollution 

and climate gradients in western Oregon and 

Washington indicated by epiphytic macrolichens.

Environmental Pollution. 145: 203-218.

Gilbert, O.L.; Giavarini, V.J. 1997. Th e lichen 

vegetation of acid watercourses in England. Th e 

Lichenologist. 29: 347-367.

Glavich, D.A. 2009. Th e distribution, rarity, and 

habitats of three aquatic lichens in the US Pacifi c 

Northwest. Th e Bryologist. 112: 54-72.

Glavich, D.A.; Geiser, L.H. 2008. Potential 

approaches to developing lichen-based critical 

loads and levels for nitrogen, sulfur and metalcontaining atmospheric pollutants in North 

America. Th e Bryologist. 111: 638-649.

Hopmans, P.; Chappell, H.N. 1994. Growth response 

of young, thinned Douglas-fi r stand to nitrogen 

fertilizer in relation to soil properties and tree 

nutrition. Canadian Journal of Forest Research. 24: 

1684-1688.

Jaff e, D.; Anderson, T.; Covert, D.; Kotchenruther, R.; 

Trost, B.; Danielson, J.; Simpson, W.; Berntsen, T.; 

Karlsdottir, S.; Blake, D.; Harris, J.; Carmichael, G.; 

Uno, I. 1999. Transport of Asian air pollution to 

North America. Geophysical Research Letters. 26(6): 

711-714.

Jovan, S. 2008. Lichen bioindication of biodiversity, 

air quality, and climate: baseline results from 

monitoring in Washington, Oregon, and 

California. Gen.Tech. Rep. PNW-GTR-737. 

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

Forest Service, Pacifi c Northwest Research Station. 

115 p.

Lilleskov, E.A. 1999. Decline of above- and 

belowground ectomycorrhizal fungal diversity 

over an atmospheric nitrogen deposition gradient 

near Kenai, Alaska. Ithaca, NY: Cornell University. 

198 p. Ph.D. dissertation.

Lilleskov, E.A.; Fahey, T.J.; Lovett, G.M. 2001. 

Ectomycorrhizal fungal aboveground community 

change over an atmospheric nitrogen deposition 

gradient. Ecological Applications. 11: 397-410.

Lilleskov, E.A.; Fahey, T.J.; Horton, T.R.; Lovett, 

G.M. 2002. Belowground ectomycorrhizal fungal 

community change over a nitrogen deposition 

gradient in Alaska. Ecology. 83: 104-115.

McCune, B.M.; Geiser, L.H. 2009. Macrolichens of 

the Pacifi c Northwest. 2nd edtion. Corvallis, OR: 

Oregon State University Press. 400 p.

Perakis, S.S.; Maguire, D.A.; Bullen, T.D.; Cromack, 

K.; Waring, R.H.; Boyle, J.R. 2006. Coupled 

nitrogen and calcium cycles in forests of the 

Oregon Coast Range. Ecosystems. 9: 63-74.

Perakis, S.S.; Sinkhorn, E.R. In press. Biogeochemistry 

of a temperate forest nitrogen gradient. Ecology.

Peterson, C.E.; Hazard, J.W. 1990. Regional variation 

in growth response of coastal Douglas-fi r to 

nitrogen fertilizer in the Pacifi c Northwest. Forest 

Science. 36: 625-640.

Porter, M.K. 2007. Regional modeling of nitrogen, 

sulfur, and mercury atmospheric deposition in the 

Pacifi c Northwest. Pullman, WA: Washington State 

University. 127 p. M.S. thesis.

Prescott, C.E.; Coward, L.P.; Weetman, G.F.; Gessel, 

S.P. 1993. Eff ects of repeated nitrogen fertilization 

on the ericaceous shrub, salal (Gaultheria shallon) 

in two coastal Douglas-fi r forests. Forest Ecology 

and Management. 61: 45-60.
98 Chapter 9—Marine West Coast Forest GTR-NRS-80

Prescott, C.E.; Kumi, J.W.; Weetman, G.F. 1995. 

Long-term eff ects of repeated N fertilization and 

straw application in a jack pine forest. 2. Changes 

in the ericaceous ground vegetation. Canadian 

Journal of Forest Research. 25: 1984-1990.

Ra, H.S.Y.; Geiser, L.H.; Crang, R.F.E. 2005. 

Eff ects of season and low-level air pollution on 

physiology and element content of lichens from 

the U.S. Pacifi c Northwest. Science of the Total 

Environment. 343: 155-167.

Ra, H.-S.Y.; Rubin, L.; Crang, R.F.E. 2004. Structural 

impacts on thallus and algal cell components of 

two lichen species in response to low-level air 

pollution in Pacifi c Northwest forests. Microscopy 

and Microanalysis. 10(2): 270-279.

Riddell, J.; Nash, T.H. III; Padgett, P. 2008. Th e eff ect 

of HNO3 gas on the lichen Ramalina menziesii.

Flora. 203: 47-54.

Stegemoeller, K.A.; Chappell, H.N.; Bennett, W.S. 

1990. Regional forest nutrition research project 

analysis history: objectives and achievements.

RFNRP Rep. No. 12. Seattle, WA: University of 

Washington, College of Forest Resources. 33 p.

UBA (UmweltBundesAmt). 2004. Manual on 

methodologies and criteria for modeling 

and mapping critical loads and levels and air 

pollution eff ects, risks, and trends. Berlin: Federal 

Environmental Agency (UmweltBundesAmt). 

Available at http://icpmapping.org/cms/

zeigeBereich/11/manual-english.html (Accessed May 

12, 2010).

Van Miegroet, H.; Cole, D.W. 1984. Impact of 

nitrifi cation on soil acidifi cation and cation 

leaching in a red alder ecosystem. Journal of 

Environmental Quality. 13: 586-590.

Whytemare, A.B.; Edmonds, R.L.; Aber, J.D.; Lajtha, 

K. 1997. Infl uence of excess nitrogen deposition 

on a white spruce (Picea glauca) stand in southern 

Alaska. Biogeochemistry. 38: 173-187.

Wilkening, K.E.; Barrie, L.A.; Engle, M. 2000. TransPacifi c air pollution. Science. 290: 65-67.
Chapter 10—Eastern Temperate Forest GTR-NRS-80 99

10 EASTERN TEMPERATE FORESTS

F.S. Gilliam, C.L. Goodale, L.H. Pardo, L.H. Geiser, E.A. Lilleskov

America, particularly those of the mixed mesophytic 

forest region (Hinkle et al. 1993). Forest types range 

from beech-maple and maple-basswood in the east to 

mixed oak-hickory in the upper Midwest, and oakhickory-pine in the Appalachians southward. Ashes 

(Fraxinus spp.), elms (Ulmus spp.), black cherry (Prunus 

serotina), yellow-poplar (Liriodendron tulipifera), sweet 

gum (Liquidambar styracifl ua), basswood (Tilia spp.), 

hackberry(Celtis occidentalis), persimmon (Diospyros 

virginiana), eastern red cedar (Juniperus virginiana), 

and dogwood (Cornus spp.) are also wide ranging. 

Th e complex and diverse forest regions include: (1) 

the mixed mesophytic forest region; (2) the western 

mesophytic forest region; (3) the oak-hickory forest 

region; (3) the former oak-chestnut forest region; and 

(4) the oak-pine forest region (Braun 1950). Each of 

these regions displays an impressive number of canopy 

co-dominant species, and occupies widely ranging 

topography on numerous soil types derived from 

sharply contrasting parent materials. Th e herbaceous 

layer, most often defi ned as the stratum of vascular 

plants ≤1 m in height (Gilliam and Roberts 2003), can 

represent >90 percent of vascular plant species richness 

of temperate deciduous forest ecosystems in this 

ecoregion (Gilliam 2007).

10.2.2 Ecosystem Responses to N Deposition

Responses of eastern hardwood forests to excess 

nitrogen (N) deposition include increases in tissue N, 

soil N cycling, nitrate (NO3

-

) leaching, decreases in soil 

carbon:nitrogen (C:N) ratio, and shifts in community 

composition, including declines in species richness and 

abundance. Th e stages of N saturation are described 

in Chapter 7. Some ecosystem responses occur as 

a result of acidifi cation and therefore are caused by 

sulfate (SO4

-

) as well as NO3

-

. Th ese responses include 

soil nutrient cation imbalances (particularly calcium 

(Ca) and magnesium (Mg)) from enhanced leaching 

of NO3

-

, especially in soils derived from intermediate 

to acidic classes of parent material (Bailey et al. 2005). 

As the result of decreases in soil nutrient cation 

10.1 Ecoregion Description

Th e Eastern Temperate Forests ecoregion extends 

from the Atlantic Coast westward into eastern Texas, 

Oklahoma, Missouri, Iowa, and Minnesota. Th e 

ecoregion description is adapted from CEC (1997). 

Eastern Temperate Forests are comprised of two fairly 

distinct ecosystems: eastern hardwood forests and the 

Southeastern Coastal Plain. Th e former includes mixed 

wood plains and central plains, as well as Ozark and 

Ouachita-Appalachian forests of the northeastern and 

midwestern states and Appalachian Mountains. Th e 

latter is composed of the southeastern and coastal plains 

extending from Cape Cod and Long Island to Florida, 

inland to and around the Appalachian mountains (Fig. 

2.2; CEC 1997).

Th e Eastern Temperate Forests ecoregion is 

distinguished by its temperate, humid climate; dense, 

diverse forest cover; and dense human population 

(>160 million). Coastal plains underlain by sedimentary 

bedrock in the southeast abut older sedimentary, 

metamorphic, and igneous bedrock of the Appalachian 

Mountains (peaking at > 2,000 m). Mixed limestonedolomite plains and hills dominate the west central 

parts, sedimentary rock dominates the northwest 

plains and plateaus, and glacial activity shaped 

northern landscapes. Soils are mostly leached, being 

nutrient-poor to calcium-rich. Surface waters include 

abundant perennial streams, areas of high lake density, 

diverse wetlands, and rich maritime ecosystems. Th e 

climate grades latitudinally from cool, continental to 

subtropical. Summers are hot and humid (27 °C to 32 

°C daily maxima); winters are mild to cool (-12 °C to 4 

°C). Precipitation (100 to 150 cm yr-1) is relatively even 

throughout the year, with summer or spring peaks.

10.2 Eastern Hardwood Forest

10.2.1 Ecosystem Description

Eastern hardwood forests, as originally described by 

Braun (1950) and further characterized by Greller 

(1988), are among the most diverse forests of North 
100 Chapter 10—Eastern Temperate Forest GTR-NRS-80

availability, decreases in net primary productivity of 

forest ecosystems (i.e., forest decline) can occur. At 

the interface between terrestrial and aquatic habitats, 

excess N can cause increases in NO3

-

 in streams and 

lakes and, particularly in extreme cases, increases in 

the mobilization of aluminum (Al3+) in freshwater 

ecosystems (Driscoll et al. 2003).

Excess N deposition can also cause a loss of biodiversity. 

Gilliam (2006) identifi ed fi ve mechanisms to explain 

such loss: (1) alteration of interspecifi c competition; 

(2) increases in herbivory; (3) decreases in mycorrhizal 

infection; (4) increases in pathogenic fungal infection; 

and (5) enhanced likelihood of invasions by nonnative 

species. Herbaceous, epiphytic lichen, and bryophyte 

layers are generally most sensitive to a variety of 

disturbances, including chronically elevated N 

deposition.

Herbaceous species. Gilliam (2006) reported that the 

response of the forest herbaceous layer to increasing 

N may include the following: (1) initial increases in 

cover of the herb layer; (2) decreases in species richness 

resulting from loss of numerous N-effi cient species; (3) 

decreases in species evenness resulting from increasing 

dominance of a few high N-requiring species; and (4) 

declines in forest biodiversity as a result of decreases in 

both richness and evenness. In addition, these studies 

suggest that the response time of herb layer species to 

increases in N availability is more rapid when ambient 

N deposition is low. Th us, a rapid response should occur 

in areas receiving low levels of atmospheric deposition 

of N, such as that found for Hurd et al. (1998) in the 

Adirondacks, whereas a delayed response would be 

expected in areas receiving high N deposition, such 

as that found by Gilliam et al. (2006a). Th is response 

pattern has been reported for other sites in eastern 

hardwood forests (Fraterrigo et al. 2009, Royo et al. 

2010). Th e relationship between plant response time to 

added N and ambient N deposition could arise because 

the eff ect is ultimately a function of relative amounts of 

N added. For a given amount of N added, the relative 

addition would be higher for a site receiving lower 

amounts of ambient N deposition.

Still other work has focused on the importance of the 

herb layer in mediating N cycling (Gilliam 2007). 

Recent work (Moore et al. 2007, Muller 2003, Welch et 

al. 2007) has demonstrated that plants of the herb layer 

facilitate N cycling in a disproportionately large way 

compared to their minimal biomass. Th is is primarily 

because herb layer foliage is generally higher in nutrient 

content and more decomposable than tree foliage.

Lichens. Lichen community-level responses to changing 

sulfur dioxide (SO2), acidity, and nutrient N deposition 

regimes vary. Increased deposition in low deposition 

areas results in species composition shifts to tolerant 

species, while in high deposition areas species richness, 

abundance, and total landscape diversity decline 

dramatically. Lichen thallus concentrations of N and 

sulfur (S) also increase or decrease, corresponding to 

changes in deposition.

10.2.3 Range of Responses Observed

Eastern hardwood forest regions with higher levels of 

N deposition (US EPA 2002; Table 10.1) exhibit some 

of the ecosystem responses previously described in this 

chapter. For example, the long-term reference watershed 

at the Fernow Experimental Forest, West Virginia, 

exhibited a 20-year pattern of increasing NO3

-

, Ca2+, and 

Mg2+ export in streamwater; throughfall deposition was 

17.5 kg N ha-1 yr-1 (Peterjohn et al. 1996). At the Great 

Smoky Mountains National Park, export of streamwater 

N is elevated (Garten 2000, Van Miegroet et al. 2001) 

at sites within the park receiving deposition of about 

32 kg N ha-1 yr-1 (Pardo and Duarte 2007, Weathers 

et al. 2006). In a broad synthesis of data from 354 

upland forest catchments across the Northeast, Aber et 

al. (2003) reported that for catchments receiving less 

than 8 kg N ha-1 yr-1, surface water NO3

-

 concentration 

rarely exceeded 1 μmol L-1; above that threshold NO3

-

concentrations in surface water increased. Th e crosscatchment analysis covered both Eastern Temperate 

Forests and Northern Forests (Chapter 7). Tree 

growth response from Wisconsin to Maine and south 

to Virginia show general increases in net growth and 

carbon storage as N deposition increases from 3 to 11 

kg N ha-1 yr-1, refl ecting an overall fertilization eff ect of 

N deposition on eastern and northern forests (Th omas 
Chapter 10—Eastern Temperate Forest GTR-NRS-80 101

et al. 2010). However, responses vary considerably 

by species. All hardwood tree species with arbuscular 

mycorrhizal associations (red maple [Acer rubrum], 

sugar maple, white ash [Fraxinus americana], yellowpoplar, and black cherry) as well as six tree species 

with ectomycorrhizal associations (balsam fi r [Abies 

balsamea], pignut hickory [Carya glabra], eastern 

white pine [Pinus strobus], quaking aspen [Populus 

tremuloides], northern red and scarlet oak [Quercus 

rubra and Q. coccinea]) had marked increases in growth 

with increasing deposition. In contrast, red pine (Pinus 

resinosa), red spruce (Picea rubens), and northern white 

cedar (Th uja occidentalis) all had statistically signifi cant 

decreases in growth with increasing deposition, and 

eight tree species (yellow birch [Betula alleghaniensis], 

eastern white pine, basswood, quaking aspen, bigtooth 

aspen [Populus grandidentata], scarlet oak, chestnut 

oak [Quercus prinus], and northern red oak; all 

ectomycorrhizal) all had decreased survivorship with 

increasing deposition. Th ese declines in survivorship 

could be due to multiple factors, including: direct eff ects 

of N itself, spatial covariation between N deposition 

and exposure to other pollutants (e.g., O3, SO4

2-), 

N-induced nutrient imbalances, interactions between N 

loading and secondary stressors such as drought or insect 

outbreaks, or suppression by competition from other 

tree species (Th omas et al. 2010).

Table 10.1—Responses to N input for the Eastern Temperate Forests ecoregion

Site N inputs

kg ha-1 yr-1

Response Comments Study

New Jersey Pine 

Barrens

<8 Decreasing 

ectomycorrhizal fungi 

morphotype richness

Dighton et al. 2004

North Carolina and 

West Virginia deposition 

gradient

>7 Stream NO3

-

 > DON Brookshire et al. 2007

Northeastern Forest 

catchments

8 Surface water NO3

- ≥ 1 

μmol L-1

Aber et al. 2003

Southeastern deposition 

gradient

9-15 Increasing foliar N 

concentration

Boggs et al. 2005

North Carolina and 

West Virginia deposition 

gradient

<10 Soil C:N <20 Brookshire et al. 2007

Michigan deposition 

gradient

<12 Arbuscular mycorrhizal 

fungal biomass decline 

and community 

composition change

van Diepen et al. 2008, 

2010

Fernow Experimental 

Forest, WV

<17.5 Herb layer changes Gilliam et al. 2006a

Fernow Experimental 

Forest, WV

17.5 Increasing streamwater 

Ca2+, Mg2+, NO3

-

Deposition as 

ambient throughfall

Peterjohn et al. 1996

Cedar Creek LTER, MN <28 Ectomycorrhizal fungi 

community response

Avis et al. 2008

Great Smoky Mountains 

National Park

32 Export of streamwater N Garten 2000, Pardo 

and Duarte 2007, Van 

Miegroet et al. 2001, 

Weathers et al. 2006

Michigan deposition 

gradient

<35 Declines in fungal 

biomass

van Diepen et al. 2007, 

2010

Cedar Creek LTER, MN <50 Ectomycorrhizal fungi 

community response

Avis et al. 2003
102 Chapter 10—Eastern Temperate Forest GTR-NRS-80

Table 10.2—Response of Eastern Temperate Forests ecosystems to experimental N 

fertilization (≤ 50 kg ha-1 yr-1)

Site Location Harvard Forest, 

Massachusetts

Millbrook, New 

York

Fernow Forest, 

West Virginia

Forest Type Red pine 

plantation

Black oak & 

red maple

Chestnut & 

northern red oak

Mixed central 

hardwoods

Type of N Deposition Wet + dry Wet + dry Wet + dry

N Deposition

kg ha-1 yr-1

9 10 15

N fertilization rate

kg ha-1 yr-1

+50 +50b +35

Form of N Fertilizer NH4NO3 NH4NO3 (NH4)2SO4

Start year 1988 1996 1989

Duration (years) 14 9 11 --Response relative to Controla--

Tree growth 0/– 0/+ + –

Mortality + 0 + 0,+

Foliar %N + + ND +, 0

Foliar %Ca 0 0 ND -, 0

NO3

-

 leaching + 0 + +

Cation loss + 0 + +

Soil C/N 0 0 ND 0

N mineralization 0 0 ND 0

Nitrifi cation + 0 ND 0

Soil respiration –

0

ND ND

Microbial biomass – – ND ND

References 1-6 1-6 7 8-11

a

Responses are indicated as increase (+), decrease (–), or no response (0) in fertilized relative to control 

plots, or no data (ND) in indicated references.

b

Started at 100 kg ha-1 yr-1 for 1996-1999, then reduced to 50 kg ha-1 yr-1 thereafter.

References: 1) Magill et al. 2000; 2) Magill et al. 2004; 3) Frey et al. 2004; 4) Bowden et al. 2004; 5) Minocha 

et al. 2000; 6) Venterea et al. 2004; 7) Wallace et al. 2007; 8) May et al. 2005; 9) Adams et al. 2006; 10) 

Edwards et al. 2006; 11) DeWalle et al. 2006; 12) Gilliam and Adams 1996.

Experimental N additions indicate broadly similar 

responses to the gradient studies above (Table 10.2). 

At the Fernow Experimental Forest, experimental 

additions of N (35 kg ha-1 yr-1) increased NO3

-

 leaching 

(Edwards et al. 2006), decreased soil nutrient cation 

availability (Gilliam et al. 1996), and decreased tree 

growth by some species but not others (May et al. 

2005, DeWalle et al. 2006). Th ese results resemble 

those from fertilization experiments at an oak stand 

at Millbrook, New York (Wallace et al. 2007), and at 

a red pine stand and an oak-maple stand at Harvard 

Forest, Massachusetts (Table 10.1; Bowden et al. 2004, 

Frey et al. 2004, Magill et al. 2000, Magill et al. 2004, 

Minocha et al. 2000, Venterea et al. 2004). In all but 

the oak-maple stand at Harvard Forest, moderate 

rates of long-term fertilization induced NO3

-

 leaching, 

which led to losses of soil base cations. Foliar N content 

increased at both the red pine and oak-maple stands 

at Harvard Forest for the duration of that experiment 

(Magill et al. 2004), whereas the increase in foliar N 

concentration at the Fernow Experimental Forest was 

transient in some species (black cherry, red maple) and 

persistent in others (yellow-poplar; DeWalle et al. 2006, 

May et al. 2005). Th ese changes in foliar N have had 

mixed consequences for tree growth. At Harvard Forest, 

fertilization enhanced growth in the oak-maple stand, 

but stimulated mortality in the mature red pine stand 

following a drought in 1995 (Magill et al. 2004). At 
Chapter 10—Eastern Temperate Forest GTR-NRS-80 103

Millbrook, New York, chronic N fertilization induced 

signifi cant mortality in a 60- to 100-year-old chestnut 

oak/northern red oak stand, but also increased the 

relative growth rate of the remaining trees, for a net 

eff ect of an overall decrease in live basal area in fi ve of six 

treatment plots after 8 years (Wallace et al. 2007). Th e 

most oak mortality occurred at fertilized sites that also 

had low Ca:Al ratios in surface organic layers (Wallace et 

al. 2007). At the Fernow Experimental Forest, some tree 

species (red maple, river birch [Betula nigra]) had no 

response to fertilization, whereas others (black cherry, 

yellow-poplar) increased growth during the fi rst 7 years, 

then had decreased growth relative to control trees in 

years 9 - 12 (DeWalle et al. 2006). Soil processes also 

often respond to chronic fertilization. Although none 

of these studies had detectable changes in soil C:N ratio 

(Magill et al. 2004, Adams et al. 2006), fertilization 

resulted in higher net nitrifi cation rates at Fernow Forest 

(Gilliam et al. 2001) and at the Harvard Forest red pine 

plots (Magill et al. 2000, Venterea et al. 2004). After 

more than a decade of N addition, both the red pine 

and the oak-maple stands at the Harvard Forest had 

decreases in microbial biomass (Frey et al. 2004) and 

depressed rates of soil respiration relative to the control 

plot (Bowden et al. 2004).

In a gradient study spanning a range of N deposition 

rates across North Carolina and Virginia, Boggs et al. 

(2005) found an increase in foliar N concentration in 

American beech, sugar maple, and yellow birch as wet N 

deposition increased from 9 to 15 kg N ha-1 yr-1. Foliar 

Ca:Al ratios were not aff ected. Forest fl oor pH and C:N 

ratios decreased and rates of net N mineralization and 

nitrifi cation increased with increasing deposition. In 

an independent gradient study spanning N deposition 

rates from 5 to 32 kg N ha-1 yr-1 in North Carolina 

and West Virginia, Brookshire et al. (2007) found 

that stream NO3

-

 and dissolved organic N (DON) 

concentrations both increased with increasing N 

deposition, although NO3

-

 increased most dramatically. 

Nitrate concentrations exceeded DON concentrations 

with >7 kg N ha-1 yr-1 of deposition. Th ey also observed 

a steep decrease in surface soil C:N ratio from 30 to 

17 as deposition increased from 5 to 10 kg N ha-1 yr-1, 

after which soil C:N ratio did not decrease further with 

increasing deposition.

Th e impact of excess N via acidifi cation was assessed 

by both Fox et al. (1989) and Adams et al. (1991) 

using a “Green-Yellow-Red” screening model to 

determine levels of acid deposition (N and S) that 

would impact Class 1 wilderness areas. At the green line 

value, vegetation health and vigor would be protected; 

at the red line value, changes in N cycling, species 

composition, or other ecosystem features could occur 

(Fox et al. 1989). For Joyce Kilmer, North Carolina, 

and Slick Rock, Tennessee, the green line value for total 

N deposition was 7 to 10 kg N ha-1 yr-1; damage would 

occur at 15 kg N ha-1 yr-1 (red line). In Otter Creek, 

West Virginia, the green line value was 7 kg N ha-1 yr-1; 

the red line value was 10 to 15 kg N ha-1 yr-1 (Fox et al. 

1989). Adams et al. (1991) used the same approach to 

determine N and S deposition levels that would prevent 

foliar injury and fi sh species shifts. Th ey set green line 

values of 5 to 8 kg N ha-1 yr-1 for Dolly Sods, West 

Virginia; Hercules Glade, Missouri; and Otter Creek, 

West Virginia. Signifi cant ecosystem damage could be 

expected at total N deposition level >20 kg N ha-1 yr-1

(Adams et al. 1991).

Herbaceous layer. Responses of the herb layer to excess 

N vary among temperate deciduous forest sites. A 

northern forest (Adirondack) ecosystem had large shifts 

in herb cover in response to modest N additions (Hurd 

et al. 1998). At the Watershed Acidifi cation Study at 

Fernow Experimental Forest, West Virginia, which has 

ambient deposition of approximately 17.5 kg N ha-1

yr-1 in throughfall (Adams et al. 1993), addition of 35 

kg N ha-1 yr-1 via aerial application has led to changes 

in herbaceous species composition. Recently, similar 

changes in herbaceous species composition have been 

observed over time on the adjacent reference (control) 

watershed, which receives only atmospheric deposition9

. 

Previous research (Gilliam et al. 2006a) at the same 

watershed had shown no signifi cant diff erences in 

herb cover, composition, or several diversity indicators 

between treatment and control watersheds for any of 

the years sampled (1991, 1992, and 1994). Gilliam et 

al. (2006a) suggested that the initial lack of observed 

9

Gilliam, F.S. Unpublished data. Professor, Department of 

Biological Sciences, Marshall University, Huntington, WV 

25755-2510.
104 Chapter 10—Eastern Temperate Forest GTR-NRS-80

response to N addition was due to high background 

levels of N deposition that had already saturated 

watershed soils. Because the ecosystem had already 

been altered by N inputs, changes in the herb layer had 

already occurred. Th is is consistent with the conclusion 

of Hurd et al. (1998) that sensitivity of the herb layer 

to N additions was largely a function of ambient N 

deposition.

Ectomycorrhizal fungi. Ectomycorrhizal fungi form 

associations with many of the dominant trees of drier 

or nutrient-poor sites in the region, including oak, 

hickory, beech, birch, and basswood (Quercus, Carya, 

Fagus, Betula, Tilia). Th ere has been less research on 

ectomycorrhizal response to N deposition in hardwood 

than in conifer forests. We expect hardwood forest types 

that have a strong ecological overlap with conifer forest 

types of this region (e.g., oligotrophic beech-dominated 

northern hardwoods and oligotrophic oak forests) to 

have similar community responses to N deposition, but 

at this time, this supposition is unsupported by data. 

In an oak savannah forest in east-central Minnesota, 

ectomycorrhizal fungal community response (reduced 

species richness of sporocarps, changes in communities 

seen on root tips and sporocarps) was seen with addition 

of 50 and 170 kg ha-1 yr-1 over 18 to 20 years (Avis et 

al. 2003). Th e treatments also involved fertilization 

with other nutrients, so are not strictly equivalent to 

N deposition. In an N-only fertilization experiment 

Avis et al. (2008) also found a change in oak forest 

ectomycorrhizal fungal communities with deposition 

plus additions of N of approximately 28 kg N ha-1 yr-1. 

However, we do not know the lower threshold for this 

response. Studies of other ectomycorrhizal community 

types (e.g., Betulaceae, Salicaceae) are lacking.

Arbuscular mycorrhizal fungi. Arbuscular mycorrhizal 

fungi form associations with many tree taxa dominant 

in this forest type, including maple, ash, yellow-poplar, 

sweet gum, and cherry (Acer, Fraxinus, Liriodendron, 

Liquidambar, and Prunus). We have very little empirical 

data on arbuscular mycorrhizal community response 

to N in forests. One study in sugar maple dominated 

forests in Michigan has found declines of arbuscular 

mycorrhizal fungal biomass and change in community 

composition in response to addition of 30 kg-1 ha-1

yr-1 over and above background deposition across a 

gradient ranging from approximately 5 to12 kg ha-1

yr-1 wet deposition (van Diepen et al. 2007, 2010; van 

Diepen 2008). Th ere is some indication that arbuscular 

mycorrhizal biomass has declined and community 

composition has shifted over the N deposition gradient 

(van Diepen et al. 2007, 2010; van Diepen 2008), 

suggesting a threshold of <12 kg ha-1 yr-1 of wet 

deposition, but it is likely that there are interactions 

with within-site factors, and defi nitive determination of 

a response threshold awaits further analysis.

Lichens. Air pollution has been a major factor in the 

depauperization of the lichen fl ora in the eastern 

hardwoods region during the past 100 years (e.g., 

see Murphy et al. (1999; Massachusetts), Smiley and 

George (1974; New York), Showman and Long (1992; 

Pennsylvania); Wetmore (1989) and Showman (1975; 

Ohio); Wetmore (1988) and McCune (1988; Indiana 

and Michigan); or Lawrey and Hale (1988; Virginia)). 

Historically, the most damaging pollutants have been 

SO2, sulfuric acid (H2SO4), 

nitrogen oxides(NOx), and 

nitric acid (HNO3). As an example, declines in lichen 

species diversity exceeding 80 percent in Cuyahoga 

National Park (Ohio) and Indiana Dunes National 

Lake Shore, between 1895 and 1988, were attributed to 

SO2

 (Wetmore 1988,1989). Nutrient N, especially in 

the form of ammonium (NH4

+

), is a newly recognized 

concern.

It is entirely possible that a widespread modifi cation 

of native lichen fl ora has taken place across the eastern 

forests due to air pollution and habitat loss. In 2007, 

after substantial, regionwide decreases in SO2

 and NOx

emissions, total wet N deposition was still 2 to 5 times 

higher throughout the eastern hardwoods ecoregion 

compared to background western U.S. localities. In 

some regions, NH4

+

 is a signifi cant portion of deposition 

(NADP 2009). Although species richness remains high 

in parts of the southern Appalachians and southeastern 

Georgia (McCune et al. 1997), north central Florida 

(DeBolt et al. 2007), and Maine (Hinds and Hinds 

1998), the paucity of historical records complicates 

confi rmation of intact lichen community composition, 

species richness, cover, and/or landscape-level diversity 

in most locations.
Chapter 10—Eastern Temperate Forest GTR-NRS-80 105

Declines in SO2

 and NOx emissions in response to the 

Clean Air Act have been linked to improvements in 

epiphytic lichen diversity and cover in the upper Ohio 

Valley (Showman 1981, 1990, 1998) and south central 

Pennsylvania (McClenahen et al. 2007) and lower 

tissue S concentrations in Virginia (Lawrey and Hale 

1988). For the Ohio and Pennsylvania study areas, 2007 

total wet N deposition was still between 6.2 and 7.2 

kg N ha-1 yr-1 in 2007 (NADP 2009). As a result, the 

returning species are moderately acid- and SO2-sensitive, 

but N tolerant (Geiser et al. 2010). Th is response is 

consistent with substantial regional decreases in acidity 

but increasing N in wet deposition between 1980 and 

2007 (NADP 2009).

More information on lichen research methodology and 

terminology can be found in Chapter 4 of this volume.

10.2.4 Critical Loads Estimates

Th e critical load for decreased tree growth and 

survivorship in response to N deposition is >3 kg N 

ha-1 yr-1, while the critical load for herbaceous species is 

<17.5 kg N ha-1 yr-1

.

 Empirical critical loads for nutrient 

N in eastern hardwood forest ecosystems can be set at 

8 kg N ha-1 yr-1 for NO3

-

 leaching (Table 10.5). Th is 

threshold applies for surface water NO3

-

 leaching in the 

northeastern United States (Aber et al. 2003), as well 

as for select forests in the southeastern United States, 

where Brookshire et al. (2007) observed stream NO3

-

concentrations to exceed those of dissolved organic N 

above 7 kg N ha-1 yr-1. Similarly, Lewis (2002) used 10 

kg N ha-1 yr-1 as a deposition criterion for “minimally 

disturbed” watersheds in a compilation of N yield from 

watersheds across the United States.

Critical loads for nutrient N calculated using the steadystate mass balance method (Pardo 2010, UBA 2004) 

were reported (Table 10.3) for Great Smoky Mountains 

National Park as ranging from 3 to 7 kg ha-1 yr-1 (Pardo 

and Duarte 2007). Earlier calculations of critical loads 

for N nutrient within Great Smoky Mountains National 

Park made using data from the Integrated Forest Study 

(Johnson and Lindberg 1992) ranged from 2.5 to 9 kg 

N ha-1 yr-1 (Oja and Arp 1998).

Critical loads for ectomycorrhizal fungi, arbuscular 

mycorrhizal fungi, and lichens have not been reported 

previously. Th e critical load for ectomycorrhizal fungi is 

expected to be much less than the N input level of 28 kg 

ha-1 yr-1 reported by Avis et al. (2008); the critical load 

for arbuscular mycorrhizal fungi is also expected to be 

less than the 35 kg ha-1 yr-1 that caused a mycorrhizal 

response in research by van Diepen et al. (2007), 

and may be less than 12 kg N ha-1 yr-1 (van Diepen 

2008). Arbuscular mycorrhizal fungal critical loads are 

tentatively set at <12 kg N ha-1 yr-1 (van Diepen 2008). 

Th e value for arbuscular mycorrhizal fungi is considered 

expert judgment, as it is based on one gradient with only 

four sites.

Lichen community-based critical loads are diffi cult to 

pinpoint in the eastern hardwoods region due to the 

confounding eff ect of acidity from sulfur-containing 

pollutants, habitat alterations, and changing climate. 

However, N critical loads are probably similar to 

those proposed for epiphytic macrolichens in other 

temperate forests of the United States (northwest 

forested mountains (3.1 to 5.2 kg N ha-1 yr-1), west coast 

marine forests (2.7 to 9.2 kg N ha-1 yr-1), Mediterranean 

Table 10.3—Previously determined critical loads of nutrient N for the Eastern Temperate 

Forests ecoregion

Site Critical load for 

nutrient N 

kg N ha-1 yr-1

Comments Study

Great Smoky Mountains 

National Park

3-7 Determined using 

steady state mass 

balance method

Pardo and Duarte 2007

Great Smoky Mountains 

National Park

2.5-9 Determined using 

steady state mass 

balance method

Oja and Arp 1998
106 Chapter 10—Eastern Temperate Forest GTR-NRS-80

California (3.3 to 5.5 kg N ha-1 yr-1)), due to species 

overlap and physiological/ecological similarities among 

epiphytic macrolichens. Application of the western 

Oregon and Washington model relating epiphytic 

lichen community composition to N deposition 

(Chapter 4, Geiser et al. 2010), to eastern hardwoods 

yields a critical load estimate of 4 to 8 kg N ha-1 yr-1. 

Th is value was calculated using a precipitation range of 

71 to 305 cm, and a lichen community-composition 

response threshold that allows as few as 25 percent 

oligotrophs and no more than 47 percent eutrophs. 

Th is response threshold is higher than that used for 

Oregon and Washington to allow for the higher 

nutrient content of leaf exudates and more neutral bark 

substrates in hardwood forests, which favor eutrophs. 

(Coniferous forests predominate in western Oregon and 

Washington.)

A number of estimates of critical loads for acidity have 

also been set for the region and are summarized in 

Table 10.4. Critical loads for N associated with acidity 

often tend to be higher than those for nutrient N; for 

this ecoregion, they were of similar magnitude. Th e red 

line/green line approach of Fox et al. (1989) has been 

used to establish critical loads for acidity for eastern 

forests. Using the green line value determined for Joyce 

Kilmer, North Carolina, the critical load for acidity for 

the overstory and herb layer was 7 kg N ha-1 yr-1. Th is 

is consistent with calculations made for this region by 

McNulty et al. (2007), who made coarse estimates for 

critical loads for total acidity across the United States. 

Based on those calculations, the contribution of N to 

those critical loads for acidity is ~10 kg ha-1 yr-1. Earlier 

assessments in this region suggest that the critical load 

for acidity has been exceeded for both terrestrial and 

aquatic ecosystems (Fox et al. 1989, Sullivan and Cosby 

2002, Sullivan and Cosby 2004, Sullivan et al. 2003).

10.2.5 Comparison to European Critical Loads

Th ese estimates for critical loads are somewhat lower 

than for comparable ecosystems in Europe. Critical 

loads for European deciduous forests are set at 10 to 15 

kg N ha-1 yr-1 for increased NO3

-

 leaching and increased 

N mineralization and nitrifi cation; at 15 to 20 kg N 

ha-1 yr-1 for alterations in tissue chemistry and increased 

susceptibility to pest outbreak; and at 10 to 20 kg 

N ha-1 yr-1 for alterations in mycorrhizal community 

composition (UBA 2004). In a review and synthesis on 

eff ects of atmospheric ammonia (NH3) on terrestrial 

vegetation that included critical load estimates, Krupa 

(2003) reported a critical load range of 15 to 20 kg N 

ha-1 yr-1 for deciduous forests of Europe.

Herbaceous layer. Critical loads for ground vegetation 

in temperate forests are set at 10 to 15 kg N ha-1 yr-1 in 

Europe (UBA 2004). Critical loads research in Europe 

often employs gradients in N deposition (Brunet et al. 

1998, Strengbom et al. 2003). Most such work is done 

in conifer forests of Europe, limiting direct comparisons 

to the herb layer of temperate deciduous forests in the 

eastern United States. Brunet et al. (1998) employed 

multivariate methods to assess potential eff ects of 

N deposition on herb layer species of oak forests by 

sampling along a gradient of N deposition in southern 

Sweden. Species that increased in cover with increased 

N deposition were generally both nitrophilous and acid 

Table 10.4—Previously determined critical loads of acidity for the Eastern Temperate 

Forests ecoregion

Site Critical load for 

N acidity 

kg N ha-1 yr-1

Response Study

Joyce Kilmer, NC; Slick 

Rock, TN

7-10 Vegetation health 

and vigor

Fox et al. 1989

Otter Creek, WV 7 Vegetation health 

and vigor

Fox et al. 1989

Otter Creek, WV; 

Hercules Glade, MO; 

Dolly Sods, WV

5-8 ANC = 25 Adams et al. 1991
Chapter 10—Eastern Temperate Forest GTR-NRS-80 107

Ecosystem Ecosystem 

component

Critical load for nutrient N kg N 

ha-1 yr-1

Reliability Response Comments Study

Eastern hardwood forests

Trees >3 # Decreased growth of 

northern white cedar, and decreased survivorship of yellow birch, scarlet and chestnut oak, quaking aspen, and basswood

Thomas et al. 2010

Eastern hardwood forests

Lichens 4-8 (#) Changes in epiphytic lichen 

community composition

Geiser et al. 2010

Eastern hardwood forests

Forests 8 ## Increased surface water 

loading of NO3

-

Aber et al. 2003

Eastern hardwood forests

Arbuscular mycorrhizal fungi

<12 (#) Arbuscular mycorrhizal 

fungal biomass decline and community composition change

Observed along a Michigan N gradient

van Diepen 2008

Eastern hardwood forests

Herbs <17.5 (#) Herb layer changes Observed at Fernow 

Experimental Forest, WV 

Gilliam 2006, 2007, Gilliam et al. 2006a

Southeastern Coastal Plain

Lichens 4-6 (#) Changes in epiphytic lichen 

community composition

Appliation of Geiser et al. 2010

Southeastern Coastal Plain

Ectomycorrhizal fungi

5-10 (#) Ectomycorrhizal fungi 

community response

Dighton et al. 2004; Lilleskov et al. 2001, 2002, 2008

Table 10.5—Empirical critical loads of nutrient N for the Eastern Temperate Forests ecoregion. Reliability rating: ## reliable; # fairly reliable; (#) expert judgment
108 Chapter 10—Eastern Temperate Forest GTR-NRS-80

tolerant. Th ey concluded that N deposition infl uenced 

vegetation directly by increasing N availability and 

indirectly by increasing soil acidity.

10.2.6 Future Research Directions and Gaps in Data

Research on critical loads for eastern hardwood forests 

of the United States lags far behind research for similar 

forests in Europe. Particularly lacking are empirical 

data on community responses to gradients of ambient 

N deposition. Plot-scale fertilizer studies have been 

conducted, but these are not as numerous as similar 

studies in Europe, and carry with them the inherent 

diffi culty of scaling up from the plot to the landscape 

scale. Studies that employ whole-watershed treatment 

with N, such as that at Fernow Experimental Forest, 

are extremely rare. Additional complexities aff ecting 

responses to N loading include interactions between N 

addition and climate events (disturbance and variation), 

prior land use, species composition, site characteristics, 

and hydrology. For example, the coincidence of 

increased N availability, drought, and soil nutrient 

cation depletion appear to have led to signifi cant tree 

mortality at sites in Massachusetts (Magill et al. 2004) 

and eastern New York (Wallace et al. 2007), suggesting 

that the incipient N saturation at the site made the 

trees more susceptible to secondary stresses. Multifactor experimental manipulations that addressed N 

interactions with climate and other stressors could help 

address some of these uncertainties.

Also useful would be more studies focusing on gradients 

in N deposition (e.g., Boggs et al. 2005, Brookshire et 

al. 2007) and on forests of varying stand age (Castro et 

al. 2007). Gradient studies would be especially useful in 

fi lling the large data gaps that exist on the response of 

mycorrhizal fungal communities to realistic levels of N 

deposition in this region, especially in the development 

of response functions to N deposition that would permit 

establishment of critical loads.

A framework grid system has been systematically 

surveyed for lichens in New England and Mid-Atlantic 

states (Maine through Virginia) by the U.S. Forest 

Service Forest Inventory and Analysis/Forest Health 

Monitoring (FIA/FHM) program (McCune 2000, WillWolf et al. 2006); matching deposition estimates and 

supplementary sites are needed to interpret these data 

to estimate critical loads. Despite major reductions in 

SO2 and NOx

, critical loads for nutrient N for epiphytic 

lichens are probably still exceeded in much of the 

region. Lichen fl oras dating back to about 1900, where 

available, would help elucidate the changes that have 

already taken place. Research is also needed to separate 

H2SO4

 and HNO3 acidifi cation eff ects from fertilizing 

N eff ects on lichen community composition, diversity, 

and physiology, with consideration to the infl uences 

of habitat alteration and climatic changes. Forests of 

Minnesota and Wisconsin, where hydrogen ion (H+

) 

and SO4

-

 deposition are close to background levels but 

NO3

-

 and NH4

+

 deposition are elevated, may off er an 

opportunity for such studies when compared to forests 

of Ohio, Pennsylvania, West Virginia, and New York 

with high acidity, S deposition, and N deposition.

10.3 Southeastern Coastal Plain

10.3.1 Ecosystem Description

Forest ecosystems of the Southeastern Coastal Plain have 

evolved under conditions of relatively high frequency of 

fi re and hurricanes (Gilliam et al. 2006b). Th e forests 

of the coastal plain contrast with other forest types in 

the Eastern Temperate Forests ecoregion in that: (1) 

coastal plain forests tend to comprise pine-dominated 

overstories of low biodiversity and herb-dominated 

understories often of very high biodiversity (Gilliam 

2007, Platt et al. 2006); and (2) coastal plain forests 

experience lower levels of atmospheric deposition of N.

10.3.2 Ecosystem Responses to N Deposition

Coastal Plain forest responses to N deposition are not 

well studied. Generally, we would expect the same 

responses to excess N as seen in eastern hardwood 

forests: increases in tissue N concentration, declines in 

biodiversity, shifts in soil N cycling and NO3

-

 leaching, 

decreases in soil C:N ratio, and shifts in community 

composition (section 10.2.2). Interactions of N 

deposition with ecosystem processes in the Southeastern 

Coastal Plain are complicated by a changing N 

deposition regime, spatially variable fi re dynamics, 

and disturbances caused by large tropical storms. 

Historically, many pine-dominated forests of the Coastal 
Chapter 10—Eastern Temperate Forest GTR-NRS-80 109

Plain have been fi re maintained. Fire varies in frequency 

throughout the region and alters the N status of the 

ecosystem (Boerner et al. 2004). In a meta-analysis, 

Wan et al. (2001) found that fi re invariably caused 

substantial loss of N from fuels (i.e., forest litter). For 

example, as much as 50 percent of litter N was lost in a 

single fi re for conifer forests.

Tropical storms are an important disturbance regime 

in forest ecosystems of the coastal plain (Gilliam et 

al. 2006b). Th ese disturbances can have a substantial 

impact of nutrient cycling in these ecosystems. It is 

possible that the frequency/intensity of tropical storms is 

increasing as a result of global warming.

10.3.3 Range of Responses Observed

Although the eff ect of S deposition on coastal plain 

ecosystems has been studied (Janicki et al. 1995), little 

research has been done on responses of the coastal plain 

ecosystem to excess N deposition.

Ectomycorrhizal fungi. One study in the Pine Barrens 

region of New Jersey found a trend of decreasing 

morphotype richness with increasing N deposition 

(Dighton et al. 2004). Th ere was signifi cantly lower 

mycorrhizal morphotype richness at a site with bulk 

deposition of 8 kg N ha-1 yr-1 than at a site with 

approximately 4 kg N ha-1 yr-1 (Dighton et al. 2004). 

Low number of sample points in the gradient weakens 

the evidence for causal linkages, but this evidence is 

consistent with that from other oligotrophic coniferdominated ecosystems that suggest thresholds of 

somewhere between 5 and 10 kg ha-1 yr-1 for community 

response (Lilleskov et al. 2001, 2002, 2008). More 

productive pine or deciduous ecosystems in the southern 

part of this ecosystem type, on richer soils or with 

signifi cant N removals via harvest and/or fi re, might 

have a higher deposition threshold. However, no data 

are available from the region with which to test this 

hypothesis.

Lichens. Acidifying and fertilizing pollutants in 

deposition (N and S), primarily from urban/industrial 

emissions, have strongly infl uenced lichen communities 

in the Southeastern Coastal Plain. Th e characteristic 

response has been loss of sensitive species resulting in 

sharply reduced species richness and reduced lichen 

cover. For example, even by 1966, epiphytic lichen 

diversity was low (14 species per plot) on the north 

shore of Long Island, up to 62 km from Brooklyn 

(New York City; Brodo 1966). Th ere was complete 

extirpation of all lichens within 19 km of Brooklyn’s 

industrial center, even though Brooklyn and the north 

shore of Long Island shared the same fl ora historically 

(Brodo 1966). Of 70 species detected around 1900 on 

Plummer’s Island, a protected habitat 15 km south of 

Washington, D.C., only 20 could be found in the early 

1990s (Lawrey 1993). Such extreme de-pauperization 

of the lichen fl ora is less pervasive in the southeastern 

states. In an analysis of FIA/FHM epiphytic 

macrolichen data from forests of Georgia, the Carolinas, 

eastern Tennessee, and Virginia, McCune et al. (1997) 

observed pollution-tolerant species and lower species 

richness in urban and industrial areas, but pollutionsensitive and ‘luxuriant and diverse’ lichens (up to 40 

species per plot) in many rural areas. Th e richer fl ora is 

consistent with generally lower deposition of S and N in 

states of the southeastern compared to the northeastern 

United States. However, lacking historic fl ora, it is not 

possible to confi rm whether species are missing from 

communities in the southeast that are experiencing 

lower S and N deposition. Some aspects of the fl ora hint 

at a widespread loss of species due to air pollution: only 

one fruticose genus was detected and cyanolichens were 

rare.

Fire has less infl uence on the N regimes of forest lichens 

compared to soil-rooted plants in the Southeastern 

Coastal Plain. Lichens that are not consumed or heatkilled during forest fi re become a propagule source for 

recolonization without long term shifts in community 

composition (Johansson et al. 2006, Sillet and Goslin 

1997). However, increased fi re frequency and short 

timber harvest cycles favors faster growing species over 

those with greater dispersal or growth rate limitations, 

shifting community composition and reducing diversity 

(DeBolt et al. 2007).

More information on lichen research methodology and 

terminology can be found in Chapter 4 of this volume.
110 Chapter 10—Eastern Temperate Forest GTR-NRS-80

10.3.4 Critical Loads Estimates

Until more research has been done in Southeastern 

Coastal Plain forests, we can only off er provisional 

critical loads. Our best estimate of the critical load for 

ectomycorrhizal community integrity is 5 to 10 kg 

N ha-1 yr-1, and even this must be regarded as highly 

uncertain and only applicable to low productivity pine 

barrens forests (Table 10.5). In other ecosystem types, 

critical loads for mycorrhizal fungi appear to be roughly 

equivalent to those for other receptors, but defi nitive 

characterization of critical loads for other receptors and 

responses awaits further research.

 Application of the western Oregon and Washington 

model relating epiphytic lichen community composition 

to N deposition (Chapter 4, Geiser et al. 2010) to 

Southeastern Coastal Plains yields a critical load estimate 

of 4 to 6 kg ha-1 yr-1. Th is value was calculated using 

a precipitation range of 102 to 178 cm, and a lichen 

community-composition response threshold that allows 

as few as 30 percent oligotrophs and no more than 34 

percent eutrophs. Th is response threshold represents 

the upper end of the response threshold range used for 

Oregon and Washington to allow for potential favoring 

of the mesotrophs and eutrophs by warmer temperature. 

Th ese species are generally more drought resistant than 

oligotrophs.

10.3.5 Comparison to European Critical Loads

Th e concept of critical loads for N was originally 

developed for forest ecosystems of Europe (Nilsson 

and Grennfelt 1988). Whereas most of these forests 

have long been altered by human use (Dambrine et al. 

2007), they generally do not experience frequent natural 

disturbances, such as fi re and extreme wind events. 

Th us, it is diffi cult to compare Southeastern Coastal 

Plain ecosystems to European ecosystems.

10.3.6 Future Research Directions and Gaps in Data

More research is needed on the eff ects of N deposition 

to ecosystems of the Southeastern Coastal Plain. 

Potential research areas include watershed and plot scale 

fertilization studies and deposition gradient studies. 

In addition, research should be conducted to examine 

how the interaction of frequent disturbance and N 

deposition aff ect the N status of the Coastal Plain. 

Research is also needed to separate H2SO4 and HNO3

acidifi cation eff ects from fertilizing N eff ects on lichen 

community composition, diversity, and physiology, with 

consideration to the infl uences of habitat alteration, 

climatic changes and changes in fi re frequency. Southern 

Georgia and north-central Florida, with the lowest N 

and S deposition, may be good locations for fi eld-based 

N addition studies.

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Magill, A.H.; Martin, M.E.; Hallet, R.A.; Stoddard, 

J.L. 2003. Is nitrogen deposition altering the 

nitrogen status of Northeastern forests? BioScience. 

53(4): 375-389.

Adams, M.B.; Nichols, D.S.; Federer, C.A.; Jensen, 

K.F.; Parrott, H. 1991. Screening procedure to 

evaluate eff ects of air pollution on eastern region 

wildernesses cited as class I air quality areas. Gen. 

Tech. Rep. NE-151. Radnor, PA: U.S. Department 

of Agriculture, Forest Service, Northeastern Forest 

Experiment Station. 33 p.

Adams, M.B.; Edwards, P.J.; Wood, F.; Kochenderfer, 

J.N. 1993. Artifi cial watershed acidifi cation on 

the Fernow Experimental Forest, USA. Journal of 

Hydrology. 150: 505-509.

Adams, M.B.; DeWalle, D.R.; Peterjohn, W.T.; 

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Chapter 11—Great Plains GTR-NRS-80 117

11 GREAT PLAINS

C.M. Clark

dwarf shrubs. Tallgrass prairie is primarily dominated 

by the taller C4

 grasses: big bluestem (Andropogon 

gerardii), indiangrass (Sorghastrum nutans), switchgrass 

(Panicum virgatum), and little bluestem (Schizachyrium 

scoparium), interspersed with C3 (cool season)11 grasses, 

forbs, and shrubs. Mixed-grass prairie is a mixture of 

the two and can shift dramatically in composition with 

variation in climate (Samson and Knopf 1994). Th e 

historical combination of low precipitation, frequent 

fi res, and coevolution with native herbivores (e.g., North 

American bison, Bison bison) maintains the persistence 

of this generally grass-dominated ecoregion (CEC 1997, 

Knapp et al. 1999, Samson and Knopf 1994). More 

recently, reductions in fi re frequency and grazing have 

contributed to encroachment by more woody vegetation 

across the Great Plains (Briggs et al. 2005).

Human activity has greatly modifi ed the structure and 

function of the Great Plains. Because of its fertile soil, 

the Great Plains have been and continue to be heavily 

used for agriculture and rangelands. It is estimated that 

of the approximately 120 million ha of historical U. S. 

prairie, nearly 96 percent, 59 percent, and 50 percent, 

respectively, of the tallgrass, mixed-grass, and shortgrass 

prairie, have been lost as a result of human activity 

(Samson and Knopf 1994). Th e remaining areas exist 

mostly as a patchwork of prairie remnants. Many of 

the species on these remnants are still considered at risk 

due to long-term fi re suppression, alteration of grazing 

regimes, and isolation, leading to unfavorable conditions 

for persistence (Briggs et al. 2005, Leach and Givnish 

1996). In the Great Plains States, 61 plant species are 

11.1 Ecoregion Description

Th e North American Great Plains are the largest 

contiguous ecoregion in North America, covering 3.5 

million square km2

, or 16 percent of the continental 

area (CEC 1997). In the United States, the Great 

Plains ecoregion encompasses a roughly triangular 

region (Figure 2.2), bordered on the west by the Rocky 

Mountains and the southwestern deserts in Texas and 

northern Mexico, on the east by the Eastern Temperate 

and Northern Forests ecoregions, and extending well 

into Canada to the north. Precipitation decreases 

from east to west, and the dry climate, combined with 

historical fi re and grazing regimes, generally limits tree 

expansion except in localized riparian areas (Briggs and 

Knapp 1995, Wells 1965). Savannas and local forest 

cover are also possible in the wetter eastern Great Plains, 

though periodic fi re and grazers play an important role 

in determining the balance of herbaceous and woody 

vegetation across the Great Plains (Briggs and Knapp 

1995). Precipitation ranges from 320 to 1020 mm 

per year, and temperatures range from -5 to 30 °C; 

both peak during the summer growing season (Bailey 

1998). Topography is generally fl at, with some plateaus 

and slightly rolling plains. Soils are generally deep 

and fertile, with little leaching of minerals and strong 

buff ering potential.

Due to the dry climate, Great Plains ecoregion 

vegetation is primarily composed of herbaceous species, 

with grasses dominant, and forbs, sedges, and shrubs 

subdominant. However, in the drier areas to the west 

and south, shrubs can codominate; in the east and 

north, as well as along riverbanks and depressions, trees 

can dominate. Th e Great Plains usually are subdivided 

into three subregions based on precipitation. From 

west to east (drier to wetter), these are the shortgrass 

prairie, mixed-grass prairie, and tallgrass prairie (CEC 

1997, Samson and Knopf 1994). Shortgrass prairie 

is primarily dominated by the C4

 (warm season)10

grasses blue grama (Bouteloua gracilis) and buff alo grass 

(Buchloe dactyloides), interspersed with succulents and 

10C4 grass: Any species of grass that uses the C4 photosynthetic 

pathway, which initially produces a 4-carbon intermediary 

during fi xation of carbon dioxide. Th is process is more 

energetically expensive but enables more effi cient water usage 

and less degradation of photosynthetic enzymes.

11C3

 grass: Any species of grass that uses the C3 photosynthetic 

pathway, which initially produces a 3-carbon intermediary 

during fi xation of carbon dioxide. Th is process is less 

energetically expensive but is also less effi cient at using water 

and leads to more degradation of photosynthetic enzymes.
118 Chapter 11—Great Plains GTR-NRS-80

listed as Federally protected (threatened, endangered, 

or extirpated), and 1029 are listed as state protected 

(NRCS 2009). Th e short-, mixed-, and tallgrass prairie 

generally correspond to the western rangelands, the 

wheat belt, and the corn/soybean regions (CEC 1997). 

Wet N deposition generally ranges from 1 to 7 kg 

N ha-1 yr-1, increasing from west to east in roughly 

equal proportions of ammonium (NH4

+

) and nitrate 

(NO3

-

): 1.3:1 ratio of NH4

+

: NO3

-

 (NADP 200912). Dry 

deposition is poorly defi ned for the area, but site-specifi c 

studies indicate that dry deposition can account for at 

least half of total deposition (Gilliam 1987, Knapp et 

al. 1998). Consequently, we estimate total N deposition 

over the Great Plains at 2 to 14 kg N ha-1 yr-1.

11.2 Ecosystem Responses to N 

Deposition

Grasslands characteristic of the Great Plains are 

commonly N-limited or co-limited by N and other 

resources (e.g., water, phosphorus (P)), and are 

dominated by plants capable of rapid response to 

changing conditions (Hooper and Johnson 1999, 

Knapp and Smith 2001, Vitousek and Howarth 1991). 

As such, the Great Plains are expected to be sensitive 

to elevated N input. Following N deposition, there is a 

complex cascade of potential ecosystem responses which 

fall into three broad categories: biogeochemical and soil 

microbial responses, plant population and community 

responses, and responses at higher trophic levels 

(Galloway et al. 2003).

Biogeochemical and soil microbial responses:

• Increase in N availability in the soil for 

microbial immobilization and plant growth

• Change in microbial community structure and 

activity (often more bacteria)

• Increase or decrease in decomposition based on 

litter chemistry and other factors

• Increase in NH3 volatilization (depending on 

soil pH)

• Increase in N leaching

• Increase soil acidifi cation

Plant individual, population, and community responses:

• Increase in plant growth and allocation to 

aboveground structures

• Increase in tissue N content

• Decrease in light levels and alteration of 

germination conditions

• Shifts in competitive interactions and 

community composition

• Loss of species; often of rare, short, long-lived, 

native, and N-fi xing species

Higher trophic level responses:

• Short-term increase in production at higher 

trophic levels due to consumption of more 

aboveground structures that are more 

nutritionally rich

• Long-term increase or decrease depending on 

whether plant community shifts toward or away 

from herbivore-preferred plant species

Th e above responses are general, and likely to be 

aff ected by several factors important in the Great Plains. 

Indeed, the structure and function of the Great Plains 

ecosystem is co-regulated by a complex interaction 

among multiple-resources, co-evolution with large 

populations of herbivores, and frequent punctuated 

disturbances, such as wildfi re, that have eff ects on 

(and are aff ected by) responses at multiple hierarchical 

levels (Knapp et al.1998, Risser et al. 1981, Seastedt 

and Knapp 1993). As such, analysis of the eff ects of N 

deposition on the Great Plains is incomplete without 

incorporation of disturbance, especially from fi re, 

herbivory, and climatic vulnerability. Unfortunately, the 

paucity of studies examining the eff ects of N addition 

at levels comparable to N deposition (discussed in the 

next section) means that little can be concluded about 

potential interactions with these important regulatory 

factors. One might expect that the removal of N via 

fi re and/or herbivores would enhance N-limitation and 

therefore sensitivity to N deposition, though this likely 

depends on regional climatic factors, the intensity of 

12Data from Great Plains sites were averaged over all 

available years to estimate N deposition and the average 

ratio of NH4

+

: NO3

-

.
Chapter 11—Great Plains GTR-NRS-80 119

herbivory, and the dominant mechanism regulating 

ecosystem structure and function (Knapp et al. 1998). 

Because of these uncertainties, we focus our analysis 

on known eff ects of relatively low input rates of N 

comparable to deposition, with some extrapolation to 

higher levels.

11.3 Range of Responses Observed

Few studies in the Great Plains have explicitly examined 

the impact of rates of N input similar to ambient N 

deposition rates. Indeed, of the 108 papers reviewed for 

this chapter that examined the eff ects of N enrichment 

on the Great Plains, only 17 included treatments ≤30 

kg N ha-1 yr-1, relevant for direct assessment of critical 

loads (range: 10 to 25 kg N ha-1 yr-1). Furthermore, 

9 of these 17 included other nutrient amendments in 

addition to N (e.g., Tilman 1987), and 12 of these 17 

are from only two geographic locations (Cedar Creek 

Ecosystem Science Reserve, Minnesota, and the Center 

for Subsurface and Ecological Assessment Research, 

Oklahoma). Th ese limitations warrant caution when 

drawing conclusions. It is unclear whether studies that 

also add non-N resources would underestimate the 

eff ect of N deposition (because other eff ects such as 

soil acidifi cation and cation loss might not occur) or 

overestimate eff ects of N deposition (because co-limiting 

nutrients were also added). However, because (1) studies 

that add only N but at higher rates show qualitatively 

similar patterns; and (2) the site that added non-N 

resources is not co-limited and is largely unresponsive 

to addition of non-N resources (Tilman 1984, 1987), 

we are confi dent in using these studies in this review. 

Th e sections that follow in this chapter are structured 

by ecosystem responses, with all three prairie types 

grouped, owing to the small number of studies. Th e fi rst 

few paragraphs describe the eff ects from low-N studies 

(≤30 kg N ha-1 yr-1) and the last paragraph in each 

section extends the review to include notable high-N 

studies (>30 kg N ha-1 yr-1).

11.3.1 Biogeochemical and Soil Microbial Responses 

in the Great Plains

Jorgensen et al. (2005) added N at 16.3 kg ha-1 yr-1 to a 

mixed-grass prairie with a background total deposition 

of 11 kg N ha-1 yr-1 in Oklahoma for 5 years and 

measured several key ecosystem responses. Soil NO3

-

increased within 1 year and remained elevated by an 

average of 250 percent over the next 5 years. Th is N was 

not effi ciently retained in the system. Indeed, leaching 

of N increased nearly twelvefold, and the dominant 

form of N-loss shifted from dissolved organic N (DON) 

(75 percent of N-losses in control plots) to NO3

-

 (85 

percent of N-losses in fertilized plots). Although litter 

decomposition was not strongly aff ected by N addition, 

there was a tendency for litter N loss to be inhibited 

after 180 days in fertilized plots. Denitrifi cation was 

not generally stimulated by added N except after large 

rain events or with supplemental carbon (C), thus this 

process appeared to be C-limited and oxygen-inhibited 

more than N-limited. Gross N transformation rates were 

not sampled within fertilized plots.

Tilman (1987) added N at a range of rates (0, 10, 

20, 34, 54, 95, 170, and 270 kg N ha-1 yr-1) to four 

Minnesota sites near the ecotone of mesic prairie and 

mixed woodland (background wet deposition 6 kg N 

ha-1 yr-1). We focus our assessment on the two old fi elds 

(fi elds abandoned from agriculture) that most closely 

represent native prairie of the Great Plains, though 

patterns exhibited following N addition to all four fi elds 

were qualitatively similar. Th is experiment also included 

addition of non-N mineral nutrients (P, potassium (K), 

calcium (Ca), magnesium (Mg), sulfur (S), and trace 

metals), and was resampled several times in its 28-year 

history to assess the impacts of eutrophication. Nitrogen 

addition at 10 and 20 kg N ha-1 yr-1 for 12 years did 

not generally increase soil NO3

-

 or net N mineralization 

rates except in the more N-poor fi eld, though both 

were elevated at higher treatment rates in both fi elds 

(Wedin and Tilman 1996). After 22 years, low rates of 

N addition (10 and 20 kg N ha-1 yr-1) in the previously 

unresponsive old-fi eld increased net N mineralization 

rates and soil extractable NO3

-

 tended to increase at all 

treatment rates (Clark et al. 2009). Th e other fi eld was 

not examined in this later study. Finally, although much 

of the added N at low rates was retained after 12 years 

in both fi elds (Wedin and Tilman 1996), this retention 

effi ciency was diminished after 22 years, as N sinks 

became saturated and species composition shifted at 

lower treatment rates (Clark and Tilman 2008, Clark et 

al. 2009).
120 Chapter 11—Great Plains GTR-NRS-80

Th e bulk of the nutrients and C in prairie ecosystems 

are found in the soil, suggesting the ultimate fate of 

deposited N is likely in complex organic matter stored 

in the soil. To explore this, Epstein et al. (2001) added 

10 kg N ha-1 yr-1 of 15N to three shortgrass steppe 

communities (C3, C4, and mixed communities) in 

northeastern Colorado, and sampled various pools 

over 3 years (background wet deposition not reported). 

Th ey found that, although the C3

 community initially 

retained more 15N after the fi rst year (especially in 

plant material, Fig. 11.1), after the third year retention 

effi ciencies were fairly low (range: 20 to 35 percent). 

In addition, after 3 years, more 15N was retained in the 

C4 and mixed C3-C4 communities primarily in the fi ne 

particulate organic matter fraction of the soil (Epstein 

et al. 2001). Another study added 25 kg N ha-1 yr-1 to 

fi ve sites across the Great Plains (Table 11.1), and found 

after 3 years that 46 to 84 percent of the added N was 

retained primarily in the fi ne particulate organic matter 

of the top 20 cm of the soil, with increasing retention in 

soils with higher soil organic C and fi ner texture (Barret 

and Burke 2002). Th e lower retention effi ciencies 

found in Epstein et al. (2001) compared with Barrett 

and Burke (2002; as well as others) was explained 

by higher losses to grazing in Epstein et al. (2001), 

though it is unknown whether higher herbivory led to 

a redistribution of 15N or actual losses (i.e., increased 

volatilization, denitrifi cation). Numerous tracer studies 

from forests corroborate the evidence that added N 

often resides in the soil organic or mineral layer rather 

than in plant tissue (Aber et al. 1998, Currie et al. 2004, 

Nadelhoff er et al. 2004).

Studies using higher rates of N addition than those 

already discussed generally confi rm the above dynamics 

and further elucidate responses following N addition. 

Nitrogen addition (40 kg ha-1 yr-1) to monocultures 

of prairie grassland species in Minnesota increased 

net N mineralization (46 percent), and respiration 

rates of the labile (21 percent) and recalcitrant (10 

percent) soil C, and signifi cantly decreased soil labile 

C (31 percent) and microbial N (20 percent) (Dijkstra 

et al. 2006, Reich et al. 2001). Th e same experiment 

demonstrated that although N addition also increased 

gross N mineralization rates (West et al. 2006) and 

plant tissue quantity and quality (discussed in next 

section), there was not an observed increase in the rate 

of decomposition (Knops et al. 2007). Meta-analyses 

have found few general relationships between N 

addition and decomposition, though decomposition 

is often inhibited in low quality tissue (high lignin:N) 

and when N addition rates are low compared with 

ambient N deposition (Knorr et al. 2005). Detailed 

decomposition studies using higher rates of N addition 

(100 kg ha-1 yr-1) have found that fi ne-scale soilsubstrate interactions, for example from possible shifts 

in microbial community composition and activity, 

likely determine whether N addition will generally 

0

10

20

30

40

50

60

70

80

90

100

C3 Mixed C4

Percentage (%) Initial 15N Retained

1995 1997

Soil - Filled Plant - Hashed

a

c

b ab

d cd

Figure 11.1—Percentage initial 15N retained in soil 

(solid) and plant compartments (hatched) for three 

different plant communities (C3, mixed, C4) in 1995 

(1 month after addition) and in 1997 (three growing 

seasons after addition). Bars are +1 SE. Difference 

letters represent statistically signifi cant differences 

(P<0.05) (Epstein et al. 2001, reprinted with kind 

permission of Springer Science+Business Media).
Chapter 11—Great Plains GTR-NRS-80 121

increase or decrease decomposition (Hobbie 2005, 

2008; Keeler et al. 2009). In tallgrass prairie where N 

addition (100 kg ha-1 yr-1) was studied in combination 

with fi re and simulated grazing (mowing), N availability 

was unaff ected by burning, was elevated with burning 

plus N addition, and was not decreased with additional 

mowing (Collins et al. 1998).

Mycorrhizal responses to N deposition in the Great 

Plains ecoregion have been documented in gradient 

studies and fertilization experiments. Using grassland 

soil data from an N gradient in the Chicago area, 

Egerton-Warburton13 reported a decline in both 

mycorrhizal colonization and spore density as total 

extractable N increased. Th ere was a steep decline in 

arbuscular mycorrhizal fungal activity between 5-10 

g g-1 extractable soil inorganic N. While N deposition 

values were not available for all of these sites, National 

Atmospheric Deposition Program (NADP) values of 

9.7 kg ha-1 yr-1 were reported from the site with highest 

arbuscular mycorrhizal fungal activity, and 12.4 kg ha-1

yr-1 from a site with reduced activity. Other studies using 

higher rates of N inputs (≥40 kg N ha-1 yr-1) generally 

confi rm the eff ects of N on microbial populations, 

reporting shifts in microbial communities from fungal 

to bacterial dominance (Bradley et al. 2006, Chung 

et al. 2007), increased activity for several key soil 

enzymes (phosphatase, -glucosidase, and peroxidase) 

(Chung et al. 2007), and decreased symbiotic microbial 

populations (Johnson 1993, Johnson et al. 2003). Th ese 

eff ects appear to be especially pronounced when soils are 

more P-rich (Egerton-Warburton et al. 2007, Johnson 

et al. 2003).

11.3.2 Plant Community Responses in the 

Great Plains

Nitrogen addition for 4 years at 16.3 kg ha-1 yr-1 to a 

mixed-grass prairie in Oklahoma (background total 

deposition of 11 kg N ha-1 yr-1)increased total plant 

biomass and led to a fi vefold increase in tall fescue 

(Festuca arundinacea) cover, although total cover was 

relatively unaff ected (Clark et al. 2003, Jorgensen et 

al. 2005). Tall fescue cover was weakly, and often 

not signifi cantly, correlated with cover of other plant 

functional groups (negatively with native C4

 grasses and 

positively with non-leguminous forbs). Species richness 

decreased with increased tall fescue cover in the second 

and third years of the study (this eff ect was not seen in 

year 1, and was not examined in later years), and was 

unrelated to litter biomass.

Table 11.1—Site characteristics and percentage of N retained 3 years after adding 15N (25 kg N ha-1 yr-1) 

to fi ve prairie communities across the Great Plains. The bulk of the retention was in fi ne-textured soil 

with greater organic C content, with less N retention in plant biomass (average of 25 percent in plants 

across sites) (Barret and Burke 2002, reprinted with permission).

Site State System MAP (cm)a MAT (°C)b

Total N 

retention (%)

Comanche National 

Grassland

Colorado shortgrass 41 11.9 84

Thunder Basin 

National Grasslands

Wyoming mixed-grass 32 7.9 76

Pawnee National 

Grasslands

Colorado shortgrass 37 9.1 64

Fort Keogh Livestock 

and Range Research 

Laboratory

Montana mixed-grass 35 7.1 61

Muleshoe National 

Wildlife Refuge

Texas shortgrass 45 14.3 46

a

MAP – mean annual precipitation 

b

MAT – mean annual temperature

13Egerton-Warbuton, L.M. Unpublished data. Chicago 

Botanic Garden, 1000 Lake Cook Road, Glencoe, IL, 60022
122 Chapter 11—Great Plains GTR-NRS-80

Wedin and Tilman (1996) examined the impacts of 

12 years of N addition at a range of rates (0 to 270 kg 

N ha-1 yr-1, background wet deposition of 6 kg N ha-1

yr-1) to two Minnesota old fi elds dominated by a diverse 

mixture of prairie grasses and forbs near the ecotone 

of prairie and mixed woodland. Even low rates of N 

addition (10 and 20 kg ha-1 yr-1) decreased the percent 

abundance of C4 grasses (dominated by the native 

bunchgrass little bluestem) by nearly 20 percent, and 

increased the abundance of C3 grasses (primarily the 

invasive rhizomatous grasses Kentucky bluegrass [Poa 

pratensis] and quackgrass [Agropyron repens]) (Wedin and 

Tilman 1996). Knops and Reinhard (2000) found that 

the specifi c leaf area for all three of these grasses tended 

to increase with N across the N addition gradient. 

However, quackgrass specifi c leaf area increased the 

most at the lowest N addition rate (10 kg N ha-1 yr-1

plus 6 kg N ha-1 yr-1 background wet deposition) and 

demonstrated the largest proportional increase on 

average (86 percent for quackgrass as opposed to 40 

percent and 28 percent for little bluestem and Kentucky 

bluegrass respectively) (Knops and Reinhart 2000). Th is 

pattern suggests plasticity as a key factor infl uencing 

compositional changes. Compositional shifts toward 

more productive N-rich species decreased root and 

litter C:N ratios, increased litter mass, decreased light 

penetration, and decreased species richness (Tilman 

1993, Wedin and Tilman 1996). Th ough these changes 

were not readily apparent after the fi rst few years (Clark 

and Tilman 2008, Tilman 1987), the losses of species 

and shifts in community structure increased through 

time, especially at lower N input levels, with a 17 

percent reduction in species number at the 10 kg N 

ha-1 yr-1 rate after 22 years (Fig. 11.2, Clark and Tilman 

2008). Tilman (1993) suggested that the loss of species 

was driven primarily by increased litter mass, reducing 

light and germination rates—a hypothesis supported by 

the weaker response found in the annually burned fi eld 

at the same research site (Tilman 1993) and from other 

related studies (Foster and Gross 1998). Other research 

suggests light limitation from competing neighbors 

drives the reduction in species (Hautier et al. 2009).

Studies using high rates of N addition generally confi rm 

the dynamics discussed above. Nitrogen addition (> 50 

kg ha-1 yr-1) to a newly restored tallgrass prairie (Baer et 

al. 2003) increased productivity and decreased species 

richness, purportedly through the same mechanism 

of reduced light and increased aboveground biomass 

and litter. Nitrogen addition at 40 kg ha-1 yr-1 to 

assembled prairie communities in Minnesota led to 

similar responses as in natural communities (increased 

biomass, increased plant N concentration, increased 

aboveground:belowground biomass ratios), and 

also led to decreased lignin:N ratios, increased root 

cellulose concentration, and short-term increases in 



10 k N h -1 -1

EHU



10 kg N ha 1 yr 1

34 kg N ha-1 yr-1

95 kg N ha-1 yr-1

HV1XPE

LRQ 

H6SHFLH

SURSRUWL



5HODWLYH

S





95 34 10

<HDU

     

Figure 11.2—Shown are the species 

numbers (relative to controls) for 

plots receiving N at three difference 

rates averaged over three Minnesota 

fi elds from 1982-2004. Dashed lines 

correspond to annual standard errors 

in control plots and arrows indicate 

the year of fi rst signifi cant (P < 0.01) 

detection of species loss for that 

treatment rate (Clark and Tilman 2008).
Chapter 11—Great Plains GTR-NRS-80 123

photosynthesis (Dijkstra et al. 2006, Knops et al. 

2007, Lee et al. 2001, Reich et al. 2001). Lee et al. 

(2001) reported that the small increases in rates of net 

photosynthesis for 13 prairie species in the fi rst year of 

treatment did not generally continue, even though leaf 

N content remained elevated. Nitrogen addition for 

8 years at 50 kg ha-1 yr-1 to a restored tallgrass prairie 

community increased annual net primary productivity, 

root N, and root tissue quality, but root C:N ratios 

remained high enough to prevent net mineralization 

of N (Baer and Blair 2008). Similar to the long-term 

studies in Minnesota (Clark et al. 2009), there was little 

evidence that N enrichment increased root biomass, soil 

C or N accrual rates, or storage of C in total, microbial, 

or mineralizable pools within this time frame (Baer and 

Blair 2008).

Higher rates of N addition (100 kg ha-1 yr-1) have 

been studied extensively in many sites, with qualitatively 

similar results. One notable study is from the tallgrass 

prairie, that added high rates of N addition in 

combination with burning and mowing. Th ey found 

that N addition greatly reduces species richness in 

combination with burning (especially of the diverse C3

forb group), a decline that is mitigated with additional 

mowing (Collins et al. 1998).

11.3.3 Other trophic responses in the Great Plains

Nitrogen addition for 2 years (16.3 kg ha-1 yr-1, 

background total deposition of 11 kg N ha-1 yr-1) to a 

mixed-grass prairie in Oklahoma had variable eff ects 

on mammalian herbivores (Clark et al. 2003, Clark 

et al. 2005, Jorgensen et al. 2005). Nitrogen addition 

increased the abundance of one species of harvest 

mouse (Reithrodontomys montanus) primarily during 

the winter, had no eff ect on a second species of harvest 

mouse (Reithrodontomys fulvescens), and increased the 

abundance of cotton rats (Sigmodon hispidus) only in 

combination with reduced predation (i.e., fencing) 

(Clark et al. 2003, Jorgensen et al. 2005).

Haddad et al. (2000) examined the impacts on insect 

communities from fertilizer addition at a range of 

rates to a Minnesota old-fi eld (0 to 270 kg N ha-1 yr-1; 

background wet deposition 6 kg N ha-1 yr-1). Nitrogen 

addition at low rates (≤20 kg ha-1 yr-1) for 14 years 

decreased insect species diversity and increased insect 

numbers and biovolume. Th is resulted from a decrease 

in the diversity of food sources and oviposition sites 

as plant species richness declined and two species 

(quackgrass and Kentucky bluegrass) came to dominate. 

Changes in the number of insect individuals at low N 

addition rates (≤20 kg ha-1 yr-1) also varied by functional 

type, with decreases in parasitoids and increases in 

herbivores (esp. dominant herbivores) (Haddad et al. 

2000). Another study used a detailed biogeochemical 

process model to examine the eff ect of N deposition (1 

to 5 kg N ha-1 yr-1) and insect herbivory on ecosystem 

level C and N dynamics (Th roop et al. 2004). 

Researchers found that although total C storage was less 

sensitive to herbivory than to N deposition, herbivory 

reduced production and soil organic C more as N 

deposition increased, infl uencing whether ecosystem C 

sequestration saturated at lower or higher N deposition 

rates (Fig. 11.3) (Th roop et al. 2004). Short-term and 

long-term trophic responses to N deposition are likely 

to diff er, and they remain understudied (Th roop and 

Lerdau 2004). Over the short term, N addition increases 

plant quantity and quality without large changes in 

composition, which should increase insect populations. 

But over longer periods, compositional changes can 

increase or decrease herbivore activity, depending 

on whether there is a shift toward or away from a 

community dominated by herbivore-preferred species 

(Th roop and Lerdau 2004).

Studies using higher rates of N addition than those 

discussed above generally confi rm the aforementioned 

dynamics, with some notable exceptions. In particular, 

Lau et al. (2008) found no eff ect of N addition (40 kg 

ha-1 yr-1) on herbivory of a common N-fi xing prairie 

legume (roundhead lespedeza [Lespedeza capitata]). 

Th is suggests that higher trophic responses to N 

deposition are likely to depend on specifi c plantherbivore interactions. Nonetheless, a review of studies 

on the eff ect of high rates of N addition on insect 

herbivory found a generally stimulatory eff ect through 

increased foliar N and decreased secondary compounds 

(Th roop and Lerdau 2004). Macroinvertebrates in the 

tallgrass prairie of Kansas yielded complex treatment 

responses between N addition (100 kg ha-1 yr-1) and 

other treatments (burning, mowing, and P addition). 
124 Chapter 11—Great Plains GTR-NRS-80

Generally, exotic earthworms and herbivorous beetle 

larvae tended to increase with fertilization, while cicadas 

were relatively unresponsive (Callaham et al. 2002, 

2003). Nitrogen addition (80 kg ha-1 yr-1) to a tallgrass 

prairie in Texas led to a decrease in arthropod richness 

as a result of an increase in nonpreferred woody species 

(Hartley et al. 2007). One of the few long-term studies 

to date of small mammal responses found a long-term 

decrease in the population of voles after 9 years of 

high-N treatment (336 kg ha-1 yr-1) to an Ohio old-fi eld 

(Hall et al. 1991).Th e decrease in the vole population 

was attributed to a shift in the plant population 

from edible species (70 percent edible species such as 

Canada goldenrod [Solidago canadensis] and clover 

[Trifolium spp.] in control plots) to inedible species (63 

to 99 percent inedible species such as annual ragweed

[Ambrosia artemisiifolia] and great ragweed [Ambrosia 

trifi da] in fertilized plots).

11.4 Critical Loads Estimates

We estimate the critical loads for shifts in ecosystem 

structure and function at 10 to 25 kg N ha-1 yr-1 for 

mixed- and shortgrass prairie and 5 to 15 kg N ha-1

yr-1 for tallgrass prairie. Th ese overall estimates are 

based on observed shifts in several key biogeochemical, 

plant, and trophic indicators following N inputs just 

above these levels (Table 11.2, 11.3). Total N input at 

27.3 kg ha-1 yr-1 (16.3 kg N ha-1 yr-1 treated and 11.0 

kg N ha-1 yr-1 of ambient N deposition) to a mixedgrass prairie site increased plant community biomass, 

increased NO3

-

 leaching, and had variable eff ects on 

higher trophic levels, suggesting that the N critical load 

must be below this level. Th us, we set the critical load 

at 10 to 25 kg N ha-1 yr-1. It is diffi cult to assign a lower 

value to this range, as we do not have a treatment level 

for which there was no negative eff ect, nor do we know 

whether ambient N deposition has caused alterations in 

the ecosystem from a prior unimpacted condition. We 

regard this estimate as fairly reliable rather than reliable 

because of this uncertainty and because it is based on 

the only known empirical study examining the response 

of mixed-grass prairie to such low levels of N. We use 

the same estimate for the shortgrass prairie (10 to 25 

kg N ha-1 yr-1) because we expect these ecosystems to 

respond similarly and we have no other empirical studies 

to consult. Th is reduces the reliability for the shortgrass 

prairie critical loads to expert judgment. Th e critical 

load for shortgrass prairie may be slightly higher or 

co-dependent on water availability, as this system is less 

sensitive to high rates of N addition (100 kg ha-1 yr-1) 

than the tallgrass prairie (Clark et al. 2007), and likely 

has a greater tendency for water limitation or colimitation of water and N (Hooper and Johnson 1999, 

Lauenroth et al. 1978). However, we urge caution due 

to the lack of studies and suggest using the 10 to 25 kg 

Figure 11.3—Infl uences of N deposition and the four 

herbivore functions on C dynamics. The four functions 

describe whether the intensity of herbivory is independent 

of the rate of N deposition (constant), increases linearly 

with N deposition (positive), decreases linearly with N 

deposition (negative), increases multiplicatively with N 

deposition (population), or whether herbivory is absent (no 

herbivory). Values are from the fi nal year of a simulation run 

that included a 2500-year climate stabilization, 30 years of 

N deposition, and 30 years of herbivory and N deposition. 

(a) Total annual biomass production (sum of above-and 

belowground production) in terms of C accumulation in g 

m-2 yr-1; (b) Total pools of soil organic C (Throop et al. 2004, 

reprinted with permission of John Wiley and Sons).

Nitrogen deposition (g N m-2 yr-1)

012345

Soil organic carbon (g C m-2) 3000

4000

5000

6000

7000

8000

Nitrogen deposition (g N m-2 yr-1)

012345

Total production (g C m-2 yr-1) 200

300

400

500

600

700

800

no herbivory 

constant

positive

negative

population 

E

a.
Chapter 11—Great Plains GTR-NRS-80 125

Table 11.2—Responses to low rates of N addition (N-add; ≤25 kg N ha-1 yr-1) for the Great Plains ecoregion. Rates of N 

deposition (N-dep) are shown for comparison.

Ecosystem component

Site N input

kg N ha-1 yr-1

N-add/N-dep

Indicator Measured Response Study

Ecotone (tallgrass prairie and oak woodland)

Cedar Creek Ecosystem Science Reserve LTER, MN 

10-20/6a Biogeochemical 

and microbial

Elevated net N mineralization rates after 22 years, not after 12 years

Clark et al. 2009, Wedin and Tilman 1996

Ecotone (tallgrass prairie and oak woodland)

Cedar Creek Ecosystem Science Reserve LTER, MN

10-20/6a Plant community Increased biomass, C3 (cool season grass) 

abundance, litter mass, decreased C4 (warm 

season grass) biomass, light penetration, tissue C:N, and species richness

Clark and Tilman 2008, Wedin and Tilman 1996, Tilman 1993, Tilman 1987

Ecotone (tallgrass prairie and oak woodland)

Cedar Creek Ecosystem Science Reserve LTER, MN

10-20/6a Higher trophic 

levels

Decreased insect diversity, and number of parasitoids, increase in total numbers of insects and number of herbivorous insects 

Haddad et al. 2000

Shortgrass prairie (3) Short-grass steppe 

LTER, CO

10/3a Biogeochemical Retention of 15N in organic soil layers more 

than in plants after 3 years

Epstein et al. 2001

Arbuscular mycorrhizal fungi

Chicago, IL area N gradient

12a microbial Decline in both mycorrhizal colonization and 

spore density

Egerton-Warburtonc

Mixed-grass prairie Center for Ecological 

Assessment Research, OK

16.3/11b Biogeochemical 

and microbial

Increased soil NO3

- and leaching, inconclusive 

effects on decomposition, no effect on denitrifi cation

Jorgensen et. al. 2005

Mixed-grass prairie Center for Ecological 

Assessment Research, OK

16.3/11b Plant community Increased plant biomass and tall fescue cover, 

decreased richness, weaker effects on cover of other species

Jorgensen et. al. 2005

Mixed-grass prairie Center for Ecological 

Assessment Research, OK

16.3/11b Higher trophic 

levels

Increases of harvest mouse (Reithrodontomys montanus) in winter, no effects on other mammals

Jorgensen et. al. 2005

Mixed-grass (2) and shortgrass (3) prairies

Various (Table 11.1) 25/1-4a Biogeochemical Retention of 15N in organic soil layers more 

than in plants

Barret and Burke 2002

aLocally measured wet depositionbTotal (wet+dry) N depositioncSee footnote 13 on page 121
126 Chapter 11—Great Plains GTR-NRS-80

N ha-1 yr-1 level as the critical load. Total N input at 16 

kg N ha-1 yr-1 (10 kg N ha-1 yr-1 treated and 6 kg N ha-1

yr-1 ambient N deposition) to a northeastern tallgrass 

prairie site led to reductions in plant richness, shifts 

in biogeochemical cycling, and changes in the insect 

community, suggesting that the N critical load must 

be below this level. Researchers at this site estimated 

the critical threshold to these old fi elds at 5.3 kg N 

ha-1 yr-1. However, due to the extremely low N status 

of this site, resulting from its geomorphology and past 

use for agriculture, we think this estimate is likely low 

for the tallgrass prairie as a whole. Th us, we estimate 

the critical load for tallgrass prairie at 5 to 15 kg N ha-1

yr-1. We regard this estimate as fairly reliable rather 

than reliable because the only available study also added 

non-N resources, and because the site is nutrient poor 

compared to other tallgrass prairie sites (Grigal et al. 

1974).

Based on research by Egerton-Warburton13 and 

NADP wet deposition, we suggest a provisional N 

critical load of 12 kg ha-1 yr-1 for mycorrhizal eff ects 

in this ecoregion. However, the lack of sites with 

lower N deposition in the gradient studied may give 

us an unrealistically high value for the critical load, 

as communities in the low deposition sites may have 

already been altered. Clearly, additional factors such as 

soil type, fi re history, and agricultural and other land-use 

history will have a large infl uence on the N status of the 

Great Plains grassland ecosystem types, and thus on the 

responsiveness of mycorrhizal fungi to N deposition.

11.5 Comparison to Critical Loads for 

Other Regions

Th e European ecosystems most comparable to the 

U.S. Great Plains are the EUNIS (European Nature 

Information System) coded habitats E1.2 (perennial 

calcareous grassland and basic steppes) and E1.7 

(non-Mediterranean dry acid and neutral grassland) 

(European Environmental Agency 2010). Bobbink et al. 

(2003) estimated a moderately reliable critical load for 

perennial calcareous grassland and basic steppes at 15 to 

25 kg N ha-1 yr-1, and a reliable critical load for nonMediterranean dry acid and neutral grassland at 10 to 20 

kg N ha-1 yr-1 (Bobbink et al. 2003). Other ecosystems 

comparable to the U.S. Great Plains (especially shortand mixed-grass prairie) include the Eurasian steppe 

that stretches from Asia into Eastern Europe. Nitrogen 

addition at 17.5 kg N ha-1 yr-1 to a representative site 

in rural China led to increases in plant biomass and 

reductions in species richness (critical load less than 20 

kg N ha-1 yr-1) (Bai et al. 2010). Our estimates for the 

Great Plains are similar to those from China and slightly 

lower than those from Europe, as we have long-term 

fi eld experiments in these lower ranges, and it is likely 

that responses in Europe may be somewhat attenuated 

due to higher historical levels of N deposition (Galloway 

et al. 2004). Indeed, several observational studies across 

Table 11.3—Critical loads for nutrient N for the Great Plains ecoregion. Reliability rating: ## reliable; # fairly 

reliable; (#) expert judgment

Ecosystem 

component

Critical load for N 

deposition

kg N ha-1 yr-1

Reliability Response Comments Study

Tallgrass 

prairie

5-15 # Biogeochemical N 

cycling, plant and 

insect community 

shifts

Need more low-N 

studies without 

other nutrients

Clark et al. 2009, Clark 

and Tilman 2008, Tilman 

1993, Tilman 1987, 

Wedin and Tilman 1996

Mixed-grass 

prairie

10-25 # Soil NO3

-

 pools, 

leaching, plant 

community shifts

More studies 

needed

Clark et al. 2003, 2005; 

Jorgenson et al. 2005

Shortgrass 

prairie

10-25 (#) Inferred from 

mixed grass

Barret and Burke 2002, 

Epstein et al. 2001

Mycorrhizal 

fungi

12 (#) Decline in arbuscular 

mycorrhizal fungal 

activity

Egerton-Warburtona

a

See footnote 13 on page 121
Chapter 11—Great Plains GTR-NRS-80 127

current N deposition gradients (Stevens et al. 2004) and 

at diff erent time periods (Bennie et al. 2006, Smart et al. 

2005) suggest that current and historical N deposition 

may lower species richness across European systems, 

though more research is sorely needed.

11.6 Future Research Directions and 

Gaps in Data

Currently, native communities of the Great Plains are 

highly fragmented as a result of extensive agricultural 

activity across this large region and are threatened 

by several global-change factors in addition to N 

deposition, including continued land-use change, 

alterations of historic fi re and grazing regimes, and 

regional climate change. Although fi re and grazing are 

expected to alter ecosystem sensitivity to N deposition, 

no studies to date have examined these critical 

interactions at rates comparable to deposition levels. 

Th e few high N addition studies suggest that ecosystems 

may be maximally sensitive if fi re regimes alone are 

restored, suggesting eff ective conservation likely requires 

the re-establishment of multiple regulating factors 

(Collins et al. 1998, Suding et al. 2004). Elevated 

carbon dioxide (CO2) under future global change 

has been hypothesized to both favor and disfavor the 

currently subordinate C3 grasses in this ecoregion, 

depending on the dominant ecological response. On 

one hand, C3 grasses may be favored if the fertilization 

eff ect of CO2 predominates and water does not become 

limiting, resulting in dramatic reductions of ecosystem 

C-sequestration (Knapp et al. 1998). However, several 

other lines of research suggest that elevated CO2

 may 

actually favor the continued dominance of the C4

grasses of this system through their elevated water use 

effi ciency during dry years (Ham et al. 1995; Knapp 

et al.1993, 1996), an eff ect that is likely dependent on 

future climate patterns (Knapp et al. 1998). Each of 

these dynamics requires further investigation, but is 

beyond the scope of this report. Nevertheless, there are 

several future research directions that will improve our 

ability to predict the impacts of N deposition. Most 

importantly, we stress the need for more empirical 

studies in more geographic locations examining the long 

term (≥10 years) impacts from low rates of N input (≤20 

kg N ha-1 yr-1) comparable to N deposition over this 

region. Th is should be pursued both in isolation and 

in combination with the multiple factors that regulate 

ecosystem structure and function across the Great 

Plains. Th is experimental work, however, is not enough, 

and should be coordinated with dynamic ecosystemvegetation models to direct limited resources towards 

areas of greater sensitivity. Th e wide range of ecosystem 

responses to high rates of N addition observed (Clark et 

al. 2007) demonstrate that not all ecosystems are equally 

N sensitive. Prior research suggests that fi ne-scale as 

well as regional-scale factors likely moderate ecosystem 

responses (Gough et al. 2000). Nonetheless, there are 

several specifi c areas that would benefi t from additional 

study as they relate to N deposition:

• Shifts in phenology: A few studies in other 

systems have examined changes in the timing 

of ecological interactions with added N which 

can greatly infl uence ecosystem functioning 

(Cleland et al. 2006).

• Decomposition: Th e impact of N addition on 

decomposition is still unresolved and largely 

unstudied at low input rates (Hobbie 2008, 

Keeler 2010, Knorr et al. 2005).

• Shifts in belowground communities and 

processes: Although fungal:bacterial population 

ratios often decrease with N addition (e.g., 

Bradley et al. 2006), no studies to date have 

addressed these dynamics at low rates.

• Modulating responses to N deposition: 

Historical and altered regimes of fi re and 

herbivory are likely to modulate ecosystem 

sensitivity to N deposition. In addition, 

grasslands are known to often be co-limited 

by other resources in addition to N (primarily 

water), though it is unknown whether critical 

thresholds for N deposition will decrease or 

increase under changing climatic regimes 

(Bobbink et al. 1998).

• Higher-order responses to N deposition: Most 

of the research to date has focused on plant 

and soil responses to N deposition. However, 

higher order responses, such as N-induced shifts 

in herbivory, pathogens, and drought stress, 

as well as modulating infl uences from fi re, 
128 Chapter 11—Great Plains GTR-NRS-80

herbivory, and climatic variability, are likely to 

greatly modify the total ecosystem response to 

N deposition (Bobbink et al. 1998, Knapp et al. 

1998, Th roop and Lerdau 2004).

• Predictive modeling: As briefl y mentioned 

above, more eff ort directed to predictive 

modeling of ecosystem sensitivity to N addition 

needs to be pursued across ecosystem types and 

indicators (Heil and Bobbink 1993).

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Chapter 12—North American Deserts GTR-NRS-80 133

12 NORTH AMERICAN DESERTS

E.B. Allen, L.H. Geiser

ranges separated by broad valleys, whereas the Colorado 

and Arizona/New Mexico plateaus consist of canyons, 

cliff s, buttes, and mesas of sedimentary rock. Sagebrush 

(Artemisia spp.) dominates these deserts, with saltbush 

(Atriplex spp.) and greasewood (Sarcobatus spp.) on more 

alkaline soils.

Th e warm deserts have higher temperatures and 

greater evaporation rates. Th e Mojave is the driest and 

hottest of the North American deserts; its recorded 

high temperature, in Death Valley, California, is 57 

 ̊C. Characteristic plants are the creosote bush (Larrea 

tridentata) and Joshua tree (Yucca brevifolia). Th e 

Sonoran Desert of Arizona and California is subtropical 

and receives most of its precipitation in the summer 

monsoon season; gentler rains may occur in the winter 

months. Its structurally diverse vegetation includes 

the paloverde tree (Parkinsonia spp.), saguaro cactus 

(Cereus gigantea), cholla (Cylindropuntia spp.), and 

agave (Agave spp.). Th e Chihuahuan Desert extends 

into southern New Mexico from Mexico. Rain occurs 

primarily in the summer, and supports smaller scrub 

vegetation characterized by American tarwort (tarbush; 

Flourensia cernua), creosote bush (Larrea tridentata), 

12.1 Ecoregion Description

North American deserts and semi-deserts extend 

from British Columbia to Mexico (CEC 1997). Th e 

ecoregion description is adapted from CEC (1997). Th ey 

are characterized by aridity (130 to 380 mm annual 

precipitation); shrub and succulent vegetation with trees 

in riparian areas and higher elevation woodlands; and 

dry, low organic matter soils that are high in calcium 

carbonate (CaCO3) in some regions. Th e ecoregion can 

be subdivided into cold northern deserts and semideserts (Columbia Plateau, Snake River Basin, Great 

Basin, greater Colorado Plateau) and warm southern 

deserts (Mojave, Sonoran, Chihuahuan) (see Table 12.1).

Th e cold deserts and semi-deserts are arid to semi-arid 

with marked seasonal temperature extremes, a winter 

moisture regime and some snow. Th ey lie in the rain 

shadow of the Cascade and Sierra Mountains to the 

west, and are blocked from moist Gulf Coast air masses 

by the Rocky Mountains to the east. Th e Columbia/

Snake River Plateaus are loess- and volcanic ash-covered 

plains; their expansive grassland and sagebrush steppes 

have been largely converted to agriculture. Th e Great 

Basin contains hundreds of north-south mountain 

Table 12.1—Location of North American Deserts of the United States.

Level 2 

Ecoregiona

Desert Location Level 3 Ecoregiona

Cold deserts Columbia Plateau Eastern Washington Columbia Plateau

Snake River Plain Southern Idaho Snake River Plain

Great Basin Southern Oregon, extreme 

southern Idaho, north and central 

Nevada, western Utah

Northern Basin and Range, Central 

Basin and Range

Greater Colorado 

Plateau 

Western Wyoming, western 

Colorado, eastern Utah, northern 

Arizona, northern New Mexico

Wyoming Basin, Colorado Plateau, 

Arizona/New Mexico Plateau

Warm deserts Mojave Southeastern California, southern 

Nevada, northeastern Arizona

Mojave Basin and Range

Sonoran Southwestern Arizona, extreme 

southern California

Sonoran Basin and Range

Chihuahuan Southern New Mexico Chihuahuan Desert

a

CEC 1997
134 Chapter 12—North American Deserts GTR-NRS-80

soaptree yucca (Yucca elata), grama grasses (Bouteloua

spp.), and tobosagrass (Pleuraphis mutica). Piñon-juniper 

woodlands occur at the higher elevations above both 

cold and warm deserts.

Small mammals, birds, and reptiles are the most 

abundant wildlife of the ecoregion. Livestock grazing 

is widespread and, together with irrigated agriculture 

and population growth, has had signifi cant ecological 

and hydrological eff ects. Major urban areas with higher 

nitrogen (N) deposition, which aff ects the adjacent and 

downwind deserts, are Phoenix, Salt Lake City, Las 

Vegas, Tucson, Albuquerque, Spokane, and Los Angeles 

(Fenn et al. 2003).

12.2 Ecosystem Responses to N 

Deposition

Responses to N deposition in deserts include alterations 

of species composition, often increases in biomass of 

exotic species, and decreases in native species. In this 

ecoregion of altitudinal, latitudinal, and landform 

diversity, vegetation types vary, but low growing shrubs 

and grasses predominate. Recently, invasive exotic 

species have been increasing in desert ecosystems (Allen 

et al. 2009, Brooks 2003). Th is may be due to increased 

N deposition, as exotic grasses may respond to higher 

soil N levels diff erently than native species (Allen et 

al. 2009). Such changes can have serious management 

implications. For example, in the agriculturally intensive 

Snake River Plain and in the Great Basin, extensive 

cheatgrass (Bromus tectorum) invasions contribute to 

increased fi re frequency, that in turn favors even greater 

cover of cheatgrass. Loss of native plants adapted to 

longer fi re intervals and the poorer nutritional quality 

of cheatgrass in turn reduces the carrying capacity of 

lands (Whisenant 1990). Fire frequency has similarly 

increased in the Mojave Desert in areas undergoing exotic 

grass invasions (Brooks et al. 2004, Brooks and Matchett 

2006). Experimental N fertilization resulted in a depletion 

of soil moisture in sagebrush steppe, which could 

potentially aff ect future plant production (Inouye 2006).

Biotic soil crusts (consisting of bacteria, green algae, 

microfungi, and crustose lichens) and mycorrhizal 

fungi constitute important parts of desert ecosystems. 

Responses to increased N deposition include reductions 

in photosynthesis and nitrogen fi xation rates; reductions 

in photosynthetic pigments and ultraviolet (UV) 

protective pigments such as scytonemin; increases 

in electrolyte leakage, increases in microbial N and 

denitrifi cation; shifts in community composition 

favoring eutrophs; and increases in microbial activity in 

nonvegetated soils.

12.3 Range of Responses Observed

Several studies have used N fertilizer in the desert to 

measure a biomass response and then compared this 

response to N inputs along N deposition gradients 

(Allen et al. 2009, Baez et al. 2007). Responses to N 

inputs are summarized in Table 12.2. A response to a 

certain level of N fertilization is assumed to be similar 

to the response along the gradient, assuming that factors 

besides N inputs are equal. Th e diffi culty with this 

approach is that gradients may have multiple factors that 

change among sites in addition to N inputs, making 

interpretation of results diffi cult. Th e combination of 

N fertilization plus gradient approaches provide the 

most reliable information currently available. Th e most 

useful N fertilization studies for setting critical loads 

are those in the range of threshold responses, as reported 

below. For instance, N fertilizer of 72 kg ha-1 yr-1 at the 

Jornada Long-Term Ecological Research (LTER) site in 

New Mexico showed growth increases by native winter 

annual forbs and decreases by summer annual forbs 

(Gutierrez and Whitford 1987), but this value is well 

above the response threshold determined from studies 

reviewed below. Similarly, 60 kg N ha-1 yr-1 increased 

mineralization rates at the Central Arizona Project 

(CAP-LTER; Hall et al. 2009). Alternatively, input of 

40 kg N ha-1 yr-1 over 3 years did not increase creosote 

bush biomass at the Mojave Global Change Facility 

near Las Vegas (Barker et al. 2006) and the authors 

concluded that annual variability in precipitation 

was the most important variable for creosote bush 

productivity.

12.3.1 Plant Community Responses

Baez et al. (2007) used an increase in the biomass of the 

dominant native blue grama grass (Bouteloua gracilis) 

as an indicator of elevated N in desert grassland at the 
Chapter 12—North American Deserts GTR-NRS-80 135

Table 12.2—Responses to N input for North American Deserts ecoregion. Reliability rating: ## reliable; # fairly reliable; (#) expert judgmentSite Response 

threshold

kg N ha-1 yr-1

Reliability Response Comments Study

Riparian habitat, Hells Canyon NRA, Columbia Plateau

3 (#) High cover of eutrophic 

lichens, enhanced N concentrations in lichen thalli

Uncertainty regarding modeled N estimate

Geiser et al. 2008, Porter et al 2007

Piñon-juniper and creosote scrub,Joshua Tree National Park, Sonoran and Mojave Desert

8.4 # Increased biomass of invasive 

grasses

N fertilization of 5 kg ha-1 y-1 plus background N deposition of 3 kg ha-1 y-1 in wet year

Allen et al. 2009

Creosote scrub,Joshua Tree National Park, Sonoran Desert

3-8 # Production of fi ne fuel of 

Mediterranean splitgrass suffi cient for fi re

DayCent model Rao et al. 2010 

Piñon-juniper Joshua Tree National Park, Mojave Desert

3-6 # Production of fi ne fuel of red 

brome suffi cient for fi re

DayCent model Rao et al. 2010

Desert grasslandSevilleta LTER

2.5-22 # Increase in blue grama 

biomass

Response observed at 20 kg N ha-1 y-1 fertilization; ambient is 2.45 kg N ha-1 y-1

Baez et al. 2007

Sagebrush steppe 7.4 # Depletion of soil moisture No increase in aboveground 

production

Inouye 2006

Desert grassland, Colorado Plateau

22 Not reliable: 

fertilization level 

too high

Increase in tiller density of the dominant cool season grass species Indian ricegrass 

No increase in warm season galleta grass.Fertilizer addition of 20 kg N ha-1 y-1; ambient deposition 2 kg N ha-1 y-1

Schwinning et al. 2005

42 Not reliable: 

fertilization level 

too high

Increase in biomass of the invasive species Russian thistle 

Short term—2 yrs fertilization

Desert scrub,Mohave Desert

32 Not reliable: 

fertilization level 

too high

Increase in exotic invasive grasses and a decline in native forbs

Creosote bush dominated ecosystem

Brooks et al. 2003
136 Chapter 12—North American Deserts GTR-NRS-80

Sevilleta LTER site in New Mexico. Plots fertilized with 

20 kg N ha-1 for one season had increased biomass of 

blue grama compared to controls. Bulk N deposition 

increased from 1.7 to 2.4 kg ha-1 yr-1 in the control plots 

between 1989 to 2004, adding an additional total of 

5.9 kg N ha-1 to the control plots from increased N 

deposition during these 16 years. However, there was 

no change in biomass of blue grama over this period, 

suggesting that 5.9 kg N ha-1 over 16 years was below 

the threshold for increased biomass in this species. Th us, 

Baez et al. (2007) concluded that the threshold for 

elevated biomass of blue grama lies somewhere between 

2.45 and 20 kg N ha-1 yr-1; including ambient deposition 

raises the upper response threshold to approximately 22 

kg N ha-1 yr-1.

Zhu et al. (2006) calculated 72.3 kg ha-1 N content 

in urban Phoenix soils compared to exurban desert 

soils that had 9.4 kg N ha-1, accompanied by elevated 

carbon (C) in urban soils. Th is increase in soil N and 

C was hypothesized to have occurred during the last 

several decades of rapid urban expansion. A more recent 

study also showed elevated N and C in remnant native 

vegetation fragments within the urban Phoenix CAPLTER site that promoted changes in N cycling, such as 

increased denitrifi cation potential and higher microbial 

N (Hall et al. 2009). Nitrogen deposition in the 

Phoenix area was measured using wet and dry bucket 

collectors. Th e highest measured values were less than 

6 kg N ha-1 yr-1, which the investigators describe as an 

underestimate because the dry bucket technique is not 

accurate (Lohse et al. 2008). Th e Community Multiscale 

Air Quality (CMAQ) model, which includes both wet 

and dry deposition, showed a maximum deposition 

of 26 kg N ha-1 yr-1 for the Phoenix area (Fenn et al. 

2003). Impacts of N deposition on vegetation were not 

reported in these studies in the Phoenix area, although 

the PALS (Patch Arid Land Simulator) model showed 

nearly a doubling in total plant production with 

simulated N addition between 2 and 26 kg ha-1 yr-1

(Shen et al. 2008). Th e model shows that even a small 

increase in N can increase biomass, but it is not clear 

how this can be used to determine a response threshold. 

However, measurements comparing urban to rural areas 

show no signifi cant change in plant biomass due to 

elevated N, likely because of multivariate gradients that 

include elevated ozone (O3

), temperature, and other 

factors.14 Additional research on responses to fertilizer N 

are under way at the CAP-LTER site.

Schwinning et al. (2005) used fertilizer levels of 0, 10, 

20, and 40 kg N ha-1 for 2 years in a desert grassland on 

the Colorado Plateau and measured an increase in tiller 

(grass shoot) density of the dominant cool season grass 

species, Indian ricegrass (Oryzopsis hymenoides), at 20 kg 

N ha-1 yr-1. Th is increase was not observed in the warm 

season grass species, galleta grass (Hilaria jamesii). Th e 

most noticeable change in this study was an increase in 

biomass of the invasive species Russian thistle (Salsola 

iberica) at the highest N level. Th e authors concluded 

that elevated N deposition will promote an invasion of 

Russian thistle and a shift in dominance to cool season 

grasses. However, current N deposition at this site may 

be as low as 2 kg ha-1 yr-1. Because the study was carried 

out for only 2 years, the levels of N inputs may be 

higher than threshold response levels for long-term N 

deposition.

Working in creosote bush desert scrub, Brooks (2003) 

fertilized at a rate of 32 kg N ha-1 yr-1 for 2 years and 

found an increase in exotic invasive grasses and a decline 

in native forbs. Allen et al. (2009) studied the impacts 

of elevated N by fertilization along an N deposition 

gradient in Joshua Tree National Park in creosote bush 

scrub and piñon-juniper woodland. Th e N deposition 

rates along the gradient ranged from 3.4 to 12.4 kg N 

ha-1 yr-1 as measured by bulk deposition samplers and 

also as modeled (Tonnesen et al. 2007). Fertilization 

rates were 5 and 30 kg N ha-1 between 2003 and 2005. 

Th ere were no vegetation biomass responses in 2003, a 

dry year. In 2004, a year with moderate precipitation, 

two of the four sites had increased invasive grass biomass 

at the treatment of 30 kg N ha-1

, but not with 5 kg N 

ha-1. Th e strongest response to N occurred in the wettest 

year, 2005, when invasive grass biomass increased 

under both 5 and 30 kg N ha-1 fertilizer in the lowest 

and highest N deposition sites. Conversely, native 

14Hall, S.J. Unpublished data comparing plant response to N 

deposition in urban and rural areas. On fi le with: Department 

of Ecology, Evolution, and Environmental Science, School of 

Life Sciences, Arizona State University, Tempe, AZ 85287-

4701
Chapter 12—North American Deserts GTR-NRS-80 137

forbs decreased in production at these sites following 

fertilization. However, another site with low N 

deposition and sparse invasive grass cover had a positive 

response by native forbs to N fertilization, suggesting 

that native forbs can respond to N in the absence of 

grasses (Allen et al. 2009).

To determine the N input level that results in vegetation 

change, the biogeochemical model daily Century 

(DayCent; Rao et al. 2010) was used to model invasive 

grass and forb productivity under varying levels of N 

deposition, soil texture, and precipitation between 2003 

and 2008 (Rao et al. 2009, 2010). Fires have become 

more frequent in the Mojave and Sonoran Deserts, and 

there may be a relationship between fi re and increased 

grass biomass under enhanced N deposition (Allen et 

al. 2009, Brooks 2003, Brooks and Matchett 2006). 

Fire risk, which was calculated as the probability of 

producing enough fi ne fuels to carry a fi re (1 T ha-1

fi ne fuel), increased exponentially when deposition was 

above 3.0 kg N ha-1yr-1; fi re risk leveled at 5.7 kg N ha-1 

yr-1 in piñon-juniper woodland, and at 8.2 kg N ha-1 yr-1

in creosote bush scrub (Rao et al. 2010).

Elevated N may also increase soil water use by plants, as 

reported by Inouye (2006) for big sagebrush (Artemisia 

tridentata) fertilized at 6 and 12 kg N ha-1 yr-1 for six 

years in Great Basin semi-desert in southern Idaho. Soil 

moisture was depleted to 140 cm and was not recharged 

except in the wettest years. No aboveground increase in 

productivity was measured, so the moisture depletion 

may have been due to increased root productivity. 

Th e site receives 1.4 kg N ha-1 yr-1 wet deposition as 

measured at a nearby NADP (National Atmospheric 

Deposition Program) station, but dry deposition is 

not known. Since there was no signifi cant diff erence 

between the two fertilizer levels, the lower level of 6 kg 

N ha-1 yr-1 (plus the background 1.4 kg N ha-1 yr-1 wet 

deposition) can be considered the response threshold 

for moisture depletion by sagebrush. Schwinning et al. 

(2005) observed an earlier onset of drought stress and 

increased water use with 20 kg N ha-1 yr-1 of fertilizer 

plus the estimated background deposition of 2 kg N ha-1 

yr-1, confi rming the observations that elevated N can 

promote increased water use in desert vegetation.

12.3.2 Arbuscular Mycorrhizal Fungi

Mycorrhizal fungi are important for plant nutrient 

acquisition, drought stress tolerance, and the hydraulic 

lifting of water, which are critical in arid lands (Allen 

2007). Th eir diversity, abundance, and functioning are 

altered by N deposition. A Phoenix-area study showed 

shifts in mycorrhizal species composition in urban 

and exurban sites that were dominated by diff erent 

plant species, but did not relate these species shifts to 

N deposition (Bills and Stutz 2009). Data from the 

Sevilleta LTER site, which is located at the junction 

of cold desert, warm desert, and shortgrass steppe, 

suggest that long-term N fertilization at 100 kg ha-1 

yr-1 can lead to changes in the community structure 

and function of arbuscular mycorrhizal fungi associated 

with grasses from semi-arid environments (Corkidi et 

al. 2002, Egerton-Warburton et al. 2007). However, 

this level of N fertilization is well beyond the critical 

load, greatly exceeds current deposition amounts (wet 

plus dry), and exceeds the threshold for changes in 

mycorrhizal functioning observed in Mediterranean 

California ecosystems. Results of mycorrhizal response 

to N deposition in the Great Plains and Mediterranean 

ecoregions (Chapters 11 and 13) also have relevance 

for the more mesic end of the cold and warm desert 

ecosystem types.

12.3.3 Lichens, Biotic Soil Crusts, and Soil Microbes

Lichens make a signifi cant contribution to biological 

diversity in North American deserts and semi-deserts; 

a recent compendium documented 1971 species from 

the greater Sonoran Desert alone (Nash et al. 2002, 

2004, 2008). However, the long periods of metabolic 

inactivity due to insuffi cient hydration, dominance of 

communities by crustose forms with minimal surface 

exposure, minimal dry deposition inputs from canopy 

drip, and lower growth rates, can make it diffi cult to 

detect community-level shifts. Even after a decade of 

exposure, Marsh and Nash (1979) found no evidence of 

community eff ects on lichens in the vicinity of the large 

Four Corners Power Plant in the Colorado Plateau. 

Nevertheless, physiological measurements indicate that 

wetted desert lichens can be as sensitive to air pollution 

as lichens from other climates. Lichens in the vicinity of 

a coal-fi red power plant in Page, Arizona, a major point 
138 Chapter 12—North American Deserts GTR-NRS-80

source of sulfur dioxide(SO2), nitrogen oxides (NOx), 

and metals, exhibited increased electrolyte leakage, 

chlorophyll degradation, and reduced N fi xation 

(Belnap 1991).

Enhanced concentrations of N in thalli of the saxicolous 

lichen, Xanthoparmelia cumberlandia, and high cover of 

eutrophic lichens on netleaf hackberry (Celtis laeviagata

var. reticulata), a small tree of riparian areas, were 

observed throughout Hells Canyon National Recreation 

Area (Geiser et al. 2008). Interagency Modeling of 

Protected Visual Elements (IMPROVE) data show that 

Hells Canyon is a regional hot spot for ammonium 

(NH4

+

) and nitrate (NO3

-

) deposition from the Snake 

River valley and possibly as far away as California. Other 

pollutants—SO2, NOx

, O3, hydrogen sulfi de (H2S), and 

fi ne particulate ammonium sulfate ((NH4)2SO4)—were 

low and comparable to other remote areas. Mean annual 

deposition of total N from CMAQ in this area between 

1990 and 1999 was 2.6 to 3.0 kg N ha-1 yr-1 (Porter 

2007), with some uncertainty surrounding the reliability 

of the deposition estimate. Signifi cantly higher thallus 

N concentrations and eutroph cover occurred within 

0.2 km of the Snake River. Ammonia (NH3) emissions 

were especially high at these sites in spring and summer 

(5 to19 ppb biweekly averages), presumably volatilizing 

from the fertilizer-laden Snake River. Average annual 

NH3 concentrations at fi ve year-long monitoring sites 

along the Snake River ranged from 1.4 to 4.7 ppb. In a 

review of European studies, Cape et al. (2009) suggested 

a long-term mean annual critical level for ammonia 

of 1 μg m-3 (1.4 ppb), based on lichen and bryophyte 

responses; local data support that estimate.

Biotic soil crusts are well developed in cold deserts and, 

in pristine areas, can account for as much surface cover 

as vascular plants (Davidson et al. 2002). Consisting of 

cyanobacteria-dominated mats of intricately enmeshed 

bacteria, green algae, microfungi, and crustose lichens, 

they play vital ecological roles by improving topsoil 

stabilization, soil fertility, surface water holding capacity, 

and seedling survival (Harper and Belnap 2001). Biotic 

soil crusts are responsive to N deposition, but because 

component species are physically inextricable from 

each other, measurements refl ect the response of the 

combined biota, along with climate, soil chemistry, and 

other environmental variables. Some examples: 

1) An aqueous solution of 0.5 g N L-1 was added 

to biotic crusts from Canyon Lands National 

Park in the Colorado Plateau (Belnap et 

al. 2008). Reductions in quantum yield, 

b-carotene, nitrogenase activity, scytonemin 

pigments, and UV protective xanthophyll 

pigments due to N alone were observed in 

early- and mid- but not late-successional crusts. 

Higher temperatures, more UV, and more 

frequent precipitation had adverse eff ects on 

all crust types. In these cases, N positively 

moderated some adverse eff ects of UV, but only 

in the late successional crusts.

2) In a gradient study in the Sonoran Desert near 

Phoenix (McCracken et al. 2008), microbes 

of wetted soils between plants responded 

(increased carbon dioxide (CO2) fl ux) readily to 

N deposition, whereas microbes of wetted soils 

under plants did not. Th e authors postulated 

that anthropogenic N could essentially decouple 

microbes from reliance on shrubs for nutrients, 

resulting in increased activity in open spaces 

between plants. In other words, whereas native 

desert microbial activity is primarily confi ned 

to zones infl uenced by plant root exudates, 

if adequate water and anthropogenic N are 

available, microbial activity can occur in soils 

outside these zones. Major shifts in microbial 

metabolism, community composition, and 

biomass in extensive bare soil surface areas are 

implied eff ects.

12.4 Critical Loads Estimates

Fertilization studies in the Sevilleta grasslands of the 

Chihuahuan Desert (Baez et al. 2007) indicate a 

critical load greater than 2.45 and less than 22 kg N 

ha-1 yr-1, based on alteration in species composition 

and biomass observed at 22 kg N ha-1 yr-1 after 1 year 

and no changes under ambient deposition (2.45 kg 

N ha-1 yr-1). Fertilization studies by Schwinning et al. 

(2005) and Brooks (2003) were not used to set the 
Chapter 12—North American Deserts GTR-NRS-80 139

critical load, as the high levels of N inputs over both 

short-term experiments could not accurately determine 

the critical load for long-term N deposition. At Joshua 

Tree National Park in the Mojave Desert, Allen et al. 

(2009) found that vegetation response was determined 

by a combination of factors, including soil properties 

and precipitation. Th ese experimental studies indicated 

that a critical load for invasive grass biomass response 

may be as low as 8.4 kg N ha-1 yr-1 (3.4 kg N ha-1 yr-1

deposited, plus 5 kg N ha-1 yr-1 fertilized), as measured 

in the cleanest site (Allen et al. 2009). Th eir results 

suggest a critical load of 8.4 kg N ha-1 yr-1 in wet years 

to prevent an increase of invasive grasses in the desert 

ecosystem (Allen et al. 2009). Th e results from the 

biogeochemical DayCent model compared well with 

the experimental fertilization studies, and suggested 

a critical load of 3.1 to 8.2 kg N ha-1 yr-1 for lower 

elevation desert dominated by the invasive common 

Mediterranean grass (Schismus barbatus), and a critical 

load of 3.0 to 5.7 kg N ha-1 yr-1 in higher elevations with 

red brome (Bromus rubens) as the invasive grass (Rao et 

al. 2010). Th e model shows that at levels higher than 

5.7 to 8.1 kg N ha-1yr-1 the fi re risk is controlled by 

precipitation, i.e., these values exceed the critical load 

for N. Th e critical load was determined by the amount 

of N modeled to produce 1 T ha-1 of fi ne fuel to carry a 

fi re, as these are regions subject to increasing grass-fi re 

cycle impacts (Brooks and Matchett 2006). Th e critical 

load for vegetation response to N deposition is set at 

3.0 to 8.4 kg ha-1 yr-1, based on the research by Allen 

et al. (2009) and Rao et al. (2010), and supported by 

evidence (Table 12.3) of soil moisture depletion in that 

range (Inouye 2006). Th is critical load is considered 

fairly reliable; the empirical study (Allen et al. 2009) 

is supported by a biogeochemical model (Rao et al. 

2010). Further research in various desert study locations 

would increase the reliability of this critical load for the 

North American Deserts ecoregion. No critical load 

specifi c to arbuscular mycorrhizal fungi can be estimated 

currently. A tentative estimate of 3.0 kg N ha-1 yr-1 based 

on data from Geiser et al. (2008) and Porter (2007) is 

suggested for lichen thallus N concentration and lichen 

community shifts in the Columbia Plateau. Th is value 

is considered expert judgment, as there was uncertainty 

associated with the deposition estimate.

12.5 Comparison to Critical Loads for 

Other Regions

No critical loads have been calculated for Europe, as 

desert regions do not occur in Europe. We are not aware 

of critical loads estimated for other continents with 

desert systems. Th e European critical level for lichen 

and bryophyte response to NH3

 (1.0 μg m-3 or 1.4 ppb; 

Cape et al. 2009) appears to be consistent with increases 

in cover of eutrophic lichens and elevated N in lichen 

thalli observed in the Hells Canyon National Recreation 

Area in the Columbia Plateau, where annual average 

NH3 was 1.4 to 4.7 ppb (Geiser et al. 2008).

12.6 Future Research Directions and 

Gaps in Data

More research is needed to determine the range of 

critical loads in the North American Deserts ecoregion. 

Finer resolution fertilization studies could more 

closely pinpoint critical loads, as could research along 

Table 12.3—Empirical critical loads of nutrient N for North American Desert ecoregion. 

Reliability rating: ## reliable; # fairly reliable; (#) expert judgment

Ecosystem 

component

Critical load for 

N deposition 

kg N ha-1 yr-1

Reliability Response Comments Study

Lichens 3 (#) Lichen community shifts, 

increase in thallus N 

concentration

Uncertainty regarding 

modeled N estimate

Geiser et al. 2008, 

Porter 2007

Shrubland, 

woodland, and 

desert grassland

3-8.4 # Vegetation response, 

community change. 

Increased biomass 

of invasive grasses; 

decrease of native forbs

Allen et al. 2009, 

Inouye 2006, Rao 

et al. 2010
140 Chapter 12—North American Deserts GTR-NRS-80

N deposition gradients. In addition, current research 

does not encompass the full range of diversity in 

North American desert ecosystems or fl ora. Only the 

far western section of the Sonoran Desert ecosystem 

downwind of Los Angeles has been studied. Th is is a 

region that contains great biological diversity and which 

is also subject to high rates of increased N deposition 

around the urban area of Phoenix. In addition, future 

research should examine the eff ects of N deposition 

on ecosystem components with longer generation 

times, including tree and cactus species. Long-term 

observations are under way in some of the fertilizer 

studies cited (Allen et al 2009).

12.6.1 Lichens and biotic crusts

Establishing critical loads is challenging for epiphytic 

lichens, biotic crusts, and other cryptogamic 

communities of the North American Deserts compared 

to other ecoregions, due to slow growth rates, the 

complex ecology of biotic crusts, and interactions of N 

deposition with changing climate and soil chemistry. 

Further research is needed to link accurately measured 

or modeled estimates of deposition with physiological 

data for indicator species, threshold N concentrations, 

or with lichen community responses. Studies that 

measure physiological responses to fertilization may hold 

the most promise.

LITERATURE CITED

Allen, E.B.; Rao, L.E.; Steers, R.J.; Bytnerowicz, A.; 

Fenn, M.E. 2009. Impacts of atmospheric nitrogen 

deposition on vegetation and soils in Joshua Tree 

National Park. In: Webb, R.H.; Fenstermaker, L.F.; 

Heaton, J.S.; Hughson, D.L.; McDonald, E.V.; 

Miller, D.M., eds. Th e Mojave Desert: Ecosystem 

processes and sustainability. Las Vegas: University of 

Nevada Press: 78-100.

Allen, M.F. 2007. Mycorrhizal fungi: highways for 

water and nutrients in arid soils. Vadose Zone. 6: 

291-297.

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nitrogen addition on soil moisture in sagebrush 

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Chapter 13—Mediterranean California GTR-NRS-80 143

13 MEDITERRANEAN CALIFORNIA

M.E. Fenn, E.B. Allen, L.H. Geiser

mixes depending on elevation and site conditions, but 

are typically characterized by dense mixed-aged stands 

composed of coniferous (pine [Pinus spp.] and often fi r 

[Abies spp.] and incense cedar [Calocedrus decurrrens]) 

and broadleaved species (commonly oaks).

13.2 Coastal Sage Scrub

13.2.1 Ecosystem Description

Coastal sage scrub is a semi-deciduous shrubland that 

occurs in the Mediterranean-type climate of southern 

and central coastal California, extending southward to 

Baja California, Mexico. Th e United States portion of 

coastal sage scrub covers some 684,000 ha (CDFFP 

2007). Coastal sage scrub is very diverse. Unfortunately, 

it is subject to extensive human impacts in rapidly 

developing coastal California. Dominant species 

throughout the range are coastal sagebrush (Artemisia 

californica), Eastern Mojave buckwheat (Eriogonum 

fasciculatum), sage (Salvia spp.), brittlebush (Encelia

spp.), and other shrub species, with highest shrub 

diversity occurring in the southern range. Th e forbs, 

primarily annual, are especially diverse, with many 

species of concern under the Endangered Species Act 

occurring throughout the range of coastal sage scrub in 

California.

13.2.2 Ecosystem Responses to N Deposition

Ecosystem responses to elevated nitrogen (N) in 

coastal sage scrub include increases in exotic invasive 

annual grasses, loss of native shrub and forb cover, 

reduced diversity of native annual forbs and arbuscular 

mycorrhizal fungi, and elevated N mineralization. 

Plants in this low-productivity ecosystem respond 

quickly to N; thus changes in species abundance 

are excellent indicators of ecosystem response to N 

pollution. Th is is unlike ecosystems with long-lived 

vegetation, such as forests or chaparral, where initial 

ecosystem responses to N are measured by changes 

in biogeocycling or nutrient runoff (Aber et al. 1989, 

Section 13.3.3).

13.1 Ecoregion Description

Th e Mediterranean California ecoregion (CEC 1997; 

Fig 2.2) encompasses the greater Central Valley, 

Sierra foothills, and central coast ranges of California 

south to Mexico and is bounded by the Pacifi c Ocean, 

Sierra Nevada Mountains and Mojave Desert. Th e 

ecoregion description is adapted from CEC (1997).

It is distinguished by its warm, mild Mediterranean 

climate, chaparral vegetation, agriculturally productive 

valleys, and a large population (>30 million). Th e Coast 

Ranges crest at 600 to 1200 m. Th e broad, fl at, Central 

Valley is drained by the Sacramento and San Joaquin 

Rivers into the Sacramento Delta and San Francisco 

Bay. Rugged Transverse Ranges, peaking at 3506 m, 

border the Los Angeles basin. Soils are complex, mostly 

dry, and weakly developed with high calcium (Ca) 

concentrations. Summers are hot and dry (>18 o

C), 

winters are mild (>0 o

C) with precipitation from winter 

Pacifi c Ocean storms; coastal fog is common May 

through July. Annual precipitation (200 to 1000 mm) 

is highly variable and droughts are common. Vegetation 

is characterized by chaparral, patches of oak woodland, 

grassland, and some coniferous forest on upper 

mountain slopes. Th e chaparral has thickened, hardened 

foliage resistant to water loss and forms a cover of closely 

spaced, mostly evergreen shrubs 1 to 4 m tall. Common 

shrubs include chamise (Adenostoma fasciculatum), 

buckbrush (Ceanothus cuneatus), and manzanita 

(Arctostaphylos spp.). Coastal sagebrush (Artemisia 

californica) and summer-deciduous plants that tolerate 

more xeric conditions are found at lower elevations. A 

blue oak-California foothills pine (Quercus douglasiiPinus sabiniana) woodland community forms a ring 

around the Central Valley, which once had extensive 

grasslands and riparian forests. Th e southern oak 

woodland extends into the Transverse and Peninsular 

Ranges and includes Southern California walnut (Juglans 

californica) and Engelmann oak (Quercus englemannii). 

Mixed-conifer forests (Minnich 1999) occur at a higher 

elevation than chaparral and include a variety of species 
144 Chapter 13—Mediterranean California GTR-NRS-80

13.2.3 Range of Ecosystem Responses Observed

Studies along anthropogenic N deposition gradients 

and in experimental N-fertilized plots have shown 

alterations in soil N cycling, plant cover and diversity, 

and arbuscular mycorrhizal fungi. Nitrogen cycling 

studies include fi eld and laboratory N mineralization 

measurements that show higher turnover of inorganic 

N under high N deposition or fertilization (Sirulnik et 

al. 2007, Vourlitis et al. 2007a, Vourlitis et al. 2007b). 

Plant cover and diversity studies were done along an 

N deposition gradient and show threshold losses of 

native annual forb species of 45 percent (a decrease from 

67 to 37 species per site) under 10 kg N ha-1 yr-1 of 

modeled and 7.8 kg ha-1 yr-1 measured N deposition.15

A study of arbuscular mycorrhizal fungi was made on 

the same gradient and showed similar threshold losses of 

mycorrhizal root infection and diversity at many of the 

same sites showing plant loss under 9 kg N ha-1 yr-1 of 

modeled N deposition (Egerton-Warburton and Allen 

2000), as described below.

Nitrogen mineralization and nitrifi cation. Several 

fertilization studies conducted in coastal sage scrub 

used high levels of N fertilizer, 50 to 60 kg N ha-1 yr-1

(Allen et al. 1998, Vourlitis et al. 2007a). However, two 

studies on N mineralization (production of ammonium 

(NH4

+

) and nitrate (NO3

-

)) along an N deposition 

gradient in coastal sage scrub showed increased rates 

of N mineralization in soils receiving 10 compared 

to 4 kg N ha-1 yr-1 (Vourlitis et al. 2007b, Vourlitis 

and Zorba 2007). Another fi eld experiment (Sirulnik 

et al. 2007) tested the eff ects of elevated N on net N 

mineralization and nitrifi cation in coastal sage scrub 

in sites with modeled (14.7 and 8.7, as modeled by 

Tonnesen et al. [2007]) and inferential (20.2 vs. 6.6 kg 

N ha-1 yr-1; Table 13.3) N deposition rates. Few changes 

in N mineralization were observed, but nitrifi cation 

(production of NO3

-

) was signifi cantly higher in high 

deposition than control plots on at least one date in all 

3 years (Sirulnik et al. 2007). Because the studies by 

Vourlitis et al. (2007b) and Vourlitis and Zorba (2007) 

detected changes in soil N dynamics at a lower level of 

N deposition (10 versus 4 kg N ha-1 yr-1) than the study 

by Sirulnik et al. (2007) (20.2 versus 6.6 kg N ha-1 yr-1), 

the former are used to report the N deposition response 

threshold for mineralization.

Vegetation cover and species richness. Coastal sage scrub is 

subject to levels of total N deposition up to 20 kg ha-1

yr-1, as estimated by the U.S. Environmental Protection 

Agency’s Community Multiscale Air Quality (CMAQ) 

model (Tonnesen et al. 2007) in inland Riverside and 

San Bernardino Counties, an area that has been rapidly 

converted to exotic annual grassland in the past 30 to 40 

years (Allen et al. 1998, Talluto and Suding 2008). Th e 

conversion to grassland is likely caused by a combination 

of elevated N deposition that promotes increased grass 

biomass and frequent fi re that prevents establishment of 

native shrubs (Allen et al. 1998, Minnich and Dezzani 

1998, Talluto and Suding 2008).

Vegetation-type conversion studies in coastal sage scrub 

have examined patterns of vegetation change, but have 

not determined the N deposition response threshold. 

Research in the adjacent North American Deserts 

ecoregion related N deposition, exotic grass biomass, 

and fuel for fi re to determine the N deposition response 

threshold (Fenn et al. 2010, Rao et al. 2010, Chapter 

12).

A fi eld survey in 200315 determined the eff ects of N 

on the native and exotic vegetation along a gradient of 

seven sites that ranged in N deposition from 8.7 to 19 

kg N ha-1 yr-1 (Fenn et al. 2010; Table 13.1). Nitrogen 

deposition values reported are modeled wet plus dry 

deposition on a 4-km grid (Tonnesen et al. 2007). In 

addition, inferential calculations of N deposition are 

reported for three sites (Table 13.1). Th e inferential 

method is used to calculate dry deposition to surfaces 

(plant canopies and soil) based on atmospheric 

concentrations of N pollutants, the deposition velocities 

of each pollutant, surface areas, and other factors (Fenn 

et al. 2009). Measurements of air quality, soil N, and 

short-term deposition to leaf surfaces (Padgett et al. 

1999) confi rm this air pollution gradient. Extractable 

soil N (NO3

-

 plus NH4

+

) ranged from 10 μg g-1 at 

the low end of the N gradient to 38 μg g-1 at the high 

15Allen, E.B. Unpublished data. Professor and natural 

resources extension specialist, Department of Botany 

and Plant Sciences and Center for Conservation Biology, 

University of California, Riverside, CA 92521.
Chapter 13—Mediterranean California GTR-NRS-80 145

end. Percentage cover of exotic grasses was positively 

correlated with soil N and atmospheric N deposition, 

while percentage cover of native shrubs and forbs was 

inversely correlated with N deposition (P < 0.001). 

Native forb richness ranged from 67 to 16 species across 

the deposition gradient; the rapid drop in native forb 

richness from 67 to 37 species per site between 9 and 

11 kg N ha-1 yr-1 indicates impacts on plant community 

richness occur at 10 kg N ha-1 yr-1. Chosen sites had 

not burned in the previous 10 years, and were thus in a 

similar successional stage. However, the high deposition 

sites (modeled as 13 to 19 kg N ha-1 yr-1) have a history 

of two or more fi res since the 1960s, while the low 

deposition sites (8 to 11 kg N ha-1 yr-1) have burned 

only once since the 1960s. Th e two sites with deposition 

of 11 kg N ha-1 yr-1 have reduced native cover and 

richness even without frequent fi res, which suggests a 

direct eff ect of N on native species, such as increased 

competition by nitrophilous exotic grasses (Yoshida and 

Allen 2004).

Arbuscular mycorrhizal fungi. Both mycorrhizal root 

infection and mycorrhizal spore species density 

decreased greatly between 8.7 and 9.6 kg N ha-1 yr-1 

of modeled N deposition (Tonnesen et al. 2007) along 

the same N deposition gradient reported above in 

coastal sage scrub vegetation in southern California 

(Egerton-Warburton and Allen 2000, Fenn et al. 

2010). Th e decrease in mycorrhizal fungi occurred 

between inferential values of 6.6 and 8.9 kg N ha-1 yr-1; 

inferential values were not available for all sites (Table 

13.2). Mycorrhizal root infection indicates the extent 

to which plants may depend on mycorrhizae for their 

nutrient uptake, and is typically reduced in extent 

by elevated soil nutrients, while spore density is an 

indication of the amount of carbon fl owing from the 

plant to the fungus and indicates the strength of the 

mutualistic relationship.

13.2.4 Estimated Critical Loads

A summary of critical loads for ecosystem responses 

in coastal sage scrub is given in Table 13.3. Given 

the rapid decrease in native species cover and forb 

richness between 9 and 11 kg N ha-1 yr-1, 10 kg N ha-1

yr-1 may be estimated as the critical load for loss of 

native diversity and cover in coastal sage scrub based 

on modeled N deposition, while 7.8 kg N ha-1 yr-1 

may be calculated using inferential data (Table 13.1). 

Percent colonization and spore counts of arbuscular 

mycorrhizal fungi in coastal sage scrub decline steeply 

at deposition levels of 9.2 kg N ha-1 yr-1 using the 

modeled data and 7.8 kg N ha-1 yr-1 using inferential 

data at many of the same sites. A critical load for loss 

of arbuscular mycorrhizal activity is estimated at 9.2 kg 

Table 13.1—Percent cover and richness (per 3 ha) of plant groups along an N deposition gradient in 

western Riverside County, California. Sites are arranged from north to south along an urban to rural 

gradient. Forb species richness in bold shows a rapid drop in richness, suggesting a critical load of 

10 kg N ha-1 yr-1 by the CMAQ model (between 9.0 and 11.1 kg N ha-1 yr-1), and 7.8 kg N ha-1 yr-1 by the 

inferential method (between 6.6 and 8.9 kg N ha-1 yr-1). 

Site Exotic grass 

% cover

Native forb 

% cover

Shrub 

% cover

Native forbs

no. of species 

per 3 ha

soil N 

μg g-1

N deposition

kg N ha-1 yr-1

 CMAQa Inferentialb

Jurupa Hills 63.5 4.0 2.2 16 37.7 19.6

Box Springs 69.2 18.5 2.4 31 32.6 14.7 20.2

Botanic Garden 36.0 25.4 0.2 20 28.9 13.4

Lake Perris 0.5 26.1 2.8 30 20.3 11.1

Mott Reserve 6.7 14.3 11.2 37 30.6 11.1 8.9

Lopez Canyon 11.1 19.6 19.3 67 9.6 9.0 6.6

Tucalota Hills 1.5 35.7 35.0 50 10.5 8.7

a

N deposition using the CMAQ model is reported as modeled wet plus dry deposition (Tonnesen et al. 2007). 

b

Deposition calculated with the inferential method also includes wet plus dry deposition to vegetation and soil; see 

footnote 20 on page 161.
146 Chapter 13—Mediterranean California GTR-NRS-80

N ha-1 yr-1 using modeled data, and 7.8 kg N ha-1 yr-1 

using the inferential method (inferential data are not 

available for all modeled areas). Th e addition of more 

sample sites to the analysis along the gradient allows 

more precise values of critical loads to be estimated, as 

in the case of modeled N for mycorrhizal fungi. One 

challenge in setting critical loads for this system is that 

the lowest modeled N deposition is 8.7 kg N ha-1 yr-1, 

and the lowest value for the inferential method is 6.6 kg 

N ha-1 yr-1, values that are already signifi cantly elevated 

above pre-industrial background deposition. Th erefore, 

the critical load as defi ned from our study should be 

considered an upper limit. Th e threshold might actually 

be lower than these values. Th e critical load, with the 

limitations described above, is considered fairly reliable.

13.3 Chaparral and Oak Woodlands 

and the Central Valley

13.3.1 Ecosystem Description

Chaparral and oak woodland ecosystems are widespread 

over the Mediterranean California ecoregion. In 

this section, the term chaparral refers specifi cally to 

California chaparral, an ecosystem type composed of 

evergreen shrublands occupying most of the hills and 

lower mountain slopes of California. Chaparral has no 

commercial value, yet it forms an important cover that 

helps maintain the integrity of the watershed. Chaparral 

encompasses about 2.5 million ha, or 6.1 percent of 

the state (Keeley and Davis 2007). Shrubs common to 

chaparral include manzanita, Ceanothus (California 

lilac; Ceanothus spp.), buckbrush, chamise, and toyon 

(Heteromeles arbutifolia). Hardwood woodlands include 

several species of deciduous and live oaks (Quercus 

spp.), box elder (Acer negundo) and big leaf maple (Acer 

macrophyllum), California buckeye (Aesculus californica), 

California ash (Fraxinus latifolia), Oregon ash (Fraxinus 

latifolia), southern California walnut, and Fremont 

cottonwood (Populus fremontii). Occasional conifers are 

California foothill pine and Douglas-fi r (Pseudotsuga 

menziesii). Th e Central Valley, now largely converted 

to agricultural production, was chiefl y a grassland and 

forbland ecosystem, with riparian belts of hardwood 

trees (CEC 1997) and expansive tule (Schoenoplectus

acutus) sedge wetlands along rivers and streams (Munz 

1973).

13.3.2 Ecosystem Responses to N Deposition

Chaparral ecosystem responses to N deposition include 

elevated NO3

-

 leaching in streamwater and nitric oxide 

(NO) emissions from soil, increased direct transport 

Table 13.2—Spore counts and percent root infection of arbuscular mycorrhizal fungi 

along an N deposition gradient in western Riverside County, California. Sites are 

arranged from north to south along an urban to rural gradient (data from EgertonWarburton and Allen 2000). Values in bold show a change in spore density and root 

infection and are interpreted to represent a critical load.

Site Spores/g soil % root 

infection

soil N 

μg g-1

N deposition

kg N ha-1 yr-1

 CMAQa Inferentialb

Jurupa Hills 23 19 86.2 19.6

Waterman Rd. 54 18 15.6 16.9

Box Springs Mt. 66 17 26.4 14.7 20.2

Mockingbird Cyn 74 24 56.8 12.5

Lake Mathews 70 23 39.5 11.1

Motte Reserve 80 27 26.1 11.1 8.9

Hemet 78 22 12.8 9.6

Lake Skinner 103 39 7.5 8.7 6.6

Santa Margarita 105 45 9.7 8.7

a

N deposition using the CMAQ model is reported as modeled wet plus dry deposition (Tonnesen et al. 2007).

b

Deposition calculated with the inferential method also includes wet plus dry deposition to vegetation and soil; 

see footnote 20 on page 161.
Chapter 13—Mediterranean California GTR-NRS-80 147

of atmospheric NO3

-

 from polluted catchments, 

N enrichment of soil and vegetation, increased N 

cycling rates in soil, and decreased diversity, species 

richness, and productivity of arbuscular mycorrhizal 

communities. In chaparral and oak woodlands, lichen 

communities have been severely altered by N deposition. 

Elevated NO3

-

 leaching often follows high atmospheric 

N inputs to the catchment due to limited biological 

assimilation or retention. Th e N response of chaparral 

ecosystems is not known to be confounded by cooccurring ozone (O3

) eff ects as it is in forested areas (see 

section 13.4). Chaparral vegetation is highly O3

 tolerant 

(Stolte 1982, Temple 1999), and even if some species 

were O3

 sensitive, California chaparral vegetation is 

quiescent in summer, when O3

 levels are high.

Factors predisposing both California chaparral and 

forests to N loss are the active nitrifying characteristics 

of the soils (Fenn et al. 1998, Fenn et al. 1993, Riggan 

et al. 1985, Vourlitis and Zorba 2007). Chaparral soils 

exposed to chronic N deposition exhibit elevated N 

cycling, as evidenced by lower carbon:nitrogen (C:N) 

ratios, higher N mineralization rates, and increased 

15N (Vourlitis and Zorba 2007). Th e high nitrifi cation 

capacity of these soils is evidenced by virtually 

complete net nitrifi cation of mineralized N. Although 

N responses at low to moderate N input levels have 

not been evaluated in grasslands in the Central Valley, 

N deposition in serpentine grasslands near San Jose, 

California, has been shown to cause the displacement 

of native herbaceous species by exotic invasive annual 

grasses (Weiss 1999).

13.3.3 Range of Ecosystem Responses Observed

Data on N deposition eff ects on chaparral are primarily 

available from the south coast (Los Angeles) air basin 

and a watershed study in Sequoia National Park in 

central California. Although short-term N addition 

studies have been done, little is known about the longterm eff ects of N enrichment on vegetation.

Nitrate leaching and nitrifi cation. Because of the virtually 

complete nitrifi cation of mineralized N in chaparral and 

forested soils in southern California, approximately 95 

percent of the inorganic N in soil occurs in the form of 

NO3

-

 (Fenn et al. 1998, Fenn et al. 1993, Vourlitis et 

al. 2007b). Nitrifi cation in chaparral and coastal sage 

scrub soils increases in response to nitrogen deposition 

(Vourlitis et al. 2007b). Data relating atmospheric 

deposition and streamwater NO3

-

 in chaparral 

watersheds are available for three areas of California: 

in several catchments in the San Dimas Experimental 

Forest, located in the San Gabriel Mountains 45 km 

northeast of Los Angeles; eight streams in the Devil 

Canyon watershed, in the western San Bernardino 

Mountains; and Chamise Creek, a catchment located 

along the southwestern slope of the Sierra Nevada 

Mountains within Sequoia and Kings Canyon National 

Parks (Fenn et al. 2003b). In addition, 56 chaparral 

streams were sampled for NO3

-

 concentrations in a 

Table 13.3—Empirical critical loads of nutrient N for coastal sage scrub ecosystems of the Mediterranean California 

ecoregion. Reliability rating: ## reliable; # fairly reliable; (#) expert judgment

Site Critical load for 

N depositiona

kg N ha-1 yr-1

Reliability Response Comments Study

Coastal sage scrub, 

W. Riverside Co.

7.8-10 # Increase in exotic grass 

cover, decrease in native 

richness and cover 

See Table 13.1 Allen unpublishedb

; 

Fenn et al. 2010

Coastal sage scrub, 

W. Riverside Co.

7.8-9.2 # Decrease in arbuscularmycorrhizal spore 

density, richness, and 

percent root infection

Table 13.2 Egerton-Warburton 

and Allen 2000; 

Fenn et al. 2010

Coastal sage scrub, 

W. Riverside Co.

10 # Increase in N 

mineralization rate

Comparison of sites 

with 4 and 10 kg N 

ha-1 yr-1deposition

Vourlitis and Zorba 

2007, Vourlitis et al. 

2007b

a

N deposition is modeled as total wet plus dry N (Tonnesen et al. 2007)

b

See footnote 15 on page 144
148 Chapter 13—Mediterranean California GTR-NRS-80

regional survey in March 1982 across the greater Los 

Angeles Basin and in northern San Diego County over 

a wide range of N deposition exposures, demonstrating 

the relationship between N deposition and NO3

-

concentrations in streamwater in these ecosystems 

(Riggan et al. 1985).

Over 20 years of atmospheric deposition and 

streamwater NO3

-

 data are available from San Dimas 

Experimental Forest watersheds (Meixner et al. 2006; 

Riggan et al. 1985, 1994). Atmospheric and streamwater 

NO3

-

 data over shorter periods are available from Devil 

Canyon (Fenn and Poth 1999, Meixner and Fenn 

2004). Th e Chamise Creek watershed was sampled 

continuously from 1983 to 1998 (Fenn et al. 2003a). 

In all three of these study areas, the catchments can be 

considered to be at Stage 2 of Stoddard’s watershed N 

saturation model (Stoddard 1994) based on the seasonal 

NO3

-

 export patterns and peak NO3

-

 concentrations. 

More information on N saturation can be found in 

Chapter 1.

Nitrate concentrations in streamwater from chaparral 

watersheds in California are the highest reported from 

North America; peak concentrations greater than 200 

to 300 μeq L-1 are common in catchments receiving 

elevated N deposition while peaks as high as 700 μeq 

L-1 have been reported (Fenn et al. 2003a, 2003b; Fenn 

and Poth 1999; Riggan et al. 1985, 1994). On the 

other hand, total NO3

-

 export is often less than in N 

saturated mesic forests, presumably because of the high 

evapotranspiration fl uxes relative to annual precipitation 

inputs. Nitrogen export in N-saturated chaparral 

catchments typically ranges from 1 to 10 kg ha-1 yr-1, 

varying greatly from year to year because of highly 

variable precipitation (Fenn and Poth 1999, Riggan et 

al. 1985).

Th e proportion of NO3

-

 exported from chaparral 

catchments in the Transverse Ranges of southern 

California that was direct throughput of atmospheric 

NO3

-

 (e.g., without biological assimilation or retention 

within the watersheds) was quantifi ed using a state-ofthe-art isotopic method developed at the University 

of California in San Diego (Δ17O as a tracer for 

atmospheric NO3

-

) (Michalski et al. 2004). During peak 

fl ow following storm events, about 40 percent of the 

stream NO3

-

 was found to originate from atmospheric 

sources that are transported through the watershed 

without any prior biological assimilation. All soil and 

aquatic NO3

-

 samples in the study had positive Δ17O 

values, unambiguously showing that every sample of soil 

and water (including groundwater) collected exhibited 

a signal of direct atmospheric NO3

-

 inputs (Michalski et 

al. 2004), thus highlighting the N-saturated condition 

of these catchments. Nonetheless, on average, only 

about 10 percent of total N deposition is exported in 

streamwater from N-saturated catchments in the San 

Dimas Experimental Forest and in Devil Canyon (Fenn 

and Poth 1999, Meixner and Fenn 2004, Meixner et al. 

2006, Riggan et al. 1985).

Trace gas losses of N. Nitrogen-saturated chaparral 

ecosystems also exhibit trace N gas emissions, primarily 

as NO, but nitrous oxide (N2

O) emissions can also 

be quantitatively important during the relatively short 

periods when soils are suffi ciently wetted (Anderson 

and Poth 1989). In more anaerobic zones of the 

groundwater system and near riparian zones, losses of 

reduced N gases are also expected to be important in N 

saturated chaparral catchments. Fluxes of NO from soil 

over a 6-month period from July to December 1986 

were estimated to be 1 kg N ha-1 in an unburned site 

and 3 kg N ha-1 in a burned chaparral site in the San 

Dimas Experimental Forest (Anderson and Poth 1989). 

Data are not available on trace gas losses from the 

near-riparian zone, or from phreatic zones in chaparral 

or forested watersheds in California with elevated N 

deposition. However, very high levels of dissolved N2

O 

were found in groundwater samples collected from a 

fault line (Fenn and Poth 1999) within the chaparral 

zone of the N-saturated Devil Canyon watershed.16 Th is 

pattern suggests that large amounts of excess N may be 

lost from these catchments via degassing of dissolved N 

trace gases (Bowden and Bormann 1986).

16Fenn, M.E. Unpublished data. Research plant pathologist, 

Forest Fire Laboratory, 4955 Canyon Crest Dr., Riverside, 

CA 92507.
Chapter 13—Mediterranean California GTR-NRS-80 149

Fire eff ects on N losses. Ecosystem responses to chronic 

N deposition must be considered within the context 

of regular fi re occurrence in chaparral. Stand-replacing 

fi re return intervals in these ecosystems are generally on 

the order of 40 to 60 years (Minnich and Bahre 1995). 

Over time, N from atmospheric deposition accumulates 

in soil, organic matter, necromass, and biomass. During 

a burn, much of the N in the aboveground N-enriched 

vegetation and detritus is released to the atmosphere 

or deposited on the soil surface, where it is mobilized 

via nitrifi cation, N trace gas emissions, NO3

-

 leaching, 

and erosion, thus resulting in large pulses of N losses 

(Anderson and Poth 1989, Riggan et al. 1994). Such 

losses are of much shorter duration and are more muted 

in burned chaparral catchments not exposed to elevated 

N deposition (Riggan et al. 1985, 1994). However, fi re 

removes only a minor component of the large N pools 

in the mineral soil. As a result, a large pool of N remains 

in the ecosystem even after a stand-replacing fi re. Th is, 

combined with continuing atmospheric inputs, results 

in continued long-term elevated NO3

-

 leaching in 

surface runoff and groundwater (Meixner et al. 2006). 

Th us, fi re alone does not appear to be an eff ective tool 

for mitigating N saturation, although simulation studies 

suggest that N losses from higher elevation forested 

regions can be mitigated if N deposition is decreased 

and prescribed fi re is applied periodically (Gimeno et al. 

2009).

Soil acidifi cation. Th e pH of chaparral soils measured at 

more than 700 San Gabriel Mountain sites in the 1970s 

was compared to measurements taken at more than 

300 sites in the 1990s. Surface soil pH values decreased 

considerably over this period (Wood et al. 1992). Values 

of pH in the 1970s typically ranged from 6.0 to 7.4 

compared to a typical range of 4.6 to 5.6 in the 1990s.17

Arbuscular mycorrhizal fungi. In a retrospective study 

from the San Dimas Experimental Forest, historical 

trends in biodiversity of the arbuscular mycorrhizal 

community were evaluated in pure stands of Eastern 

Mojave buckwheat and chamise established in the 

1940s on a homogenized fi ne sandy loam soil (EgertonWarburton et al. 2001). Th is analysis was done by 

evaluating the diversity of arbuscular mycorrhizal fungal 

spores in archived soils collected periodically from 1937 

to 1999, and from examinations of mycorrhizae on root 

samples in 1999. It was observed that N enrichment 

of the soils increased over the same period in which 

mycorrhizal communities experienced dramatic changes 

(Egerton-Warburton et al. 2001). Diversity, species 

richness, and productivity of the arbuscular mycorrhizal 

community had deteriorated severely by 1969. Th ree 

previously common mycorrhizal genera disappeared 

from the mycorrhizal spore community in soil, and one 

large-spored genus (Gigasopora) was no longer found in 

the rhizosphere soil. Nitrogen enrichment also enhanced 

the proliferation of potentially less mutualistic species 

of small-spored Glomus, which may have implications 

for plant community succession in the face of chronic 

N deposition (Egerton-Warburton et al. 2001). 

Th roughfall N fl uxes in the San Dimas Experimental 

Forest in the early 1980s were 23 kg N ha-1 yr-1 (Riggan 

et al. 1985) and total deposition was estimated at 35 kg 

N ha-1 yr-1 (Meixner et al. 2006).

Lichen responses. About half the epiphytic lichen species 

known to occur during the late 1800s and early 1900s 

on coast live oak throughout the chaparral and oak 

vegetation zones of the Los Angeles basin (Hasse 1913) 

have subsequently disappeared (Ross 1982). Th is 

dramatic reduction in species richness was initially 

attributed solely to oxidizing pollutants, chiefl y O3

(Nash and Sigal 1999, Ross and Nash 1983), but 

recent research provides unequivocal evidence for an 

N deposition eff ect (Riddell et al. 2008). In the Los 

Angeles air basin, throughfall N deposition in forests 

downwind of the urban areas can reach 25 to 70 kg ha-1

yr-1 (Fenn et al. 2008). Th is deposition includes nitric 

acid (HNO3), a strong gas-phase acid which exhibits 

diurnal patterns paralleling O3 concentrations. In the 

mountains downwind of Los Angeles, 24-hr means of 

up to 27.3 ppb have been recorded, whereas remote 

location means have been ≤0.1 ppb (Bytnerowicz and 

Fenn 1996). Unlike O3

, once HNO3 is produced, it 

rapidly deposits to surfaces. Menzies’ cartilage lichen 

(Ramalina menziesii Tayl.), a dominant epiphyte of

oaks throughout the Coast Ranges of California and 

17Wood, H.B. Unpublished data on fi le with Mark Fenn, 

Research plant pathologist, Forest Fire Laboratory, 4955 

Canyon Crest Dr., Riverside, CA 92507.
150 Chapter 13—Mediterranean California GTR-NRS-80

historically common in the Los Angeles basin (Hasse 

1913) was treated with nitric acid at 7 to 25, and 19.9 

to 25 μg m-3 during month-long fumigations to mimic 

diurnal patterns in Los Angeles (Riddell et al. 2008). 

All specimens in both treatments experienced marked 

declines in chlorophyll content, carbon exchange 

capacity, and membrane integrity compared to controls. 

In addition to oxidant and acidic air pollutants, Riddell 

et al. (2008) also recognized habitat loss, lower humidity 

due to urbanization, and increased fi re incidence as 

factors in the disappearance of Menzies’ cartilage and 

other lichens from the Los Angeles basin.

Epiphytic lichen communities in more northerly parts 

of the ecoregion are also showing a clear response to 

N. Th e Greater Central Valley, comprising the Central 

Valley proper and the surrounding chaparral- and oak 

woodland-dominated central Coast Ranges and Sierra 

foothills, encompasses many urban centers and is 

among the most agriculturally intensive regions of the 

United States. Using a nonmetric multidimensional 

scaling procedure and vector overlays of environmental 

variables, Jovan and McCune (2005) ordinated and 

scored lichen communities surveyed between 1998 and 

2001 at 118 sites in the Greater Central Valley along an 

NH3 emissions gradient (Fig. 13.1). Th ere were strong 

correlations between NH3 and percentages of eutrophic 

lichen richness and abundance. In addition, total N 

deposition and other N species, including HNO3

, 

NO3

-

, and NO2, were all positively correlated with 

lichen scores, suggesting that multiple N-containing 

compounds, not just NH3, are impacting lichens in the 

Greater Central Valley (Jovan 2008). Eutrophic lichens 

benefi ted from increasing N availability, comprising 

>50 percent of total abundance in lichen communities 

at air scores <0.0 (Fig. 13.2). Higher air scores indicate 

better air quality. Eutroph abundance at sites with the 

best air quality ranged from 20 to 50 percent. Th erefore 

an air score of 0.0 may be a reasonable lichen response 

threshold. CMAQ data were not available to the authors 

at the time of publication, but overlay of a 4-km grid 

of modeled 2002 CMAQ data (Tonnesen et al. 2007) 

over lichen survey site coordinates provides a more 

complete understanding of N deposition at the Jovan 

and McCune (2005) sites. Of the 53 lichen survey sites 

in the dataset with clean air scores (>0.0), only three 

occurred in grid cells where CMAQ predicted >5.5 kg 

ha-1 yr-1 total N deposition.

13.3.4 Estimated Critical Loads

In chaparral catchments exhibiting the symptoms 

of N excess discussed above, N input values range 

from 5 to 33 kg ha-1 yr-1 as throughfall (Meixner and 

Fenn 2004), except for the case of the chaparral/oak 

woodland lichen response studies in the greater Central 

Valley, where N deposition was simulated with the 

CMAQ model (Tonnesen et al. 2007). In the seven 

catchments sampled in Devil Canyon, throughfall N 

deposition ranged from 14 to 33 kg ha-1 yr-1. Because 

all seven sites had elevated NO3

-

 concentrations in 

stream water, the lowest deposition value can be 

considered to be at or above the critical load. Chamise 

Creek in Sequoia National Park is an N-saturated site 

based on elevated NO3

-

 leaching (Fenn et al. 2003a, 

2003b); throughfall N deposition18 was 10 kg ha-1

yr-1. Lichen survey sites with clean air scores suggest a 

preliminary critical load of 5.5 kg N ha-1 yr-1. In a test 

of the broader applicability of a model developed from 

western Oregon and Washington lichen community 

data to temperate forests in general, Geiser et al. (2010), 

essentially replicated this estimate. For the Greater 

Central Valley dataset (Jovan and McCune 2005) and 

response threshold (50 percent eutrophs), Geiser et 

al. (2010) calculated a critical load of 3 to 6 kg N ha-1

yr-1 for the study area, increasing with mean annual 

precipitation from 17 to156 cm. An N critical load for 

serpentine grassland invasion by exotic annual grasses 

was determined along a roadside deposition gradient at 

a site downwind of San Jose, California, and west of the 

Central Valley. Th e critical load of 6 kg N ha-1 yr-1 was 

based on the N deposition level at the point where the 

grass invasion was visibly diminished (Fenn et al. 2010, 

Weiss 1999). Considering the deposition and response 

data from all the ecosystem components considered 

here, we suggest a deposition range for the critical load 

of 3.1 to 14 kg N ha-1 yr-1. All the reported responses 

(see Table 13.4) occur within this range.

18Homyak, P.M. 2009. Personal communication. Graduate 

research associate, Department of Environmental Sciences, 

University of California, Riverside, CA 92521.
Chapter 13—Mediterranean California GTR-NRS-80 151

In the San Dimas Experimental Forest, with a reported 

throughfall N deposition of 23 kg ha-1 yr-1, the 

critical load is exceeded for NO3

-

 leaching, changes in 

mycorrhizal spore community changes, and elevated NO 

emissions from soil. Th e critical load for NO3

-

 leaching 

is known to be much lower than the deposition at 

San Dimas as discussed above, but the critical load for 

mycorrhizal community changes and for N trace gas 

emissions cannot be determined because we only have 

data from San Dimas. Likewise, the critical load for 

increased nitrifi cation has not been established due to 

insuffi cient corresponding deposition data at sites where 

nitrifi cation has been measured. Th us, the critical load for 

these responses are not included in the critical loads tables.

13.4 Mixed-conifer Forest

13.4.1 Ecosystem Description

Air pollution eff ects studies in California forests 

have focused largely on the mixed-conifer forests 

found at mid-elevation sites in the Transverse Ranges 

(particularly the San Bernardino Mountains) in the Los 

Angeles air basin and in the southwestern Sierra Nevada 

(Fenn et al. 2003b). Because of the high sensitivity to 

Figure 13.1—Air quality scores 

for the greater Central Valley 

divided into air quality zones. 

Reprinted from Jovan 2008.
152 Chapter 13—Mediterranean California GTR-NRS-80 Table 13.4—Empirical critical loads of nutrient N for chaparral and Central Valley ecosystems of the Mediterranean California ecoregion. Reliability rating: ## reliable; # fairly reliable; (#) expert judgment Site Critical load for N depositiona kg N ha-1 yr-1 Reliability Response Comments Study Sacramento & San Joaquin Valleys, Coast Ranges & Sierra Foothills 3-6 # Change in lichen communities Shift to eutroph dominance, precipitation range of 17-156 cm. Geiser et al. 2010 San Francisco Peninsula, near San Jose, California 6 ## Annual grass invasion, replacing native herbs in a serpentine grassland Critical load based on a local roadside gradient; Serpentine grassland site is actually west of the Central Valley. Weiss 1999, Fenn et al. 2010 Sacramento & San Joaquin Valleys, Coast Ranges & Sierra Foothills 5.5 # Changes in lichen communities Eutrophic lichens increased at expense of species adapted to lower N availability. Fenn et al. 2010, Jovan and McCune 2005, Jovan 2008, this chapter Chamise Creek, Sequoia NP 10 # Streamwater [NO3-] > 14 μM This NO3 threshold indicates the beginning of N saturation. Fenn et al. 2003b, Fenn et al. 2003a, Fenn et al. 2010 Devil Canyon, San Bernardino Mountains 14 # Streamwater [NO3-] > 14 μM Estimated N deposition in throughfall in the 7 catchments was 14-33, so critical load is lowest value. Meixner and Fenn, 2004, Fenn and Poth 1999, Fenn et al. 2010

Los Angeles Basin Atmospheric 

concentrations 

of 7-25 μg 

HNO3 m-3

# Loss of lichen species 

richness and harm to physiological processes

About half of the 1900 lichen fl ora in the Los Angeles basin is currently extirpated, attributed to O3 and N 

pollutants.

Riddell et al. 2008; Hasse 1913; Ross, 1982; Sigal and Nash 1983

aN deposition is reported as throughfall (Fenn and Poth 2004, Fenn et al. 2008) except for the serpentine grassland study and lichen studies in chaparral. In the Los Angeles 

Basin the range of atmospheric concentrations shown for HNO3 indicate the concentrations at which lichens are impacted. For changes in lichen communities in the Central 

California valleys, coast ranges and Sierra foothills, N deposition values for determining the critical loads were simulated using the CMAQ model (Tonnesen et al. 2007). Deposition data for Chamise Creek, Sequoia National Park provided by P.M. Homyak; see footnote 18 on page 150
Chapter 13—Mediterranean California GTR-NRS-80 153

O3

 of ponderosa pine (Pinus ponderosa) and the closely 

related Jeff rey pine (Pinus jeff reyi), most of the research 

has been carried out in sites with a predominant 

component of these species. Associated tree species that 

co-occur in varying mixes with ponderosa and Jeff rey 

pine include white fi r (Abies concolor), sugar pine, 

(Pinus lambertiana), incense-cedar, and California black 

oak (Quercus kelloggii). In parts of the southwest Sierra 

Nevada, Douglas-fi r and giant sequoia (Sequoiadendron 

giganteum) are also found (Minnich 1999, Peterson 

and Arbaugh 1992). Shrub understory in mixed-conifer 

forest is mostly open or absent (cover <20 percent), 

depending on altitude and climatic gradients. Bracken 

fern (Pteridium aquilinum var. pubescens) is common 

in more mesic sites. See Minnich (1999) for a more 

complete description of common understory species in 

mixed-conifer forests.

13.4.2 Ecosystem Responses to N Deposition

Research on air pollution eff ects in mixed-conifer forests 

began in the 1960s after the initial discovery in the 

1950s of unusual symptoms in ponderosa pine that 

were later confi rmed to be caused by O3

. Studies on N 

deposition eff ects in forests began in the San Bernardino 

Mountains (Fenn 1991) in some of the original O3

study plots established in the 1970s (Arbaugh et al. 

2003). Ecosystem responses to N deposition include 

NO3

-

 leaching, trace gas losses of N, soil acidifi cation, 

lichen community change, altered forest nutrient 

cycling, and loss of understory diversity. Because of the 

co-occurrence of O3

 and N deposition in the mixedconifer forest of California, the ecological eff ects of N 

deposition cannot be evaluated in isolation from the 

signifi cant impacts of O3

 on nutrient cycling and plant 

growth, development, community succession, and 

phenology. Th e combined ecological eff ects of N and O3

are dramatic in areas of high pollution exposure (Fenn et 

al. 2003b, Grulke et al. 2009).

13.4.3 Range of Ecosystem Responses Observed

Most characteristics which predispose an ecosystem 

to NO3

-

 leaching (e.g., climatic conditions, actively 

nitrifying soils) discussed for chaparral ecosystems 

in section 13.3 also apply to forested catchments, 

as many mixed-conifer catchments grade into 

chaparral vegetation at lower elevations. Indeed, 4 of 

11 catchment streams monitored to determine the 

relationship between N deposition and NO3

-

 leaching 

in the mixed-conifer forest were accessible for sampling 

only at the lower elevation chaparral portion of the 

catchment (Fenn et al. 2008). In both chaparral and 

mixed-conifer forests, a key factor predisposing these 

ecosystems to N loss with chronic N deposition is the 

temporal asynchrony between the period of greatest 

biotic demand (spring and early summer) and the winter 

when approximately 85 percent of the precipitation and 

the greatest surface runoff occurs (Fenn and Poth 1999).

Nitrate leaching. A threshold streamwater NO3

-

concentration for a site experiencing relatively low 

atmospheric N deposition (14 μM) was empirically 

determined to identify watersheds that are exporting 

elevated levels of NO3

-

 (see Fenn et al. 2008, 2010 for 

more details on these methods and rationale). Based on 

regression analysis of throughfall N deposition versus 

streamwater NO3

-

 concentrations, the N deposition 

level at which NO3

-

 leaching exceeds the “unpolluted” 

-site threshold is approximately 17 kg ha-1 yr-1 (Fenn et 

al. 2008). Th is level of N deposition has been measured 

in the regions of the Transverse Ranges in southern 

California most exposed to air pollution and along the 

western edge of the Sierra Nevada (Breiner et al. 2007; 

Fenn et al. 2008, 2010).

Figure 13.2—The relationship between air quality score 

and the proportion of nitrophyte (eutroph) abundance in 

California’s greater Central Valley. Reprinted from Jovan 

2008.

-1.5 -1.0 -0.5 0.0 0.5 1.0 1.5

Air quality scores

0.0

0.2

0.4

0.6

0.8

1.0

Proportion of nitrophyte abundance

High N/pH Low N/pH

r = 0.78
154 Chapter 13—Mediterranean California GTR-NRS-80

Although many factors aff ect N loss rates, the 

available data suggest that within a given catchment, 

N deposition fl uxes to the higher elevation forested 

portions of catchments are greater than to the lower 

chaparral and scrub vegetation portions (Meixner 

and Fenn 2004). Th is is presumably because of the 

much greater surface area of forested stands, which are 

effi cient collectors of dry and cloud water deposition, 

and because of greater occurrence of fog and cloud 

deposition at the upper levels of these catchments (Fenn 

et al. 2000, Meixner and Fenn 2004). However, the 

greater input to the forested portion of the catchment 

is likely counterbalanced, to some degree, by greater 

biotic N demand of the forest. Soil leachate and streams 

in the mixed-conifer zone are known to export high 

levels of unassimilated NO3

-

 (Michalski et al. 2004). In 

catchments covered by mixed-conifer forests at higher 

elevations, NO3

-

 concentrations in streamwater decrease 

within the lower elevation chaparral zone because of 

in-stream and near-stream N uptake and consumptive 

processes along the descending course of the stream 

(Meixner and Fenn 2004).

Trace gas losses of N. Fluxes of trace gas emissions 

from soil of N-saturated sites in the mixed-conifer 

forest appear to be very similar to those from 

N-saturated chaparral sites. Annual fl uxes of NO 

in high N deposition chaparral and forested sites 

generally range from 2 to 3 kg ha-1 yr-1 (Anderson 

and Poth 1989, Fenn and Poth 2001), although there 

is considerable uncertainty because of temporal and 

spatial heterogeneity in fl ux measurements. Emissions 

from soil are higher after a burn and after soil wetting. 

In short-term measurements, fl uxes of NO and N2

O 

increase markedly after soil wet-up, particularly in N 

saturated sites (Anderson and Poth 1989, Fenn et al. 

1996). Annual fl ux rates are probably greater in high 

precipitation years than the estimates given above.

Soil acidifi cation. Because of the Mediterranean climate 

and typical high base saturation of soils in semiarid forests, the possible soil acidifi cation eff ects of 

N deposition in California have not received nearly 

as much attention as the eff ects of N as a nutrient. 

However, soils have acidifi ed at an accelerated rate in 

high deposition sites. At Camp Paivika in the western 

San Bernardino mountains, soil pH values in the 

top 25 cm of soil ranged from 4.8 to 5.6 in the early 

1970s, compared to recent values ranging from 3.1 to 

4.319 (Fenn and Poth 1996, Wood et al. 2007). Th e 

unusually low pH values and rapid acidifi cation at 

Camp Paivika are due to high throughfall N deposition 

(70 kg ha-1 yr-1) and underlying stone lines at 40 to 60 

cm and 130 to 170 cm depths which control the spatial 

and temporal fl ow of soil percolates containing high 

concentrations of NO3

-

. At another site with the stone 

lines but much lower throughfall N deposition (6 kg 

ha-1 yr-1), soil acidity was relatively constant (pH 5.1 in 

1975; pH 5.0 in 2004) during the last 30 years (Wood 

et al. 2007).

Soil pH is highly correlated (r2

 = 0.99) with throughfall 

N deposition in the San Bernardino Mountains 

(Breiner et al. 2007). Based on this regression, soil 

pH is estimated to be 4.8 when deposition is 17 kg 

ha-1 yr-1. Soil base saturation is also much lower at 

high N deposition sites (30 to 80 percent) compared 

to low deposition sites (85 to 100 percent) in the San 

Bernardino Mountains19 (Fenn et al. 1996). Th e low pH 

and lower base saturation of the mineral soil does not 

inhibit active nitrifi cation, but a recent study suggests 

that nitrifi cation is dominated by heterotrophic nitrifi ers 

(Jordan et al. 2005).

Ectomycorrhizal fungi. Fungal communities of 

ectomycorrhizae were studied from November 2005 to 

June 2007 at a high (71 kg N ha-1 yr-1; Camp Paivika) 

and low (7 kg N ha-1 yr-1; Camp Osceola) deposition site 

in the San Bernardino Mountains east of Los Angeles 

(Sirajuddin 2009). Th e eff ects of added N (150 kg ha-1

yr-1) within each of the plots over a 9-year period were 

also investigated.

Basidiomycetes tended to decline with both N 

fertilization and at the high N deposition plot. Russula

(fi ve species) and Rhizopogon (three species) were present 

at Camp Osceola, but were lost with N fertilization 

and were not found at Camp Paivika. Cortinarius spp. 

19Wood, Y.A. Unpublished data on fi le with Mark Fenn, 

Research plant pathologist, Forest Fire Laboratory, 4955 

Canyon Crest Dr., Riverside, CA 92507.
Chapter 13—Mediterranean California GTR-NRS-80 155

also decreased with N fertilization. Th e exceptional 

basidiomycete was a Lactarius sp., which increased in N 

fertilized plots at both locations. In general, ascomycetes 

increased with N availability; Cenococcum geophilum in 

particular increased quite dramatically. Wilcoxinia rehmii

was the exception, in that this ascomycete declined with 

N enrichment, both via deposition and fertilization. 

In summary, N deposition and N fertilization shifted 

the ectomycorrhizal fungal communities in a similar 

direction, toward ascomycetes. Basidiomycetes generally 

declined and some key species disappeared with N 

fertilization or deposition (Sirajuddin 2009).

Lichen responses. Lichen-based responses to N deposition 

for the mixed-conifer zone of the Sierra Nevada 

(Northwest Forested Mountains ecoregion) likely 

have application to the mixed-conifer zone of the 

Mediterranean California ecoregion, especially in the 

southwest Sierras and Transverse Ranges, where there 

was a historic species overlap. Th e Sierra study (Fenn 

et al. 2008) utilized a threshold N concentration in the 

wolf lichen (Letharia vulpina (L.) Hue), an epiphyte 

on tree trunks and branches, from a low N deposition 

site. As the threshold is exceeded with increasing N 

deposition, the abundance of ecologically important 

oligotrophic species begins to decline, shifting lichen 

community composition from oligotroph-dominated 

to mesotroph- and eutroph-dominated (Fenn et al. 

2008). Oligotrophs are sensitive to even small increases 

in N; mesotrophs tolerate moderate inputs; eutrophs 

are fast growing, ‘weedy’ species of small ecological 

importance that thrive under conditions of high N 

availability. Lichen surveys in the Sierra Nevada mixedconifer zone were co-located with or near throughfall 

deposition monitoring sites. Similar studies can no 

longer be done in the San Bernardino Mountains 

because most air pollution sensitive species have long 

since disappeared (Nash and Sigal 1999, Sigal and Nash 

1983). In relatively low N deposition sites in the Sierras 

(<3.1 kg N ha-1 yr-1), oligotrophic lichen species still 

predominate. At a throughfall N deposition level of 5.2 

kg N ha-1 yr-1, the lichen community completely shifts 

from oligotroph to mesotroph/eutroph dominance. 

Oligotrophs are extirpated at 10.2 kg N ha-1 yr-1 (Fenn 

et al. 2008). Conservation of oligotrophic species is 

important because they make integral contributions 

to food webs, nesting material and insect habitat, and 

nutrient and hydrologic cycles (McCune et al. 2007).

Forest sustainability. Nitrogen deposition, in concert 

with O3 injury eff ects, is contributing to a decrease 

in the sustainability of mixed-conifer forests in the 

San Bernardino Mountains. Nitrogen deposition 

increases aboveground growth of coniferous and 

deciduous species of the mixed-conifer forest (Fenn 

and Poth 2001). Ozone and increased N fertility 

result in decreases in both C allocation belowground 

and fi ne root biomass (Grulke et al. 1998). Ozone 

causes premature foliar senescence and abscission. Th e 

combined eff ects of O3

 and elevated N deposition are 

increased C storage in the bole and woody aboveground 

biomass and accelerated foliar turnover and litter 

production (Arbaugh et al. 1999). Th is also results 

in increased accumulation of C and N in the forest 

fl oor. In these O3

-impacted and N-saturated mixedconifer stands (N deposition 25 to 71 kg ha-1 yr-1) 

(Fenn et al. 2008), aboveground N pools are much 

higher than in stands in areas receiving relatively low 

deposition (Arbaugh et al. 1999, Fenn et al. 2005, 

Grulke et al. 2009). For example, at an N-saturated site 

(N deposition 71 kg ha-1 yr-1) in the San Bernardino 

Mountains about 30 percent of ecosystem N is stored in 

the thick forest fl oor, compared to about 11 percent at 

an N-limited site (N deposition approximately 7 kg ha-1

yr-1) (Arbaugh et al. 1999, Fenn et al. 2005).

Ozone and elevated N deposition cause specifi c changes 

in forest tree C, N, and water balance that enhance 

individual tree susceptibility to drought, bark beetle 

attack, and disease, and when combined, contribute to 

whole ecosystem susceptibility to wildfi re (Grulke et al. 

2009). Th e N deposition levels at which these eff ects 

begin to occur are not well defi ned, but regression 

analysis indicates a 25 percent reduction in fi ne root 

biomass at 17 kg ha-1 yr-1 (Fenn et al. 2008); dramatic 

shifts in tree phenology and C allocation are evident at 

a site with N deposition of 39 kg ha-1 yr-1 (Fenn et al. 

2008, Grulke and Balduman 1999). Uncharacteristically 

deep litter layers develop in mixed-conifer forests 

impacted by air pollution. Elevated O3

 and N deposition 

decrease the proportion of whole tree biomass in foliage 

and roots, the latter eff ect increasing tree susceptibility 
156 Chapter 13—Mediterranean California GTR-NRS-80

to drought and beetle attack. Because both foliar and 

root mass are compromised, carbohydrates are stored 

in the bole over winter. Elevated O3

 increases drought 

stress by signifi cantly reducing plant control of water 

loss. Th e resulting increase in canopy transpiration, 

combined with O3

 and N deposition-induced decreases 

in root mass, signifi cantly increases tree susceptibility 

to drought stress. Th e combination of these eff ects 

with increased sequestration of bole carbohydrates 

may contribute to successful host colonization and 

population increases of bark beetles (Jones et al. 2004) 

and possible enhancement of dwarf mistletoe infections, 

although the latter response has not been confi rmed.

Understory. Understory diversity in mixed-conifer forests 

in the San Bernardino Mountains in southern California 

was recently compared to studies done 30 years prior, 

in 1973 (Allen et al. 2007). Both O3

 concentrations 

and particularly N deposition decline from west to east 

along the air pollution gradient. Nitrogen deposition 

in throughfall ranges from 9 to 71 kg ha-1 yr-1 at the 

understory study sites (Fenn et al. 2008). Biodiversity 

loss was pronounced in the sites receiving the highest 

N deposition and is due to the establishment of exotic 

invasive species that have become abundant. In three of 

six sites, including the two westernmost polluted sites, 

20 to 40 percent of species were lost between 1973 

and 2003. In the highest deposition sites, understory 

cover equaled 30 to 45 percent and was equally divided 

between native and exotic species. At lower deposition 

sites, understory cover was 3 to 13 percent and was 

dominated by native species. Because of confounding 

factors such as precipitation and possibly local 

disturbances, a simple correlation between air pollution 

and patterns of native and invasive species cover 

and richness was not found. However, observational 

evidence and expert opinion suggest that increased N 

deposition and precipitation in the westernmost sites are 

the primary factors contributing to reduced biodiversity 

and increased cover by invasive species (Allen et al. 

2007).

Co-occurring O3

 may be indirectly contributing to the 

establishment of exotic species as well. Ozone causes 

premature foliage loss in pine, while N deposition 

stimulates foliar growth, leading to greater litter 

production and accumulation in the forest fl oor (Fenn 

et al. 2003b, 2005). Many native plant species are 

not able to establish where dense litter accumulates. 

However, stickywilly (Galium aparine), a weedy annual 

with both native and introduced forms, thrives under 

these conditions, which include the acidifi ed N-rich 

soils that underlie the thick litter layer. Portions of the 

high-pollution study sites in the western San Bernardino 

Mountains burned in October 2003. Formerly shady 

sites in the burned areas are now covered by exotic 

annual brome grasses. Subsequent clearing of dead trees 

and brush is further exposing the soil to invasive species 

(Allen et al. 2007).

13.4.4 Critical Loads Estimates

Th e N deposition input values for the sites included 

in this synthesis range from 1.2 to 71.1 kg ha-1 yr-1 as 

throughfall, with the lowest values in the northern Sierra 

Nevada and the highest deposition in the western San 

Bernardino mountains in the Los Angeles air basin. Th e 

best defi ned empirical critical loads are for lichen eff ects, 

developed for forests of the Sierra Nevada, with the most 

protective critical load for incipient eff ects on lichens 

of 3.1 kg N ha-1 yr-1. Th is lower critical load is based 

on the low N deposition-site threshold concentration 

of N (1.0 percent N) in the widespread wolf lichen; at 

higher tissue concentrations of N, lichen community 

eff ects begin to increase (Fenn et al. 2008). Two 

additional lichen-based critical loads were determined 

for mixed-conifer forests in California: at throughfall 

N deposition of 5.2 kg ha-1 yr-1, oligotroph dominance 

shifts to mesotroph/eutroph dominance, and at 10.2 

kg ha-1 yr-1 oligotrophic species are extirpated. Th ese 

critical load estimates are considered as reliable (Table 

13.5). In applying lichens critical loads, the highest 

threshold—that of oligotroph extirpation—would not 

be used under conditions where the integrity of the 

lichen community was considered important. We do 

not include it in the summary of lichens for the U.S. 

(Table 19.4). Application of the western Oregon and 

Washington model (Geiser et al. 2010; Table 4.1) to 

the mixed-conifer forests yields a critical load estimate 

of 4 to 6 kg ha-1 yr-1 over a precipitation range of 41 

to 127 cm. Th is estimate uses the mid-range response 

threshold (oligotroph dominance shift) instead of the 

lowest threshold, as it assumes that eutrophs naturally 
Chapter 13—Mediterranean California GTR-NRS-80 157

Table 13.5—Empirical critical loads of nutrient N for mixed-conifer forest ecosystems of the Mediterranean California ecoregion. Reliability rating: ## reliable; # fairly reliable; (#) expert judgmentSite Critical load for 

N depositiona

kg N ha-1 yr-1

Reliability Response Comments Study

Sierra Nevada range

3.1-10.2b ## Lichen community 

changes

Critical load of 3.1 for exceedance of a N concentration threshold in Letharia vulpina; critical load of 5.2 for shift from oligotroph

to mesotroph/eutroph dominance of lichen communities; critical load of 10.2 for extirpation of oligotrophs

Fenn et al. 2008, 2010

Sierra Nevada range

4-6 # Lichen community 

changes

Modeled estimate using the oligotroph to mesotroph/eutroph dominance shift as response threshold and 41-127 cm precipitation range

Geiser et al. 2010 

San Bernardino mountains and southern Sierra Nevada range

17 ## [NO3

-] > 14 μM Based on regression of throughfall vs. peak 

streamwater NO3

- concentrations. A threshold of 

0.2 mg NO3-N L-1 (14.3 μM) identifi ed catchments 

beginning to “leak” N. Daycent simulations gave similar results

Fenn et al. 2008, 2010

San Bernardino mountains

17 # Reduced fi ne root 

biomass

Based on regression of throughfall N deposition and fi ne root biomass in ponderosa pine

Fenn et al. 2008, Grulke et al. 1998

San Bernardino mountains

25.9 # Soil acidifi cation Based on regression of throughfall N deposition 

and mineral soil H+. Soil pH of 4.6 was chosen as the critical value.

Breiner et al. 2007

San Bernardino mountains

24-33 (#) Biodiversity of 

understory: percent cover and no. of species/ha

Based on plant surveys in 1970s and 2003. N deposition data: Fenn et al., 2008 and unpublished data

Allen et al. 2007

San Bernardino mountains

39 (#) Forest sustainability Based on dramatic shifts in plant phenology and 

C allocation. Caused by combined effects of O3

and N deposition. Leads to increased bark beetle mortality and wildfi re risk. N deposition data from Fenn et al. 2008

Grulke and Balduman 1999; Grulke et al. 1998, 2009; Jones et al. 2004

aN deposition is reported as throughfall for all of the critical loads in this table.bUtilizing this critical load value would lead to a condition where lichen community composition had shifted toward dominance by eutrophs to the exclusion of oligotrophic 

(N-sensitive) species. Throughout this document, we use a threshold that limits shifts toward dominance by eutrophs. Thus, the value of 10.2 kg N ha-1 yr-1 does not 

appear in the chapter summarizing the critical loads for all ecoregions (Chapter 19).
158 Chapter 13—Mediterranean California GTR-NRS-80

comprise up to 47 percent of native lichen communities, 

due to the warm climate and prevalence of hardwood 

substrates. Th e critical load which would prevent 

elevated NO3

-

 leaching and further reductions in fi ne 

root biomass of ponderosa pine (Fenn et al. 2008) is 

estimated at 17 kg N ha-1 yr-1. Th e critical load of 17 kg 

N ha-1 yr-1 for reductions in fi ne root biomass (Table 

13.5) is preliminary; data from a greater number of 

sites are needed to refi ne the critical load. However, the 

available information on fi ne root biomass responses 

demonstrate that when N deposition is suffi cient to 

increase NO3

-

 leaching from soil and in streamwater, 

undesirable physiological eff ects are also apparent in 

plants. Th e critical load for streamwater NO3

-

 is reliable, 

while for fi ne root biomass eff ects a fairly reliable rating 

is assigned (Table 13.5).

Based on a critical mineral soil pH value of 4.6 

(MacDonald et al. 2002), an empirical critical load of 

26 kg N ha-1 yr-1 was also estimated and considered 

to be fairly reliable (Table 13.5), but the basis of this 

critical pH value has a tenuous connection to any fi rm 

ecosystem-specifi c biological or ecological eff ects or 

responses (Breiner et al. 2007, Wood et al. 2007). Th e 

estimated critical loads of 24 to 39 kg N ha-1 yr-1, which 

would maintain forest sustainability and biodiversity in 

the understory community, are included in Table 13.5 

and are only established as expert judgment. Th ese are 

uncertain, as few data are available from research sites in 

the case of forest sustainability (Grulke and Balduman 

1999, Grulke et al. 1998, Jones et al. 2004) and 

because of confounding factors in regard to understory 

biodiversity (Allen et al. 2007). Th e critical load for 

eff ects on ectomycorrhizal communities in California 

forests cannot be determined from current data because 

only two sites and a high N treatment level (150 kg ha-1

yr-1) were studied (Sirajuddin 2009).

13.5 Critical Loads Comparisons 

Across Ecosystem Types

Th e lowest critical load estimated for these California 

Mediterranean ecosystem types (Table 13.6) was 3.1 kg 

ha-1 yr-1 for the mixed-conifer forest. Th is value is based 

on the throughfall N deposition input leading to N 

concentrations in lichen thalli (Letharia vulpina) of the 

Sierra Nevada that exceed that of the low N deposition 

site threshold (Fenn et al. 2008). At a throughfall N 

deposition critical load of 3.5 to 5.9 kg ha-1 yr-1 the 

lichen community in mixed-conifer forests is predicted 

to shift from oligotroph dominance to mesotroph/

eutroph dominance. Th e lichen community responses to 

simulated N deposition resulted in a critical load of 3.1 

to 6.4 kg N ha-1 yr-1 for chaparral, oak woodland, and 

Central Valley lichen communities. Lichen communities 

of coastal sage scrub have not been studied with regard 

to air pollution eff ects. Th us, lichens are not yet an 

available bioindicator in coastal sage scrub.

Th e critical load for invasion of annual grasses in a 

nutrient-poor serpentine grassland (6 kg N ha-1 yr-1) 

was similar to the high end of the range for the lichen 

critical load in chaparral and forests (Table 13.6). Th is 

grassland critical load is based on a site to the west of 

the Central Valley and downwind of the San Jose/San 

Francisco metropolitan area. Th e serpentine grassland 

critical load may be a reasonable estimate for California 

grasslands in more fertile soils of the Central Valley 

and other regions of California, considering studies 

suggesting that serpentine grasslands are less responsive 

to added N than other California grasslands and are 

also less prone to exotic invasions (Harrison and Viers 

2007). Th e critical load for NO3

-

 leaching in chaparral 

catchments (10 kg ha-1 yr-1) is about 60 percent of that 

for mixed-conifer forest (17 kg ha-1 yr-1). Th e lower 

critical load in chaparral is believed to be due to the 

lower biological N demand and N retention capacity 

of chaparral catchments that are characterized by steep 

slopes and sandy soils. Another factor for the lower 

chaparral critical load may be the small size (4.3 ha) of 

the Chamise Creek catchment (Li et al. 2006) upon 

which the chaparral critical load of 10 kg ha-1 yr-1 is 

based. Smaller catchments are often associated with 

lower N retention capacity (Meixner and Fenn 2004). 

Th us, it isn’t clear if the critical load value for NO3

-

leaching from Chamise Creek is generally applicable to 

chaparral catchments.

For all vegetation types, catchment characteristics that 

increase the contact time of soil solution NO3

-

 within 

the soil profi le and with plant roots are expected to 

increase N uptake and increase the critical load. Stand 
Chapter 13—Mediterranean California GTR-NRS-80 159

Table 13.6—Empirical critical loads of nutrient N for the Mediterranean California ecoregion. Reliability rating: ## reliable; # fairly reliable; (#) expert judgmentEcosystem Critical load for 

N depositiona

kg N ha-1 yr-1

Reliability Response Comments Study

Coastal Sage Scrub

7.8-10 # Exotic invasive grass 

cover, native forb richness, arbuscular mycorrhizal richness

Modeled and inferential N deposition estimates and published data for mycorrhizae, unpublished data for vegetation survey.

Allen unpublished b, EgertonWarburton and Allen 2000, 

Tonnesen et al. 2007

Chaparral 3.1-14 # NO3 leaching; changes 

in lichen communities

Critical load for NO3 leaching of 10 kg N ha-1 yr-1 is 

based on one year of throughfall data in Chamise Creek and an additional year of throughfall data from adjacent Ash Mountain, both in Sequoia National Park. Lichen critical load of 5.5 kg N ha-1 yr-1 is from modeled N deposition data and 

published data for lichens. Lower estimated value for the lichen critical load of 3.1 kg N ha-1 yr-1 

is from application of a model developed in the Pacifi c NW including the effect of precipitation.

Fenn et al. 2003a; Fenn et al. 2003b; Fenn et al. 2010; Geiser et al. 2010; Jovan 2008; Jovan and McCune 2005; Meixner et al. 2006; Meixner and Fenn 2004; Riggan et al. 1985

Mixed-conifer forest

3-39 ## Lichen chemistry and 

community changes; NO3 leaching; soil 

acidifi cation; reduced fi ne root biomass

The lowest critical load is based on lichen tissue chemistry above the clean site threshold. NO3

leaching and fi ne root biomass critical load is 17 kg N ha-1 yr-1, and soil acidifi cation critical load is 

26 kg N ha-1 yr-1.

Allen et al. 2007; Breiner et al. 2007; Fenn et al. 2008 ; Fenn et al. 2010 ; Grulke and Balduman 1999; Grulke et al. 1998, 2009; Jones et al. 2004 

Serpentine grassland

6 ## Annual grass invasion, 

replacing native herbs

Critical load based on a local roadside gradient; Serpentine grassland site is actually west of the Central Valley.

Weiss 1999; Fenn et al. 2010

aN deposition in the mixed-conifer sites is reported as throughfall. Critical loads for coastal sage scrub and for lichen communities in chaparral (5.5 kg N ha-1 yr-1) are based on 

CMAQ simulated N deposition.

bSee footnote 15 on page 144
160 Chapter 13—Mediterranean California GTR-NRS-80

management is also expected to aff ect the critical load 

for a given site. For example, a decrease in the site N 

capital due to fi re or stand thinning operation, followed 

by vigorous vegetation regrowth and increased N 

demand, could increase site N retention and thus lead 

to a higher critical load. Th e critical load for biological 

eff ects in coastal sage scrub is relatively low (10 to 11 

kg ha-1 yr-1) and may be lower, considering the lack 

of low N deposition control sites in estimating the 

critical loads. A major factor contributing to the N 

sensitivity of coastal sage scrub vegetation is the fact that 

N accumulates to high concentrations in the surface 

soils during the long summer dry period and the entire 

rooting zone is probably only thoroughly leached of 

NO3

-

 in unusually high rainfall years (Wood et al. 

2006).

Mediterranean ecosystems exposed to chronic 

N deposition often cycle between periods of 

N accumulation and periods of N availability, 

mobilization, uptake, and loss. Th is latter period is 

presumably the season of greatest ecosystem sensitivity 

to the eff ects of excess N. Critical loads estimates are 

needed that protect ecosystems from detrimental eff ects 

during these periods of greatest sensitivity. Th e empirical 

critical loads in this study are based on the predominant 

ecological conditions, which include fi re suppression, 

co-occurrence of O3

 and chronic N deposition, and 

recurring drought. Critical load, by defi nition, refers to 

the threshold deposition below which adverse biological 

eff ects are not observed. However, some critical loads in 

this chapter (and others) are tied to ecosystem processes 

or responses that don’t always have a biological link 

within the terrestrial ecosystem that has yet been well 

established or defi ned (e.g., nitrate leaching).

13.6 Comparison to European 

Critical Loads

Few empirical N critical loads studies for European 

Mediterranean ecosystems have been published. Based 

on a European-wide modeling eff ort, Kuylenstierna 

et al. (1998) concluded that Mediterranean scrub has 

low to moderate sensitivity to N deposition, with a 

lower critical load of 15 kg N ha-1 yr-1. A relatively 

small area of the Mediterranean scrub exceeds this 

level of deposition (Kuylenstierna et al. 1998). Lorenz 

et al. (2008) used the simple mass balance approach 

and reported low nutrient N critical load values for 

Spain, ranging from 3.5 to 10.5 kg ha-1 yr-1, or possibly 

even lower in some instances. Th ese modeled critical 

load values appear to be lower than expected based on 

empirical studies (Fenn et al. 2008, Michopoulos et al. 

2008, this chapter). Modeled deposition to coniferous 

forest for the year 2000 shows N deposition of 10 to 

20 kg ha-1 yr-1 over most of Mediterranean Europe 

(Simpson et al. 2006), showing the potential for 

exceedance of the N critical load in many forested sites.

A critical load of 5 kg N ha-1 yr-1 as bulk deposition in 

forest clearings can be estimated for Spanish fi r (Abies 

pinsapo) forests in southern Spain (Torres-Cañabate 

et al. 2008). Th roughfall N deposition from an earlier 

study was reportedly 12 kg N ha-1 yr-1 in these forests 

(Torres-Cañabate et al. 2008). At the reported level of 

bulk deposition, N enrichment of soil and foliage were 

evident, along with increased potential net nitrifi cation 

and emissions of N2O from soil compared to a more 

N-limited site. However, bulk N deposition was higher 

at the N-limited site than the site showing symptoms 

of N saturation, making interpretation of these fi ndings 

diffi cult (Torres-Cañabate et al. 2008). We estimate 

a tentative N critical load for incipient NO3

-

 leaching 

at a beech forest in central eastern Greece of 13 kg 

N ha-1 yr-1, based on data reported by Michopoulos 

et al. (2008), but more data are needed to determine 

if the critical load is actually higher. Catchments in 

northern Italy along the Swiss border (Rogora 2007) are 

strongly N saturated (at level 2-3 of Stoddard’s [1994] 

classifi cation), but these are not typical Mediterranean 

ecosystems. A notable example of a Mediterranean 

forest where deposition exceeds the N critical load is 

an Aleppo pine site in Athens, Greece (Michopoulos 

et al. 2004). Because this site is presumably well above 

the critical load, with annual throughfall deposition 

as high as 38 kg N ha-1 yr-1, the critical load cannot be 

determined.

In summary, few European Mediterranean sites have 

been reported to be above the N critical loads based on 

empirical data. On the other hand, broad-scale models 
Chapter 13—Mediterranean California GTR-NRS-80 161

of N deposition and nutrient N critical loads predict 

that critical loads exceedance is not uncommon in the 

European Mediterranean region (Lorenz et al. 2008, 

Simpson et al. 2006). Lorenz et al. (2008) calculated low 

critical load values in Spain because of low plant growth 

rates, low timber harvest, and low levels of N export in 

seepage water due to low precipitation, resulting in low 

net N export from the ecosystems. It was concluded 

that, in these cases, the critical load can be exceeded 

by relatively low N deposition. However, modeled 

deposition and critical loads need to be validated 

with empirical data, particularly considering the cooccurrence of elevated O3 levels and the pronounced 

seasonality in precipitation and N deposition inputs 

and outputs in these unique Mediterranean ecosystems 

(Fenn et al. 2008). DeVries et al. (2007) also 

recommended that for these cases the empirical critical 

loads for N should be determined.

Th e N critical loads in European Mediterranean 

ecosystems appear to be generally similar to those 

in California, but the lack of reported empirical 

critical loads for Mediterranean Europe prevent any 

strong comparisons. It seems safe to assume that in 

both regions, lichens are the most sensitive terrestrial 

responders to N deposition, and are impacted in regions 

of suffi cient exposure. Lichen responses to NH3 have 

been reported from Italy and Portugal, but critical load 

estimates are not available due to lack of deposition 

data (Frati et al. 2008, Pinho et al. 2009). However, 

recent studies on the responses of epiphytic lichen 

communities to NH3 exposure indicate similar dose 

responses in Europe and the Mediterranean California 

ecoregion. In an oak woodland in Portugal, also with a 

Mediterranean climate, the estimated critical level for 

lichen community diversity impacts was 1.4 to 1.7 μg 

m-3 (Pinho et al. 2009). Th e estimated critical level for 

epiphytic lichen community changes in mixed-conifer 

forests in the southern Sierra Nevada of California 

falls within this same range.20 Recently a critical level 

of 1.0 μg m-3 as a long-term average concentration 

was proposed for NH3

 eff ects on epiphytic lichen 

communities based on European studies (Cape et 

al. 2009, Sutton et al. 2009) and a critical level of 

3.0 μg m-3 was proposed for eff ects on herbaceous 

plant communities (Cape et al. 2009). Th e California 

serpentine grassland results presented in this chapter are 

in general agreement with the herbaceous plant critical 

level for NH3

. Monthly average NH3 concentrations 

were 0.3 to 0.7 μg m-3 at a site where impacts were 

minimal, compared to NH3 concentrations of 1 to 3.5 

μg m-3 at sites where plant communities were strongly 

impacted by invasive species.21

However, in regard to broader ecosystem eff ects from 

excess N, as suggested by Rodà et al. (2002) and in 

contrast to the conclusions of Lorenz et al. (2008) 

cited above, the N-limited state of some European 

Mediterranean ecosystems may be enhanced by 

the repeated nutrient losses from fi re and biomass 

harvesting over the centuries. Under these N-demanding 

conditions, the critical loads for soil, vegetative, and 

hydrologic responses to N deposition may be higher 

than those reported for California. Th ese eff ects of 

land-use history on the critical loads are expected to vary 

across the landscape. Information on critical loads for all 

vegetation types in Mediterranean ecosystems has been 

identifi ed as a major research gap (Bobbink et al. 2003). 

Because of the dearth of critical load work in European 

Mediterranean ecosystems, the work summarized herein 

from California Mediterranean ecosystems may be 

useful to the European scientifi c community.

13.7 Future Research Directions and 

Gaps in Data

To better understand the eff ects of chronic N deposition 

in Mediterranean ecosystems, further research is needed 

on levels of N input and mechanisms and fl uxes of 

N loss and storage, the eff ects of fi re on N pools and 

processing, and the response of plant, lichen, and 

mycorrhizal communities to deposition. Gaseous N losses 

have not been well quantifi ed and may be an important 

process by which a signifi cant proportion of the excess N 

accumulation is reduced (Anderson and Poth 1989, Fenn 

21Weiss, S.B.; Bytnerowicz, A. Unpublished data on fi le with 

Mark Fenn, research plant pathologist, Forest Fire Laboratory, 

4955 Canyon Crest Dr., Riverside, CA 92507.

20Bytnerowicz, A.; Jovan, S. Unpublished data on fi le 

with Andrzej Bytnerowicz, research ecologist, Forest Fire 

Laboratory, 4955 Canyon Crest Dr., Riverside, CA 92507.
162 Chapter 13—Mediterranean California GTR-NRS-80

et al. 1996, Fenn and Poth 2001). A large proportion of 

trace gas losses are expected to occur in the riparian and 

near-riparian zones, possibly as degassing of dissolved 

N2

O from groundwater-fed streams. Further streamwater 

measurements are needed across N deposition gradients 

using Δ17O as a tracer for atmospheric NO3

-

 (Michalski et 

al. 2004) to determine the threshold percent of exported 

NO3

-

 as direct throughput of unassimilated atmospheric 

NO3

-

 associated with N-saturated catchments. As stands 

with lower air pollution or N fertilization at realistic levels 

are studied, we may fi nd that lower levels of N deposition 

also have eff ects on ecosystem dynamics. Research is 

needed on the best management practices for improving 

water quality in chaparral and mixed-conifer forest 

catchments. Questions remain regarding the level of N 

deposition reduction necessary to reverse the symptoms 

of saturation. It would be valuable to characterize the 

spatial patterns of N deposition throughout catchments, 

but this is diffi cult because of the steep terrain in many 

chaparral and forested catchments.

Research is needed on the eff ects of fi re in 

Mediterranean ecosystems. Fire is a natural part of 

Mediterranean ecosystems, and its interactions with N 

deposition are complex. In coastal sage scrub, fi re may 

interact with increased N deposition to promote the 

growth of exotic grass fuel and cause rapid degradation 

in native vegetation (Allen et al. 1998, Minnich and 

Dezzani 1998, Talluto and Suding 2008). In chaparral 

and mixed-conifer forests, fi re in combination with 

decreased N deposition may help mitigate N saturation 

(Gimeno et al. 2009). Studies are needed to test the 

eff ectiveness of periodic fi re as a management tool in 

N-saturated catchments, considering that 65 to 85 and 

80 to 95 percent of site N capital in forest (Johnson 

et al. 2009) and chaparral (Rundel and Parsons 1980, 

Rundel and Vankat 1989) ecosystems is stored in the 

mineral soil, which is unimpacted by fi re disturbance 

(Meixner et al. 2006). Recurring fi re is a normal process 

in California chaparral and forest ecosystems and 

prescribed fi re can be applied without causing signifi cant 

alterations in nutrient cycling or plant nutrient status 

(Murphy et al. 2006).

Little is known about the eff ects of N deposition 

on plant biodiversity in chaparral and in the forest 

understory, including the putative indirect eff ects on 

plant communities as a result of the impacts of N 

deposition on mycorrhizal communities. Nitrogen 

addition studies at appropriate doses are also needed to 

determine the N critical load for biodiversity eff ects in 

California grasslands. Additional analysis is needed to 

separate lichen responses due to NH3

 from those caused 

by other forms of N deposition in the oak woodlands, 

chaparral, and forest communities, and to account for 

the infl uence of precipitation on lichen responses to 

N deposition over the regional landscape. Th e original 

interpretation of lichen community work in the greater 

Central Valley was limited by available air pollution 

data. Th e simulated CMAQ deposition data presented 

herein has allowed for preliminary estimates of lichenbased critical loads in chaparral and oak woodlands. Th is 

should be complemented with empirical deposition data 

to better defi ne the critical load for lichen community 

responses.

More work is needed to determine the critical load 

for acidity in chaparral ecosystems. Evidence strongly 

suggests that atmospheric deposition has acidifi ed 

chaparral soils over the past several decades, but N 

deposition data are not available from enough sites to 

determine empirical relationships. Modeled estimates 

of the critical loads for acidity have not yet been made 

for chaparral ecosystems. Better defi nition is needed of 

the N input levels that contribute to decreases in forest 

sustainability by way of physiological perturbations, 

greater sensitivity to bark beetles, and increased drought 

stress and mortality (Grulke et al. 2009). Further 

comparisons of modeled and empirical critical loads are 

needed for California ecosystems and for Mediterranean 

ecosystems in general.

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170
Chapter 14—Southern Semi-Arid Highlands GTR-NRS-80 171

14 SOUTHERN SEMI-ARID HIGHLANDS

M.E. Fenn, E.B. Allen

In the uplands, ecosystem responses to N at current 

deposition inputs likely include changes in lichen 

communities. We hypothesize that if N deposition 

increased two- to threefold (to approximately 12 to 18 

kg ha-1 yr-1), many eff ects reported for mixed conifer 

forests (see Chapter 13) would become apparent within 

a few years as N accumulates in the ecosystem. Th ese 

expected eff ects would include elevated nitrate (NO3

-

) 

leaching and nitrogenous trace gas emissions from 

soil, impaired root production, altered mycorrhizal 

community composition and function, increased 

susceptibility to pests, and possible eff ects on understory 

biodiversity.

14.3 Range of Responses Observed

As we were not able to identify U.S. studies 

documenting a particular ecosystem response at a given 

level of N input for this ecoregion, we focus instead 

on the reported deposition for the U.S. ecoregion and 

documented ecosystem responses not linked to a specifi c 

N level.

Th e Chiricahua National Monument, a semi-arid site 

with sparse vegetation cover, is designated a Class I 

airshed under the Clean Air Act, with management 

objectives of clear views to 100 miles in any direction 

(NPS 2006). Th reats to air quality include smelting and 

power plants in Mexico, but especially future plans for 

power plants within 50 miles of the border. Because 

both smelting and coal-fi red power plants produce 

sulfur oxides (SOx

), the greater ecological threats in past 

decades may have come from sulfur (S) rather than N 

deposition. However, based on temporal trends in the 

National Atmospheric Deposition Program (NADP) 

and Clean Air Status and Trends Network (CASTNET) 

data at the Chiricahua site, N deposition is now greater 

than S deposition. Sulfur emissions have decreased in 

this area since around 1999 (US EPA 2008b), when 

smelters in Playas, New Mexico, and Cananea, Sonora, 

Mexico, closed. Total S deposition over the past 

several years as reported by NADP and CASTNET 

14.1 Ecoregion description

Most of the Southern Semi-Arid Highlands ecoregion 

occurs in Mexico, but a portion centered on the 

Chiricahua Mountains of Arizona and New Mexico 

lies in the United States. Th e ecoregion description is 

adapted from CEC (1997). Rainfall is 300 to 600 mm, 

with biseasonal distribution (winter and summer). Th e 

mountains are volcanic in origin, with valleys and plains 

of alluvial sediments. Th e area lies at the intersection 

of the Sonoran and Chihuahuan Deserts, the Rocky 

Mountains, and the Sierra Madre (CEC 1997, Fig. 2.1). 

Consequently, this region is high in plant and animal 

diversity, having components of all four ecosystem 

types. Vegetation includes desert grassland interspersed 

with desert scrub (especially mesquite [Prosopis spp.] 

and acacia [Acacia spp.]) at low elevations, oak (Quercus 

spp.) and juniper (Juniperus spp.) at intermediate 

elevations, and coniferous forests at high elevations.

14.2 Ecosystem Responses to N 

Deposition

Potential nitrogen (N) deposition impacts to the 

southern highlands are likely to be similar to those 

described for Mediterranean California and North 

American Deserts ecoregions (Chapters 13 and 12), 

where similar vegetation types occur. In the lowlands, 

ecosystem responses to N deposition might include 

increases in exotic species, an eff ect which can lead to 

the accumulation of fi re-sustaining fuel loads and thus 

increased fi re frequency and severity. Other potential 

eff ects to lowland ecosystems include alteration 

of mycorrhizal fungal diversity, abundance, and 

functioning; shifts in lichen communities; and alteration 

of biotic crust composition as reported for other desert 

areas (see Chapter 12).

At intermediate elevations, ecosystem responses 

to N deposition might include alteration of lichen 

community composition or functional groups. Further 

increases in N deposition might aff ect mycorrhizae and 

herbaceous plant communities.
172 Chapter 14—Southern Semi-Arid Highlands GTR-NRS-80

is approximately 1.2 kg ha-1 yr-1. Even if the dry 

deposition of S is underestimated, S deposition levels 

would still be considered low.

Th e NADP and CASTNET data report total annual N 

deposition of 2 to 3 kg ha-1 yr-1 over the past 15 years 

at the Chiricahua site (US EPA 2008a). Average annual 

precipitation at the monitoring site is approximately 

400 mm. Reported values of wet N deposition are 

three times greater than those for dry N deposition, 

suggesting that the CASTNET modeling approach 

is underestimating dry deposition and thus total 

deposition, as has been reported by CASTNET data for 

other semi-arid sites in the western United States (Fenn 

et al. 2009). According to NADP data, wet deposition 

of ammonium (NH4

+

) is similar to wet deposition of 

NO3

-

. However, dry deposition of ammonia (NH3) 

is not measured, and as a result, total N deposition is 

almost certainly greatly underestimated.

In our judgment, bulk N deposition at the Chiricahua 

NADP site is likely closer to 4 to 5 kg ha-1 yr-1. 

Depending on leaf area index, deposition to canopies 

in areas with signifi cant vegetation cover would be 

even higher. Nitrogen deposition in this region was 

estimated to be 4 to 7 kg ha-1 yr-1, by the Environmental 

Protection Agency’s Models-3/Community Multiscale 

Air Quality (CMAQ) model (Byun and Schere 2006) 

and 2002 emissions data (Fenn et al. 2003b). Based on 

these estimates, we hypothesize that N deposition in the 

Chiricahua National Monument occurs at levels that 

can aff ect sensitive ecosystem components such as lichen 

community composition. Other possible eff ects of 

incipient N enrichment are community shifts of annual 

plant species (Fenn et al. 2003a). However, lichen 

communities likely have also been aff ected by historical 

S emissions from copper smelters in the region.

In the San Simon Valley, near Portal, Arizona, and 

adjacent to the Chiricahua National Monument, the 

winter annual plant community has undergone marked 

community change associated with the sustained 

proliferation of stork’s bill (Erodium cicutarium), the 

Eurasian invasive winter annual plant (Schutzenhofer 

and Valone 2006). As a result, winter annual 

plant diversity has declined drastically and species 

composition has shifted. Possible factors contributing 

to these plant community changes are changes in 

climate, including decreased winter precipitation and 

more frequent drought, N deposition from agricultural 

fi elds and copper smelters (Bytnerowicz 200922), and 

changes in the rodent community (Ernest et al. 2000). 

Preliminary atmospheric concentration data for NO3

-

and nitric acid (HNO3) measured from December 2006 

to January 2007 in the San Simon Valley were higher 

than expected for a remote site.22

14.4 Future Research Directions and 

Gaps in Data

Little is known about the eff ects of N deposition in 

the Southern Semi-Arid Highlands, as they have not 

been studied. N deposition needs to be measured 

within the major ecosystem types. Ecological eff ects 

should be investigated by surveying for sites showing 

N accumulation and enrichment associated with N 

deposition. Sulfur deposition levels and eff ects from 

past and current S deposition should also be evaluated. 

Where lichen communities occur, they can be surveyed 

for community shifts and S and N accumulation as 

early sentinels of ecosystem change and response to air 

pollution.

22Bytnerowicz, A. 2009. Personal communication. Ecologist, 

Pacifi c Southwest Research Station Forest Fire Laboratory, 

4955 Canyon Crest Drive, Riverside, CA 92507.
Chapter 14—Southern Semi-Arid Highlands GTR-NRS-80 173

LITERATURE CITED

Byun, D.; Schere, K.L. 2006. Review of the governing 

equations, computational algorithms, and 

other components of the Models-3 Community 

Multiscale Air Quality (CMAQ) modeling system.

Applied Mechanics Reviews. 59: 51-77.

CEC (Commission for Environmental Cooperation). 

1997. Ecological regions of North America. 

Toward a common perspective. Montreal, Canada: 

Commission for Environmental Cooperation. 

71 p. Available at http://www.cec.org/fi les/pdf/

BIODIVERSITY/eco-eng_EN.pdf (May 24, 2010).

Ernest, S.K.M.; Brown, J.H.; Parmenter, R.R. 2000. 

Rodents, plants, and precipitation: spatial and 

temporal dynamics of consumers and resources.

Oikos. 88: 470-482.

Fenn, M.E.; Baron, J.S.; Allen, E.B.; Rueth, H.M.; 

Nydick, K.R.; Geiser, L.; Bowman, W.D.; Sickman, 

J.O.; Meixner, T.; Johnson, D.W.; and Neitlich, P. 

2003a. Ecological eff ects of nitrogen deposition in 

the western United States. BioScience. 53: 404-420.

Fenn, M.E.; Haeuber, R.; Tonnesen, G.S.; Baron, 

J.S.; Grossman-Clarke, S.; Hope, D.; Jaff e, D.A.; 

Copeland, S.; Geiser, L.; Rueth, H.M.; Sickman, 

J.O. 2003b. Nitrogen emissions, deposition, 

and monitoring in the western United States. 

BioScience. 53: 391-403.

Fenn, M.E.; Sickman, J.O.; Bytnerowicz, A.; Clow, 

D.W.; Molotch, N.P.; Pleim, J.E.; Tonnesen, 

G.S.; Weathers, K.C.; Padgett, P.E.; Campbell.; 

D.H. 2009. Methods for measuring atmospheric 

nitrogen deposition inputs in arid and montane 

ecosystems of western North America. In: Legge, 

A.H., ed. Developments in environmental science. 

Relating atmospheric source apportionment 

to vegetation eff ects: Establishing cause eff ect 

relationships. Vol. 9. Amsterdam: Elsevier: 179-228.

NPS (National Park Service). 2006. Chiricahua 

National Monument—air quality. Washington, 

DC: U.S. Department of Interior, National Park 

Service, Chiricahua National Monument. Available 

at http://www.nps.gov/chir/naturescience/airquality.

htm (Accessed May 12, 2010).

Schutzenhofer, M.R.; Valone, T.J. 2006. Positive and 

negative eff ects of exotic Erodium cicutarium on 

an arid ecosystem. Biological Conservation. 132: 

376-381.

US EPA (U.S. Environmental Protection Agency). 

2008a. Clean air status and trends network—

Chiricahua, total N. Washington, DC: U.S. 

Environmental Protection Agency. Available at 

http://epa.gov/castnet/javaweb/charts/CHA467_totn.

png (Accessed March 28, 2011).

US EPA (U.S. Environmental Protection Agency). 

2008b. Clean air status and trends network—

Chiricahua, total S. Washington, DC: U.S. 

Environmental Protection Agency. Available at 

http://epa.gov/castnet/javaweb/charts/CHA467_tots.

png (Accessed March 28, 2011).
174
Chapter 15—Temperate Sierra GTR-NRS-80 175

15 TEMPERATE SIERRAS

M.E. Fenn, L.H. Geiser

15.3 Range of Responses Observed

Little research has been done on N deposition eff ects in 

forests of the Temperate Sierras ecoregion in Arizona 

and New Mexico. However, a combination of modeling 

and monitoring were used to estimate N deposition 

fl uxes for forests northeast of Phoenix, Arizona, a rapidly 

expanding metropolis (Fenn et al. 2003). Estimated 

deposition fl uxes to the forested area ranged from 13 

to 23 kg ha-1 yr-1. Further measurements are needed to 

confi rm these data and to determine the extent of the 

aff ected areas, but these estimates suggest that current 

levels of N deposition are likely to have signifi cant 

impacts on these forests. Th e spatial extent of aff ected 

areas may be relatively limited.

Because we found no studies relating N inputs to 

ecosystem responses in the United States, we summarize 

research conducted in the Mexico City air basin (Fig. 

15.1) within this ecoregion, as it may shed some light 

on the levels at which responses may be observed. 

Mexico City (Fig. 15.1) is located within the larger 

Basin of Mexico, sometimes called the valley of Mexico. 

Th e 9600 km2

 basin is defi ned by mountain ranges 

surrounding the valley fl oor. Th e Mexico City air basin 

refers to the ground-level atmosphere of the basin. In 

Mexican mountain pine (Pinus hartwegii) forests in 

the Mexico City air basin, the primary documented 

response to N deposition was nitrate (NO3

-

) leaching 

(Fenn et al. 1999, Fenn et al. 2002b). Foliar N was 

sometimes higher in the more polluted sites, but this 

result was inconsistent. Also, soil N levels did not always 

correspond with N deposition (Fenn et al. 1999, 2002b, 

2006). Th e inconsistent response of foliar and soil N 

may be due to the inherent high soil N fertility of the 

volcanic soils prevalent in the Mexico City air basin 

(Fenn et al. 2002b, 2006). Th e role of N deposition 

in the historical and current deterioration of native 

lichen communities is unclear, as lichen community 

impacts in the Mexico City Air Basin began decades ago 

15.1 Ecoregion Description

Much of the Temperate Sierras ecoregion is located 

in Mexico, but a northern portion of this ecoregion is 

found in the southwestern United States (mainly in 

Arizona and New Mexico) (CEC 1997, Fig. 2.1). Th e 

ecoregion description is adapted from CEC (1997). 

Forested montane regions are typically surrounded 

at lower elevations by arid and semi-arid deserts, 

scrublands, or grasslands. Vegetation can be evergreen 

or deciduous, and is primarily composed of conifers 

and oaks which grow from 10 to 30 m, sometimes 

reaching 50 m. Th is vegetative cover may comprise 

from one to three tree layers, one or two shrub layers, 

and an herbaceous stratum. Th is forest community is 

characterized by about 3000 vascular plant species, 30 

percent of which are endemic to Mexico (CEC 1997).

15.2 Ecosystem Responses to N 

Deposition

Responses to N deposition in this ecoregion (Fenn 

et al. 1999, 2002a, 2002b) appear to be similar to 

those in the mixed conifer forest of the Mediterranean 

California ecoregion (Chapter 13). Th ese eff ects 

include elevated foliar N concentrations, increased 

soil N cycling rates, increased N export in solution 

and via trace gas emissions, increased susceptibility to 

pests, changes in native species composition, enhanced 

establishment of exotic invasive species, and changes 

in mycorrhizal and epiphytic lichen communities. 

Although data on N deposition and eff ects are scarce 

for the U.S. portion of the Temperate Sierras ecoregion, 

under current simulated estimates (3 to 6 kg N ha-1 

yr-1) of total N deposition (Fenn et al. 2003) and 

National Atmospheric Deposition Program (NADP) 

wet deposition data, we expect that the most likely 

eff ects occurring in some areas under current deposition 

levels are likely to be eff ects on lichen communities 

and possibly eff ects on herbaceous or grass understory 

communities.
176 Chapter 15—Temperate Sierra GTR-NRS-80

when sulfur (S) emissions and deposition were much 

greater than current levels and greater than nitrogen 

oxides (NOx

) emissions. However, it is likely that as S 

deposition levels have decreased in the Mexico City air 

basin in the past decade (Fenn et al. 2002a), the relative 

importance of N deposition on the remaining lichen 

communities has increased.

In the Mexico City air basin, NO3

-

 concentrations 

in stream and spring water were highly elevated in 

forested catchments south and southwest of Mexico 

City. Nitrogen deposition in throughfall at the highly 

N-saturated Desierto de los Leones National Park 

was 18.5 kg ha-1 yr-1 in 1996/1997 (Fenn et al. 1999). 

Nitrate concentrations in springs and streamwater at 

this site were frequently 20 to 90 μeq L-1 (Fenn et al. 

2002b), indicating that this site is strongly N saturated. 

Elevated NO3

-

 concentrations were also found at the 

Ajusco site, where N deposition is believed to be lower 

than at the Desierto site (Fenn et al. 2002a), but annual 

deposition fl uxes have not been quantifi ed. Because the 

Andisols in these study sites are typically high in N, 

even without elevated N deposition, no clear patterns 

of increased net N mineralization, nitrifi cation, soil 

N or foliar N were observed (Fenn et al. 2002b). 

Nitrogen fertilization sometimes results in little plant 

growth response (Fenn et al. 2002b, 2006); preliminary 

evidence suggests a depressive growth eff ect from 

added N in the Mexico City air basin in some cases, 

possibly because added N induces greater phosphorus 

(P) defi ciency (Fenn et al. 2006, 2002b). Widespread 

P defi ciency, presumably because of the high P-fi xation 

capacity of these Andisols (Shoji et al. 1993), and 

the high inherent N fertility in these study sites, may 

enhance the tendency of these catchments to leach NO3

-

with increasing atmospheric N deposition.

Figure 15.1—The Mexico City air basin, showing 

major mountain ranges, towns, and Mexico City. 

The high N deposition forested sites studied, 

Desierto de los Leones and Ajusco, are in the 

southwestern portion of the basin. The low 

deposition forest sites are in the southeastern 

portion of the basin within the Rio Frio and 

Nevada mountain ranges.

TEZONTLALPAN RANGE

TEPOTZOTLAN RANGE

LAS CRUCES RANGE

ZEMPOALA RANGE

CHICHINAUTZIN RANGE

NEVADA RANGE

RIO FRIO

RANGE

PACHUCA RANGE

RANGE

GUADALUPE

SANTA CATARINA RANGE

PATLICHIQUE RANGE

PITALLAS RANGE

Pachuca

Tizayuca

Zumpango

Cuautitlán

Xochimilco

Chalco

Amecameca

Texcoco

Teotihuacán

Calpulalpan

19°30'

98°45'

0 10 20

KILOMETERS

TEPOZTECO HILL

POPOCATEPETL VOLCANO

TEPOZAN HILL

L E G E N D

Basin of Mexico

Valley floor

(2300 m)

Town

98°45'

19°30'

divide

IZTACCIHUATL VOLCANO

PASO DE CORTES

AJUSCO

DESIERTO DE

LOS LEONES

MEXICO

CITY

Texcoco

Lake

















Chapter 15—Temperate Sierra GTR-NRS-80 177

At the Ajusco Mexican mountain pine site located 

south and downwind of Mexico City, an unknown 

soil factor was highly toxic to growth of river redgum 

(Eucalyptus camaldulensis) seedlings and the percentage 

of short roots colonized by putative symbiotic fungi 

in association with external mycelium was severely 

reduced (Fenn et al. 2006, Perea-Estrada et al. 2005). 

In the absence of fertilization, the percentage of 

eucalyptus roots that were ecto- and endomycorrhizal 

was consistently lower in high deposition sites. When 

eucalpytus seedlings were fertilized with P alone or a 

combination of N and P, a strong positive plant growth 

response occurred and the levels of root-associated fungi 

no longer diff ered between the polluted and clean sites 

(Perea-Estrada et al. 2005). Furthermore, at two of the 

three study sites in the Mexico City air basin, added N 

caused a signifi cant reduction in the number of short 

roots of eucalyptus seedlings colonized by symbiotic 

fungi with external mycelium (Perea-Estrada et al. 

2005). Further mechanistic studies are needed before 

fi rm conclusions can be reached regarding the role of N 

deposition in the strongly growth-depressive eff ects of 

soil from the Ajusco site. However, the negative eff ects 

of added N on symbiotic root fungi and the positive 

eff ects of added P on plant growth and root-associated 

fungi suggest that excess N may exacerbate P defi ciency, 

particularly if symbiotic fungi important for P uptake 

are inhibited by high N levels in soil.

15.3.1 Lichens

Lichen responses to air pollutants in the Temperate 

Sierras ecoregion have not been studied, except in the 

Mexico City air basin. A series of studies contrasted fi r 

and oak forests in Desierto de los Leones and Sierra de 

las Cruces (both south of Mexico City) to those of El 

Chico National Park, a relatively unpolluted site about 

100 km to the northeast. Epiphytic lichen diversity 

at Desierto de los Leones is severely impoverished 

compared to El Chico National Park. Desierto de los 

Leones may have lost nearly 50 percent of its lichen 

species and lichen abundance is reduced by 60 percent, 

presumably as a result of the severe air pollution at 

this site. Based on historical herbarium collections, the 

decline in lichen diversity appears to coincide with the 

period of accelerated industrial and population growth 

of Mexico City since the 1930s and 1940s (Zambrano 

et al. 2002). Short-term lichen transplant experiments 

in Sierra de las Cruces, a montane region just southsouthwest and downwind of the urbanized zone, 

showed 30 percent lower carbon fi xation and 15 to 25 

percent chlorophyll degradation compared to samples 

in El Chico National Park (Zambrano and Nash 2000, 

Zambrano et al. 1999). Th ese results suggested that 

chronic air pollution (ozone (O3), N- and S-containing 

compounds) is a major cause of lichen decline in 

forests surrounding the city, along with a variety of 

other anthropogenic disturbance factors (Zambrano et 

al. 2002). In the most recent study, Valencia-Islas et 

al. (2007) showed that the composition and location 

of phenolics in thalli of the sensitive parmotrema 

lichen (Parmotrema stuppeum) and tolerant Asahina’s 

cartilage lichen (Ramalina asahinae) was consistent 

with their contrasting survival in forests stressed by 

O3 (oxidative air pollution). Such responses illustrate 

potential confounding eff ects to lichen communities 

by codominant acidifying and oxidative pollutants in 

contrast to environments dominated solely by nutrient 

N.

15.4 Critical Loads Estimates

We estimate a N critical load for NO3

-

 leaching from 

pine stands in the Mexico City air basin of 15 kg ha-1 yr-1

(Table 15.1). Th is critical load is based on streamwater 

data from two regions of low N deposition east of 

Mexico City and two regions south and southwest of 

Mexico City (Fenn et al. 1999, 2002b). At the Desierto 

de los Leones site, where throughfall N deposition was 

18.5 kg ha-1 yr-1 in 1996/1997, NO3

-

 concentrations in 

springs and streamwater were frequently 20 to 90 μeq L-1

(Fenn et al. 2002b), indicating that this site is strongly 

N saturated and well above the critical load. Elevated 

NO3

-

 concentrations were also found at the Ajusco site, 

where N deposition is believed to be lower than at the 

Desierto site (Fenn et al. 2002a), but annual deposition 

fl uxes have not been quantifi ed. Uncertainty in this 

estimated critical load is largely due to an insuffi cient 

number of study sites spanning a range of deposition 

fl uxes and because deposition data are from only 1 year 

for throughfall from one high and one low deposition 

site (Fenn et al. 1999). Also note that high levels of NO3

-

leaching in the Mexico City air basin occur both during 
178 Chapter 15—Temperate Sierra GTR-NRS-80

the rainy summer growing season and during the winter 

dry season (Fenn et al. 1999, 2002b).

Although lichen species have disappeared and lichen 

communities have been severely altered in the Mexico 

City air basin, it is diffi cult to estimate a N critical 

load for lichen eff ects because of the lack of historical 

pollution exposure and deposition data and the potential 

eff ects of multiple pollutants. Emissions of S, N and 

heavy metals have been historically high; O3 exposure 

until recently has been among the highest in the world 

(Fenn et al. 2002a). Geiser et al. (2010) developed a 

method for relating N deposition to lichen community 

response. Th ey select a threshold for the shift in the 

community composition from oligotroph- to eutrophdominated. Th is approach is described in detail in 

Chapter 4. For U.S. portions of the ecoregion, our best 

estimate of N critical loads for epiphytic lichens is 4 to 7 

kg ha-1 yr-1. Th is estimate assumes a response threshold 

of about 60 percent eutrophs, mean annual precipitation 

of 30 to 180 cm, and general applicability of the Geiser 

et al. (2010) model to temperate forests.

15.5 Comparison to Critical Loads for 

Other Regions

Th ere are no relevant lichen data from Europe for 

comparison with this ecoregion; critical loads developed 

for North American deserts or Northwestern Forested 

Mountains may have some applications. Th e lichen 

data from Mexico, summarized above, demonstrate 

a response in lichen community structure, but no 

conclusions can be made regarding a N critical load 

because trace metals, O3, and S and N deposition were 

all at high levels during the decades of rapid industrial 

and population growth in the Mexico City air basin 

(Fenn et al. 2002a).

Th e estimated critical load for NO3

-

 leaching in the 

Mexico City air basin (15 kg ha-1 yr-1) is lower than that 

of mixed conifer forests in Mediterranean California. 

Th is diff erence may be due to the high N fertility 

and low P availability of the Andisols common in 

the Mexico City air basin. Because N is not strongly 

limiting in the Mexico City air basin, presumably, 

less added N is needed to induce N losses from the 

ecosystem. Th e critical load for NO3

-

 leaching in the 

Mexico City air basin is uncertain because it is based 

on only 1 year of deposition data, but assuming the 

deposition data approximate longer term inputs, we 

conclude that the Mexican forests are more prone to 

NO3

-

 leaching than the forests of California (Fenn et al. 

2008). In the Mexico City air basin with N deposition 

of 18.5 kg N ha-1 yr-1, NO3

-

 concentrations in stream 

and springwater were commonly at high levels (20 to 

90 μeq L-1), even during the growing season (Fenn et 

al. 2002b). By comparison, a throughfall deposition of 

17 kg N ha-1 yr-1 was determined as the critical load for 

incipient increases in NO3

-

 leaching (peak values above 

14 μeq L-1) in California mixed conifer forests during 

the winter dormant wet season (Fenn et al. 2008). 

Table 15.1—Empirical critical loads of nutrient N for Temperate Sierras ecoregion forests in the Mexico City air 

basina

. Reliability rating: ## reliable; # fairly reliable; (#) expert judgment

Site Critical Load for 

N depositionb

kg N ha-1 yr-1

Reliability Response Comments Study

New Mexico and 

Arizona forests

4-7 (#) Epiphytic lichens Assumes 60% eutroph 

threshold and 30-180 cm 

precipitation

Based on 

application of 

Geiser et al. 2010 

model

Las Cruces and 

Chichinautzin 

Ranges S/SW of 

Mexico City

15 # Elevated NO3

-

 in 

stream and spring 

waters

Data are from Mexican 

mountain pine sites in the 

Desierto de los Leones 

National Park and Ajusco

Fenn et al. 1999, 

2002b

a A variety of studies in the Mexico City air basin have demonstrated that air pollution is an important factor contributing to major 

changes in lichen communities, to the decline of sacred fi r, and possible reductions in plant growth, root production and mycorrhizal 

symbionts. However, because of insuffi cient data and the co-occurrence of multiple pollutants, the role of N deposition is unclear and 

critical loads for N deposition cannot be determined without further study.

b Nitrogen deposition was measured as throughfall (Fenn et al. 1999).
Chapter 15—Temperate Sierra GTR-NRS-80 179

Th e critical load for incipient NO3

-

 leaching in the 

Mexico City air basin cannot be determined without 

streamwater sampling and N deposition measurements 

from sites at intermediate levels of N deposition. 

Nitrogen deposition levels leading to NO3

-

 leaching in 

the Mexico City air basin appear to be within the range 

described from surveys of temperate forest catchments 

in Europe and North America (Dise and Wright 

1995, Gundersen et al. 2006, MacDonald et al. 2002, 

Stoddard et al. 2001).

15.6 Future Research Directions and 

Gaps in Knowledge

Future research could be directed at mountain forests 

outside of Phoenix, Arizona, to determine actual 

deposition fl uxes and ecosystem impacts at diff erent 

deposition levels. Follow up studies are needed in forests 

surrounding Mexico City to better understand nutrient 

cycling responses to N deposition of forests growing 

on N-rich but P-limited Andisols. Preliminary results 

indicate the value of developing critical load for N 

eff ects on mycorrhizal fungi, root production, and lichen 

communities, and of further refi nement of the estimated 

critical load for NO3

-

 leaching. Emissions of nitrogenous 

trace gases from soil in response to N deposition should 

also be investigated, particularly considering the high 

N concentrations in soil and the role of these gases in 

smog formation and climate change. Th e possible role 

of N deposition in combination with S deposition and 

elevated O3 exposure merits further investigation as 

contributing or causal factors in the severe decline of 

sacred fi r (Abies religiosa ) in the more polluted regions 

of the Desierto de los Leones National Park (AlvaradoRosales and Hernández-Tejeda 2002, Fenn et al. 

2002a).

Forest Inventory and Analysis lichen community surveys 

have been completed for the parts of this ecoregion 

in Arizona and New Mexico. Analysis of these lichen 

data along depositional N gradients estimated by 

NADP and the Environmental Protection Agency’s 

Models-3/Community Multiscale Air Quality (CMAQ) 

simulation model (Tonnesen et al. 2007) could refi ne 

the lichen-based critical load estimate for this ecoregion. 

More research is also needed to better understand 

the separate eff ects of acidifying and oxidizing versus 

fertilizing N-containing/N-derived pollutants on 

vegetation and lichens.

LITERATURE CITED

Alvarado-Rosales, D.; Hernández-Tejeda, T. 2002. 

Decline of sacred fi r in the Desierto de los Leones 

National Park. In: Fenn, M.E.; de Bauer, L.I.; 

Hernández-Tejeda, T., eds. Urban air pollution and 

forests: Resources at risk in the Mexico City air basin. 

Ecological Studies Series, volume 156. New York, 

NY: Springer-Verlag: 243-260.

CEC (Commission for Environmental Cooperation). 

1997. Ecological regions of North America. 

Toward a common perspective. Montreal, Canada: 

Commission for Environmental Cooperation. 

71 p. Available at http://www.cec.org/fi les/pdf/

BIODIVERSITY/eco-eng_EN.pdf (Accessed May 

24, 2010).

Dise, N.B.; Wright, R.F. 1995. Nitrogen leaching 

from European forests in relation to nitrogen 

deposition. Forest Ecology and Management. 71: 

153-161.

Fenn, M.E.; de Bauer, L.I.; Hernández-Tejeda, T. 

2002a. Summary of air pollution impacts on 

forests in the Mexico City air basin. In: Fenn, M.E.; 

de Bauer, L.I.; Hernández-Tejeda, T., eds. Urban air 

pollution and forests: Resources at risk in the Mexico 

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New York, NY: Springer-Verlag: 283-297.
Chapter 16—Tropical and Subtropical Humid Forests GTR-NRS-80 181

16 TROPICAL AND SUBTROPICAL HUMID FORESTS

S.J. Hall

16.1.1 Subtropical Humid Forests of Hawaii: 

Ecosystem Description

Subtropical humid forests in Hawaii are diverse, varying 

with rainfall and elevation, time since disturbance, 

degree of invasion, topography, and soil age. Native 

forest communities include the common dominant 

trees ohi’a lehua (Metrosideros polymorpha) and koa 

(Acacia koa), tree ferns, other common wet forest 

tree species (e.g., ʻōlapa [Cheirodendron spp.]) and 

Hawaiian holly (Ilex anomala), and a diverse understory 

of epiphytes and shrubs. Rates of endemism in the 

Hawaiian Islands are among the highest in the world, 

in part due to their isolated geography. Notable among 

these communities is a diversity of endemic birds and 

lobelioids, which are a special group of fl owering plants 

in the Campanulaceae (bellfl ower) family. Few, purely 

native Hawaiian wet forest communities exist at low 

elevations but instead consist of a mix of native and 

exotic species (Zimmerman et al. 2008). However, at 

higher elevations, tracts of intact humid forest remain in 

protected areas on most islands (Ricketts et al. 1999).

16.1.2 Subtropical Humid Forests of Puerto Rico: 

Ecosystem Description

Subtropical humid forests of Puerto Rico are also diverse 

biologically and geographically, occurring primarily as 

secondary growth in former agricultural or pasture lands 

or as primary or successional growth within remnant, 

protected forest reserves. Th ese ecosystems are composed 

of various plant communities that are associated with 

gradients in tidal inundation, rainfall, elevation, soil 

characteristics determined by geology, and time since 

disturbance (e.g., human, landslides, hurricanes) (Ewel 

and Whitmore 1973). Subtropical wet forests on 

volcanic soils include tabonuco forest at lower elevations 

in various stages of succession, dominated by abundant 

epiphytes and the tabonuco tree (Dacryodes excelsa) 

among others. Colorado and dwarf cloud forests grow 

at higher elevations, associated with trees such as palo 

colorado (Cyrilla racemifl ora) and nemoca cimarrona 

(Ocotea spathulata), respectively. Forests on serpentine 

16.1 Ecoregion Description

Tropical humid forests of the United States are located 

below 1000 m in elevation and experience average yearround temperatures between 20 °C to 26 °C, receive 

more than 1500 mm of precipitation annually, and 

experience fewer than three dry months per year (CEC 

1997; Fig. 2.2). In addition to tropical humid forests, 

we include in this chapter lower montane and low 

elevation tropical and subtropical humid forests with 

precipitation >1000 mm and mean annual temperatures 

>16 to 20 °C. Regions within the United States that 

contain forests of this type include coastal, low elevation, 

and lower montane regions: (1) on the windward slopes 

of the Hawaiian Islands, and (2) across Puerto Rico. 

Additionally, included in this chapter are (3) forests of 

the coastal plains of southern Florida within the greater 

Everglades wetland ecosystem, although mangrove and 

other saturated or seasonally saturated ecosystems are 

biogeochemically quite distinct and thus may respond 

diff erently than other, upland tropical forests discussed 

in this chapter. Tropical and subtropical humid forests 

occur in warm climates with plentiful rainfall and are 

thus well suited for agriculture. Consequently, many of 

the remaining forests of this type in the United States 

are in various stages of recovery following agricultural 

abandonment and are composed of a mix of native and 

exotic species. Native or primary forests in this class are 

highly fragmented and threatened by anthropogenic 

changes in land use and hydrology, and by invasive 

species. However, protected areas that contain forest 

communities less impacted by human activity compared 

to surrounding areas occur within each of the three 

regions. Some of the largest protected areas of tropical 

and subtropical forests in the United States include 

Hawai’i Volcanoes and Haleakala National Parks and 

a suite of forest reserves in the Hawaiian Islands; El 

Yunque National Forest in Puerto Rico; mangroves, 

tropical hammocks, and other tree islands within the 

Everglades and Biscayne National Parks; and protected 

pineland preserves in Dade and Monroe Counties, 

southern Florida.
182 Chapter 16—Tropical and Subtropical Humid Forests GTR-NRS-80

or limestone-derived soils are more xerophytic, 

supporting sclerophyllous, evergreen vegetation 

(Ewel and Whitmore 1973). Subtropical forests on 

the humid northern and southern coasts of Puerto 

Rico contain mixed stands of red, black, and white 

mangroves (Rhizophora mangle, Avicennia germinans, and 

Laguncularia racemosa) (Tomlinson 1986).

16.1.3 Tropical Forests of Southern Florida: 

Ecosystem Description

Tropical forests in southern Florida include pine 

rocklands and hardwood hammocks in the uplands, 

various ‘tree islands’ that are surrounded by occasional 

or persistent water, and coastal mangrove communities. 

Florida pine rocklands and tropical hardwood hammocks 

are located on shallow soils of limestone outcroppings 

in higher elevations of the Everglades (Lodge 2004). 

Pineland ecosystems are fi re-maintained, dominated 

by Florida slash pine (Pinus elliottii) and saw palmetto 

(Serenoa repens). Tropical hardwood hammocks develop 

in the absence of fi re, and are composed of numerous 

species characteristic of the Caribbean such as gumbo 

limbo (Bursera simarouba) and mahogany (Swietenia 

mahogani). Pinelands and hardwood hammocks are 

highly threatened by urbanization and land-use change, 

composing 2 percent of their former range (Ricketts et 

al. 1999). Mangroves are extensive along the southern 

Florida coastline, providing habitat for a variety of 

marine, estuarine, and terrestrial species of mammals, 

birds, reptiles, invertebrates, and fi sh. Th ese forests are 

dominated by red, black, and white mangroves, and 

buttonwood (Conocarpus erectus). Various other tree 

islands exist in the Everglades ecosystem, maintained by 

seasonal hydroperiods and serving as biogeochemical and 

fl oral hotspots on the fl ooded landscape (Sklar and van 

der Walk 2002). Common species within tree islands in 

the Everglades include the strangler fi g (Ficus aurea) and 

coco plum (Chrysobalanus icaco), among others (Wetzel 

et al. 2008).

16.2 Ecosystem Responses to N 

Deposition

Few studies have explored the impacts of nitrogen (N) 

enrichment on ecological processes in tropical and 

subtropical humid forests worldwide, and none have 

explicitly assessed the consequences of atmospheric N 

deposition in the three tropical ecoregions of the United 

States. However, exceptional, long-term N addition 

experiments within native ecosystems of Florida and 

Hawaii have contributed greatly to our understanding 

of tropical forest functioning (Table 16.1; Feller et 

al. 2007, 2009; Vitousek 2004). Existing studies used 

high rates of fertilizer addition primarily to assess 

nutrient limitation of vegetation and thus are not 

useful in identifying thresholds of response to N inputs. 

Nonetheless, these fertilization experiments and others 

do suggest that plant growth and ecosystem processes 

in some humid tropical and subtropical forests may be 

susceptible to N enrichment while others may be less 

susceptible, depending on numerous factors that alter 

forest nutrient status.

In the following sections we summarize the types and 

range of ecosystem responses to N fertilization observed 

in native U.S. tropical and subtropical forests. However, 

because the consequences of N deposition have not been 

well studied in these regions, evidence is also included 

from fertilization studies in other tropical forests that 

may operate similarly to those in Hawaii, Puerto Rico, 

and southern Florida. Data from non-U.S. forests are 

used only as supporting evidence; thus, information in 

this chapter is not intended to be a thorough review 

of ecosystem responses to N deposition or fertilization 

in tropical/subtropical forests worldwide. Information 

from non-U.S. regions is noted where included.

Nutrient limitation of primary production cannot 

be assessed using indirect measures of ecosystem N 

availability or cycling (Tanner et al. 1998). However, 

without abundant data, we broadly characterize tropical 

forests as relatively ‘N-rich’ or ‘N-poor’ and summarize 

the responses observed in these systems following N 

fertilization or those expected following N deposition.

In N-rich tropical and subtropical forests where primary 

production or nutrient cycling may be limited by factors 

other than N, N enrichment is not expected to change 

plant growth or species composition, although some 

evidence suggests that tissue chemistry, decomposition, 

and herbivory may be altered in some ecosystems 

(Benner and Vitousek 2007, Feller et al. 2007, Ostertag 
Chapter 16—Tropical and Subtropical Humid Forests GTR-NRS-80 183

and Hobbie 1999, Ostertag and Verville 2002). On 

the other hand, N addition to N-rich tropical forests 

may immediately increase N losses as gases (NOx

, 

N2O, or N2) or in drainage, but this response depends 

in large part on soil properties, redox conditions, and 

hydrological pathways (Hall and Matson 2003, Herbert 

and Fownes 1995, Lohse and Matson 2005, Templer

et al. 2008). Rivers that drain N deposition-aff ected 

tropical forests may experience increases in nitrate 

(NO3

-

) relative to dissolved organic N, which is the form 

of N that typically dominates export of undisturbed 

tropical catchments (Downing et al. 1999, Lewis et al. 

1999; reviewing tropical forests in the Americas and 

worldwide).

In tropical and subtropical forests where N availability 

is relatively low, ecosystem responses to atmospheric 

N deposition are expected to be similar to temperate 

forests and include increased plant growth, elevated 

rates of soil N cycling, increased foliar N concentration 

Table 16.1—Responses to long-term N-fertilization of native tropical and subtropical humid forests of 

the United States and related ecosystems.a

Site Location Hawaii,

Lower montane forests

Florida and Belize,

Coastal mangrove forests

Forest Type N-limited,

M. polymorpha

(ohi’a lehua)

P-limited,

M. polymorpha

(ohi’a lehua)

N-limited,

Florida 

P-limited,

Belize

Type of N Deposition Wet, dry, fog Wet Wet Wet + dry

N Deposition kg ha-1 yr-1 17 < 1 4 2

N Fert. Rate kg ha-1 yr-1 100 100 >100 >100

Form of N Fertilizer Urea/NH4NO3 Urea/NH4NO3 Urea Urea

Start year 1985 1991 1997 1997

Duration (years) 11-13 5-7 2-4 2-4 --Response relative to Control b--

Tree growth + 0 + 0

Foliar %N +/0 + + 0

Herbivory ND ND + -

Plant community 

composition

+ 0 ND ND

Epiphyte abundance ND 0 ND ND

Decomposition 0 + ND ND

Soil C:N 0 0 ND ND

N mineralization + + ND ND

Nitrifi cation + + ND ND

Microbial biomass N 0 0 ND ND

N-oxide emissions + + ND ND

NO3

-

 leaching + + ND ND

Soil pH 0 0 0 0

References 1-7 1-8 9-11 11,12

a Studies from Belize included to illustrate differences in ecosystem response to N fertilization in N-limited (Florida) vs. 

P-limited mangrove ecosystems.

b Responses are indicated as increase (+), decrease (–), or no response (0) in fertilized relative to control plots, or no data 

(ND) in indicated references.

References: 1) Vitousek et al. 1993; 2) Harrington et al. 2001; 3) Ostertag and Hobbie 1999; 4) Lohse and Matson 2005; 

5) Hall and Matson 2003; 6) Ostertag and Verville 2002; 7) Cordell et al. 2001; 8) Benner et al. 2007; 9) Caccia and Boyer 

2007; 10) Feller et al. 2003a; 11) Feller et al. 2007; 12) Howarth et al. 1996.
184 Chapter 16—Tropical and Subtropical Humid Forests GTR-NRS-80

and herbivory of some plant species, possible changes 

in community composition, and eventually elevated 

N losses as NO3

-

 or as gases as ecosystems become N 

saturated (Erickson et al. 2001, Feller et al. 2007, Hall 

and Matson 2003, Lohse and Matson 2005, Ostertag 

and Verville 2002, Vitousek and Farrington 1997). 

However, unlike in some temperate forests, atmospheric 

N deposition may be retained fi rst in the canopy rather 

than the forest fl oor (Wanek et al. 2007), for example, 

stimulating nutrient cycling and growth within epiphyte 

communities that intercept cloud and rainwater (Clark 

et al. 2005, Reich et al. 2003; both in Costa Rica).

Alkaline soils in Puerto Rico and southern Florida can 

be expected to have high acid neutralizing capacity 

that may buff er acidity derived from N deposition. In 

contrast, highly weathered, poorly buff ered volcanic soils 

in Puerto Rico and Hawaii or peat soils in Florida may 

be more susceptible to the negative eff ects of acidity, 

including elevated aluminum mobility and phosphorus 

(P) fi xation in soils that can further limit primary 

production (Asner et al. 2001; Matson et al. 1999, 

2002; all review tropical forests worldwide). Alkaline sea 

salt aerosols and dust likely enter coastal tropical forests 

but it is unclear whether these inputs will mitigate 

the impacts of acidity generated from deposition and 

microbial N cycling in soils (Evans et al. 2004 for 

Florida; Vogt et al. 2007 for a review of tropical forests 

worldwide).

16.3 Range of Responses Observed

Although small in their geographical extent within the 

United States, tropical and subtropical humid forests 

harbor an extraordinary diversity of soil types, species, 

plant assemblages, landscape characteristics, and 

landscape histories that will likely vary in their sensitivity 

to N enrichment (Nardoto et al. 2008 in Brazil; 

Townsend et al. 2008 for tropical forests worldwide). 

Furthermore, few studies have tested nutrient limitation 

of plant growth in tropical or subtropical forests or have 

explored soil and microbial responses to N additions 

in these regions, further compounding the uncertainty 

about how these forests will respond to N deposition.

16.3.1 N-Rich Tropical and Subtropical Forests

Worldwide, many tropical and subtropical forests are 

rich in N relative to other limiting factors, such as 

P, and are thus expected to respond to N deposition 

diff erently than N-limited forests in the temperate 

zone. For example, some tropical forests grow on 

geomorphically stable, highly weathered soils that 

contain few remaining primary minerals, leading to 

low available pools of rock-derived nutrients such as 

potassium (K), calcium (Ca), magnesium (Mg), or P 

relative to N (Chadwick et al. 1999, Walker and Syers 

1976). Other tropical forests, such as those in south 

Florida, are low in available P relative to N due in 

part to their growth on geologic substrates that lack 

P-bearing minerals or are subject to abiotic conditions 

that restrict P solubility (Noe et al. 2001, Troxler Gann 

et al. 2005). Additionally, many tropical forests support 

an abundance of legumes and other N-fi xing plant 

species that can lead to large available N pools in soil 

(Vitousek et al. 2002), including both mature forests 

and successional plant communities that are recovering 

from agricultural abandonment (Erickson et al. 2001). 

Table 16.2—Empirical critical loads of nutrient N for the Tropical and Subtropical Humid Forests ecoregion. 

Reliability rating: (#) – expert judgment; no data are available for this ecosystem category. The critical load is 

based on expert judgment and knowledge of ecosystems which may function similarly.

Ecosystem 

component

Critical load for 

N deposition 

kg N ha-1 yr-1

Reliability Response Comments Study

N-poor tropical 

and subtropical 

forests

5-10 (#) Changes in community 

composition; NO3

-

 leaching, 

N trace gas emissions

Critical load for N-poor forests based 

on estimates for Southeastern 

Coastal Plain forests

ND

N-rich tropical 

and subtropical 

forests

<5-10 (#) NO3

-

 leaching, N trace gas 

emissions

Critical load for N-rich forests should 

be lower than for N-poor forests 

based on possibility of N losses

ND
Chapter 16—Tropical and Subtropical Humid Forests GTR-NRS-80 185

In these cases, plant growth can be limited by elements 

or factors other than N, which may limit the ability 

of these forests to retain atmospheric N deposition 

within plant reservoirs. Because N cycles in relative 

excess compared to other nutrients in these systems, 

these humid tropical forests exhibit naturally high rates 

of NO3

-

 leaching and N trace gas emissions compared 

to temperate forests (Davidson et al. 2007, Lewis et 

al. 1999, Matson and Vitousek 1990, Martinelli et al. 

1999; all in tropical forests worldwide). Th us, it has 

been hypothesized that N-rich, P-limited wet tropical 

forests are ‘naturally N saturated’ and will have little 

ability to retain additional N in biotic pools, possibly 

leading to elevated N losses to aquatic systems and the 

atmosphere (Hall and Matson 1999).

Few studies have tested this hypothesis at low levels of 

N inputs that simulate expected rates of atmospheric 

deposition in humid tropical forests of the United States 

and abroad. In a dose-response experiment in lower 

montane forests of Hawaii, Hall and Matson (1999) 

found immediate and elevated nitrous oxide (N2

O) 

and nitric oxide (NO) emissions from soils after fi rsttime application of 15 kg N ha-1 as ammonium nitrate 

(NH4NO3) in a P-limited forest on the island of Kaua’i 

that received <1 kg N ha-1 yr-1 in ambient wet deposition 

(Carillo et al. 2002, Hall and Matson 1999). However, 

N-oxide emissions from this forest were not elevated 

following application of 5 kg N ha-1. Similar patterns 

were found after a dose-response experiment in an 

N-rich, P-poor lowland tropical rain forest in southeast 

Asia that experienced higher rates of soil N-oxide 

emissions compared to unfertilized controls after 

application of 15 kg N ha-1 (Hall et al. 2004).

If excess N is not eff ectively retained in biotic reservoirs, 

N-rich tropical forests might also be expected to respond 

to N deposition with elevated NO3

-

 losses. However, 

acidic, highly weathered soils often contain variably 

charged, subsurface, clayey horizons that can impede 

water infi ltration and retain anions such as NO3

-

 on 

clay particles. For example, in the same Hawaiian lower 

montane forests described earlier, Lohse and Matson 

(2005) found that NO3

-

 leaching was restricted by clay 

layers in the P-limited forest after both fi rst-time and 

long-term N fertilization with 50 kg N ha-1, despite lack 

of retention by biota, whereas NO3

-

 was readily leached 

after similar rates of fertilization in an N-limited forest 

that was well drained (Lohse and Matson 2005). Other 

factors associated with humid tropical forests may also 

favor NO3

-

 retention, even when N cycles in relative 

excess. For example, forests in Puerto Rico support high 

rates of dissimilatory nitrate reduction to ammonia 

(NH3) (DNRA), an anaerobic microbial process that 

consumes NO3

-

 anions (Silver et al. 2001). DNRA was 

once thought to occur primarily in aquatic sediments 

but now appears to be a NO3

-

 retention pathway favored 

by variable redox conditions characteristic of upland 

tropical forest soils (Pett-Ridge et al. 2006, Templer et 

al. 2008).

In N-rich humid tropical forests, it is expected 

that atmospheric N deposition will have relatively 

small direct eff ects on plant growth or community 

composition. For example, long-term N fertilization 

with 100 kg N ha-1 yr-1 had no eff ect on epiphytic 

cyanolichen abundance, diversity, or N-fi xation capacity 

in P-limited and N+P co-limited lower montane rain 

forests in Hawaii (ambient N deposition < 1 kg N ha-1 

yr-1 and ≤ 17 kg N ha-1 yr-1, respectively) (Benner et al. 

2007, Benner and Vitousek 2007, Carillo et al. 2002). 

Similarly, in an N-rich Hawaiian lower montane forest, 

N additions had no eff ect on plant growth or species 

richness and composition (Harrington et al. 2001, 

Ostertag and Verville 2002). Some responses to N 

deposition in upland tropical forests hold for mangrove 

ecosystems as well: fertilization with >100 kg N ha-1 

yr-1 in a P-limited, dwarf black mangrove ecosystem in 

Belize had no impact on plant growth or morphology 

(Feller et al. 2007). Although N deposition may not 

alter plant growth in forests that are not limited by 

N, chronic N inputs to acidic, poorly buff ered humid 

tropical forest soils may aff ect soil properties such as pH, 

aluminum mobility, and P fi xation that may indirectly 

decrease tropical forest growth over the long term (Asner 

et al. 2001, Matson et al. 1999; both describing tropical 

forests worldwide). However, few studies have tested 

these hypotheses in the fi eld.

While plant growth or community composition in 

relatively N-rich tropical and subtropical forests may 

not be aff ected by N deposition, tissue chemistry of 
186 Chapter 16—Tropical and Subtropical Humid Forests GTR-NRS-80

some plant species may be more sensitive to changes 

in N supply. For example, N additions increased foliar 

N concentrations and the decomposability of leaf and 

fi ne roots in P-limited Hawaiian lower montane forests 

(Ostertag and Hobbie 1999), and they increased litter N 

concentrations in lowland Panamanian forests that were 

not limited by N or P (Kaspari et al. 2008). However, 

based on the few studies that have measured these 

responses, this eff ect is not consistent across ecosystem 

types. For instance, N additions to a P-limited red 

mangrove stand in Belize and an N+P-limited red 

mangrove stand in Panama had no eff ect on foliar N 

or N:P ratios, or the amount of leaf area damaged by 

herbivores (Feller and Chamberlain 2007, Lovelock et 

al. 2004).

16.3.2 N-Poor Tropical and Subtropical Forests

While many humid tropical forests are relatively N 

rich, numerous processes can replenish rock-derived 

nutrients or increase N losses relative to other elements, 

thus promoting N limitation or conservative N cycling 

(LeBauer and Treseder 2008). Th ese processes include 

deposition of transported dust or sea salt aerosols, 

erosion, fi re, and other disturbances such as forest 

clearing that increase N losses or N sinks in vegetation 

(Chadwick et al. 1999, Davidson et al. 2007, Porder et 

al. 2005, Scatena and Lugo 1995; in Amazon forests). 

For example, much of the forested land area in Puerto 

Rico is successional, as >75 percent of the landscape 

is aff ected by hurricanes or landslides each century 

(Lugo and Scatena 1996, Walker et al. 1996). Th ese 

disturbances can cause N losses to the atmosphere and 

water that may remain elevated for years until biological 

sinks are re-established (Erickson and Ayala 2004, 

McDowell et al. 1996, Schaefer et al. 2000). Combined, 

these spatially and temporally heterogeneous factors 

will likely cause variation in the extent of N retention 

in humid tropical and subtropical forests in response to 

atmospheric N deposition.

Measures of nutrient availability are often correlated 

with the responsiveness of plant growth to fertilization 

but are not suffi cient to establish nutrient limitation 

(Tanner et al. 1998). However, evidence in the literature 

suggests that many tropical forests in the United States 

with low N availability will respond to N deposition in 

ways similar to temperate forests, with increased rates 

of primary production and foliar N concentrations, 

and potentially increased soil N cycling and N gas 

emissions over time. For example, emissions of N2

O 

and NO were elevated after 6 years of complete nutrient 

additions (300 kg N ha-1 yr-1 and 100 kg P ha-1 yr-1 plus 

cations; ambient deposition of 4 to 9 kg N ha-1 yr-1) in 

a >60-yr-old tropical forest in Puerto Rico that cycled 

N conservatively (Erickson et al. 2001, Ortiz-Zayas 

et al. 2006). In a strongly N-limited, lower montane 

forest on the island of Hawai’i, NOx

 emissions were 

elevated, but relatively low after application of ≥25 to 

50 kg N ha-1, but long-term N additions of 100 kg N 

ha-1 yr-1 increased rates of net N mineralization and 

nitrifi cation in soils (ambient deposition of 17 kg N 

ha-1 yr-1, primarily as fog from active volcanic sources) 

(Carillo et al. 2002, Hall and Matson 2003). Long-term 

N additions to this N-limited Hawaiian lower montane 

forest also altered species composition and decreased 

species richness, moss cover, and native seedling 

abundance (Ostertag and Verville 2002). However, 

decomposition rates were not aff ected by N additions, 

suggesting that carbon (C) mineralization rates were 

limited more strongly by C quality than nutrient 

availability (Hobbie 2000).

Growth of mangrove forests in Florida is limited by N 

in some regions and P in others, varying across small 

scales within mangrove stands (Feller et al. 2003b, 

2007) and across landscape gradients of tidal inundation 

that restrict availability of nutrients under saline 

conditions (Boyer 2006, McKee 1995). In an N-limited 

black mangrove ecosystem in Florida (ambient 

deposition ~ 4.1 kg N ha-1 yr-1) (Caccia and Boyer 

2007), N additions of >100 kg N ha-1 yr-1 increased 

growth (wood allocation and shoot elongation), altered 

foliar chemistry (increased N and P resorption effi ciency 

and foliar N concentrations; decreased C:N ratios of 

green and senescent tissue), and increased herbivory 

by some species (Feller et al. 2007). Plants are known 

to decrease N resorption from senescing tissue with 

increasing N availability, although the strength of this 

process varies widely across species (Norris and Reich 

2009). In N-limited lagoons in Florida, increased N 
Chapter 16—Tropical and Subtropical Humid Forests GTR-NRS-80 187

resorption in N fertilized mangrove stands may have 

resulted from simultaneous changes in P dynamics 

(Feller et al. 2007). In contrast, N fertilization had no 

eff ect on leaf area ratio, specifi c leaf area, photosynthetic 

rates, or hydraulic traits of N-limited black mangrove 

stands (Lovelock et al. 2006). Other work shows that 

the impact of N enrichment may also depend on species 

composition. For example, Feller et al. (2003b) and 

Lovelock et al. (2004) found that N fertilization had 

no impact on N resorption effi ciencies in N+P-limited 

red mangroves in Panama or N-limited (seaward 

fringe), N+P-limited (transition zone), or P-limited 

(forest interior) red mangroves in Belize. Also, unlike 

in N-limited black mangroves, N fertilization in N+P 

limited red mangroves in Panama had no eff ect on the 

fraction of leaf area damaged by herbivores (Lovelock et 

al. 2004).

Although N may be retained in some N-poor tropical 

forests as in temperate systems, the mechanisms of N 

retention may diff er depending on the structure of 

tropical plant communities. For example, high yearround leaf area of tropical forests may act to retain 

N in canopies, intercepting N deposition before it 

reaches microbial and plant communities of the soil and 

forest fl oor (Bakwin et al. 1990, Sparks et al. 2001). 

In a study of cloud forests in Costa Rica, epiphytes 

within the canopy retained 50 percent of the N applied 

experimentally in the laboratory and fi eld, leading to 

higher foliar N concentrations in bryophyte species 

(Clark et al. 2005). In mangrove ecosystems of Belize, 

ammonium (NH4

+

) in bulk precipitation was retained in 

the canopy in an N-limited stand, but not in a P-limited 

stand, decreasing throughfall fl uxes of N compared to 

rainwater (Wanek et al. 2007). Th ese data suggest that 

change in composition or chemistry of foliage or canopy 

epiphyte communities may indicate initial stages of N 

saturation of tropical forest ecosystems.

Finally, recent work suggests that after disturbance in 

some tropical forests, N limitation may be short-lived. 

For example, tropical forests in the Amazon supported 

low rates of N cycling early in succession following 

agricultural abandonment, but returned to ‘leaky’ N 

cycling similar to N-rich, mature tropical forests within 

decades (Davidson et al. 2007).

16.4 Critical Loads Estimates

Estimates of critical loads of N in the U.S. Tropical and 

Subtropical Humid Forests ecoregion are uncertain due 

to lack of data across N deposition gradients or from 

low-level N addition studies. Furthermore, critical loads 

will depend on the form of N inputs and numerous 

ecosystem characteristics that infl uence N retention, 

most of which vary considerably across tropical and 

subtropical forests. Th ese factors include, among others, 

the form of N deposition (NH4

+

 vs. NO3

-

), the extent 

of N limitation of primary production; spatial and 

temporal heterogeneity in soil organic matter, waterholding capacity, and moisture that can infl uence 

microbial N cycling processes (e.g., nitrifi cation and 

dentrifi cation); subsurface soil properties that may 

alter N leaching (e.g., retention or loss of NH4

+

 or 

NO3

-

 from deposition will depend on soil texture and 

cation vs. anion exchange capacity); and composition of 

heterogeneous and diverse plant communities. However, 

in the absence of data, we estimate critical loads of 5 to 

10 kg N ha-1 yr-1 for N-poor tropical and subtropical 

forests (e.g., some mangrove ecosystems in Florida, early 

successional forests or forests growing on pedogenically 

young substrates), within the same range as forests in 

the Southeastern Coastal Plain that may have similarly 

weathered soil types and experience warm year-round 

temperatures. Critical loads for N-rich tropical and 

subtropical forests may be lower than this level if N 

deposition rapidly stimulates pathways for N loss. Th us, 

we estimate critical loads of <5 to 10 kg N ha-1 yr-1 for 

N-rich tropical and subtropical forests.

16.5 Comparison to Critical Loads for 

Other Regions

Few estimates exist for N critical loads in tropical and 

subtropical forests. Studies in south China recommend 

high critical loads of 50 to >150 kg N ha-1 yr-1 to 

prevent departure from current conditions based on 

fertilization experiments across a range of ecosystem 

types (reviewed in Duan 2009). However, N deposition 

to Chinese subtropical forests ranges from at least 8 to 

50 kg N ha-1 yr-1 (Chen and Mulder 2007a, Chen and 

Mulder 2007b). Th us, it is impossible to determine 

whether current conditions represent a change from 

conditions prior to elevated N deposition. Furthermore, 
188 Chapter 16—Tropical and Subtropical Humid Forests GTR-NRS-80

species composition within Chinese subtropical forests 

varies considerably from tropical and subtropical forests 

in the United States. As a result, it is unclear whether 

critical load estimates developed for China are applicable 

to forests reviewed in this chapter.

16.6 Future Research, Gaps in Data, 

Gaps in Knowledge

Currently, tropical and subtropical forests in the 

United States are highly fragmented and are threatened 

primarily by land-use change, invasive species, 

and nutrient pollution from agricultural runoff . 

Consequently, to date no studies have evaluated the 

ecological consequences of atmospheric N deposition in 

these regions. Refi ned estimates of N critical loads for 

this ecoregion will require studies that explore ecological 

processes across N deposition gradients or following 

low-level N additions.

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Chapter 17—Wetlands GTR-NRS-80 193

17 WETLANDS

T.L. Greaver, L. Liu, R. Bobbink

Fens, freshwater marshes, freshwater swamps, and 

riparian wetlands are characterized by ground and 

surface water inputs that are often on the same 

order of magnitude as precipitation (Koerselman 

1989). Fens generally receive some drainage from 

surrounding mineral soils and, in many ecological 

aspects, are transitional between bogs and freshwater 

marshes. Freshwater marshes are characterized by tall 

graminoids, while freshwater swamps are forested 

wetlands dominated by various species (e.g., bald 

cypress [Taxodium distichum], pond cypress [Taxodium 

ascendens], white-cedar [Chamaecyparis thyoides], or red 

maple, [Acer rubrum]). Riparian wetlands are linear 

systems along rivers and streams and formed by the 

fl ooding of these streams or rivers.

Intertidal wetlands (e.g., salt marshes and mangrove 

swamps) develop on sheltered coasts or in estuaries 

where they are inundated by water with a periodicity 

that refl ects the combination of tidal cycles and 

freshwater inputs (e.g., precipitation and ground water). 

Marine water input depends on the proximity to the 

coast; wetlands that occur nearest to the coast are the 

most saline and those towards the inland are primarily 

freshwater habitats (Archibold 1995, Mitsch and 

Gosselink 2000).

17.2 Ecosystem Responses to N 

Deposition

Known responses to N enrichment are generally 

derived from nutrient-addition studies in the fi eld and 

observations along gradients of N deposition. A variety 

of ecological endpoints are evaluated, such as altered soil 

biogeochemistry, increased peat accumulation, elevated 

primary production, changes in plant morphology, 

changes in plant population dynamics, and altered 

plant species composition (US EPA 2008). Morris 

(1991) suggested the eff ect of N deposition on wetland 

ecosystems depends on the fraction of rainfall (a proxy 

for atmospheric N deposition) in its total water budget; 

the sensitivity to N deposition was suggested as bogs > 

17.1 Ecosystem Description

Th e U.S. Clean Water Act defi nes wetlands as “those 

areas that are inundated or saturated by surface or ground 

water at a frequency and duration suffi cient to support, 

and that under normal circumstances do support, a 

prevalence of vegetation typically adapted for life in 

saturated soil conditions”. Frequent water saturation 

of the soil creates anaerobic conditions and results 

in slow decomposition of organic matter. Wetlands 

in the United States support more than 4200 native 

plant species, of which 121 are federally threatened 

or endangered (NRCS 2009). Wetlands also off er key 

ecosystem services, such as sequestering carbon (C), 

providing habitat, regulating fl ood, maintaining water 

quality, and stabilizing coastal slope (US EPA 1993). 

About 107.7 million acres of wetlands are widely 

distributed in the conterminous United States (Fig. 

17.1), 95 percent of which are freshwater wetlands and 5 

percent are estuarine or marine wetlands (FWS 2005).

Depending on geology, position in the watershed, and 

climate, very diff erent types of wetland ecosystems 

may develop. Wetlands have numerous classifi cations 

in North America. In general, marshes (dominated 

by herbaceous species) and swamps (dominated by 

woody species) occur on mineral soils or shallow peats. 

Peatlands, including ombrotrophic (raised) bogs, 

minerotrophic (true) fens, and transition (poor fens), 

are true peat-accumulating communities (Vitt 1994). 

Wetlands can be divided into three hydrology-based 

categories that are useful for discussing nitrogen (N) 

sources and cycling: bogs; fens, marshes, swamps, and 

riparian wetlands; and intertidal wetlands.

Ombrotrophic bogs are acidic, moss-dominated 

freshwater wetlands that receive water inputs primarily 

via precipitation. Th ey develop where precipitation 

exceeds evapotranspiration and where there is some 

impediment to drainage of surplus water (Mitsch and 

Gosselink 2000). Bogs are especially common in the 

cool boreal zones of North America.
194 Chapter 17—Wetlands GTR-NRS-80

Fig 17.1—Distribution of wetlands in the United States. Source: Dahl and Allord 1997.
Chapter 17—Wetlands GTR-NRS-80 195

fens, marshes, swamps > intertidal wetlands. At the most 

sensitive end of the spectrum, ombrotrophic bogs receive 

exogenous nutrients (from external sources) primarily 

from precipitation, and the species in these ecosystems 

are adapted to low levels of N (Bridgham et al. 1995, 

1996; Shaver and Melillo 1984). Intertidal wetlands are 

at the other end of this spectrum; in these ecosystems, 

N loading from marine/estuarine water sources exceed 

that from atmospheric inputs by one or two orders 

of magnitude (Morris 1991). Building on these early 

observations by Morris (1991), it is now generally 

accepted that nutrient budgets in wetland ecosystems 

are tightly linked with local hydrology (Bobbink et al. 

2003). In general, N deposition is expected to have a 

larger eff ect on ecosystems with more closed N cycles, 

such as bogs, than on those with open N cycles, such 

as intertidal wetlands. Most data for U.S. wetlands 

indicates nutrient-related eff ects, unlike Europe, where 

acidifi cation eff ects in wetlands are well documented.

17.3 Range of Responses Observed

Both freshwater and estuarine wetlands tend to be 

N-limited ecosystems (LeBauer and Tresseder 2008, 

US EPA 1993). Most studies evaluate short-term 

fertilization with high N loading that typically exceeds 

atmospheric deposition. Th erefore, it is diffi cult to assess 

how long-term N deposition will aff ect N limitation, 

especially considering that the degree of N limitation 

varies among wetlands across the United States (Bedford 

1999).

17.3.1 Freshwater Wetlands

Freshwater wetlands harbor large numbers of plant 

species that have evolved under N-limited conditions. 

Th ese include quillworts (Isoetes spp.), of which three 

species are federally endangered; insectivorous plants 

such as the endangered green pitcher (Sarracenia 

oreophila); and the genus Sphagnum, of which there 

are 15 species listed as endangered by some eastern 

states. An extensive literature survey on temperate 

North American wetlands found that, in general, plant 

communities under high nutrient supply had low species 

richness when compared among diff erent wetland types 

(Bedford et al. 1999).

Sphagnum moss. Th e genus Sphagnum dominates 

ombrotrophic bogs and some nutrient poor fens in 

the northern United States and Canada. Th e species 

composition of bogs in both countries is similar, 

therefore, the Canadian data are included in the 

following discussion. Th ese mosses effi ciently capture 

atmospheric N deposition with retention rates between 

50 and 90 percent; much of the variation is due to the 

depth of the water table (Aldous 2002a, 2002b). Studies 

conducted on four species of Sphagnum in Maine (2 

to 4 kg N ha-1 yr-1 wet deposition) and New York (10 

to 13 kg N ha-1 yr-1 wet deposition) document that 

higher wet deposition of N resulted in higher tissue N 

concentrations and greater net primary productivity 

(NPP), but lower bulk density (Aldous 2002a; Table 

17.1). Note that these results are for wet deposition and 

not total deposition of N.

A study of Sphagnum fuscum in six Canadian peatlands 

showed a weak, although signifi cant, negative 

correlation between NPP and modeled total N 

deposition when deposition levels were greater than 3 kg 

N ha-1 yr-1 (Vitt et al. 2003; Table 17.1). A study of 23 

ombrotrophic peatlands in Canada with total modeled 

N deposition ranging from 2.7 to 8.1 kg N ha-1 yr-1

showed that peat accumulation increases linearly with N 

deposition. However, in recent years this rate has begun 

to slow, indicating a limited capacity for N to stimulate 

accumulation (Moore et al. 2004; Table 17.1).

In summary, documented responses to N deposition 

by Sphagnum mosses in the United States and Canada 

are limited to changes in N tissue concentration, bulk 

density, and NPP. Th ere is additional information from 

European ecosystems that prolonged N deposition may 

ultimately lead to N saturation, leaching of excess N, 

and ultimately a shift in community composition. For 

example, N additions to a boreal bog in Europe resulted 

in a collapse of the Sphagnum mat followed by increased 

growth of ericaceous shrubs and the sedge, tussock 

cottongrass (Eriophorum vaginatum) (Wiedermann et 

al. 2007). In more southerly bogs in Europe, excess 

N permitted an invasion of vascular plants normally 

held at bay due to low N concentration because of the 

N-fi ltering capacity of mosses (Lamers et al. 2000).
196 Chapter 17—Wetlands GTR-NRS-80

Table 17.1—Response to N inputs of biological and chemical indicators for freshwater wetlandsSite N Input

kg N ha-1 yr-1

Indicator Responses Reference

Peatlands:23 ombrotrophic peatlands in Canada

2.7 to 8.1 Peat accumulation Peat accumulation increases linearly with N deposition: y = 2.84x +0.67, r2 =0.32, 

P<0.001, where x is wet atmospheric N deposition from 1990-1996 (g m-2 yr-1), y 

is 50-year peat N accumulation (g m-2 yr-1). However, in recent years this rate has 

begun to slow, indicating limited capacity for N to stimulate accumulation

Moore et al. 2004

Six Canadian peatlands

3 Sphagnum fuscum At sites with deposition levels ranging from 0.7 to 152 kg N ha-1 yr-1, a weak, 

although signifi cant, negative correlation between NPP and N deposition when deposition levels were greater than 3 kg N ha-1 yr-1: y = 150 - 3.4x, R2=0.01, 

P=0.04, where x is N deposition rate (kg N ha-1 yr-1), y is NPP (g m-2 yr-1).

Vitt et al. 2003

Poor fen in NW Ontario, Canada

4.6 Sphagnum 

magellanicum

Growth was stimulated by 4.6 kg N ha-1 yr-1 as NO3

ˉ

 -N. However, additional N 

loading (18 kg N ha-1 yr-1 as NO3

ˉ

 -N and 18 kg N ha-1 yr-1 as NH4

+-N) did not further 

stimulate growth

Rochefort et al. 1990

Two bogs, Adirondacks (NY)

10 to 13 wet 

deposition

Sphagnum spp.

Polytrichum commune

Comparison of low (2 to 4 kg N ha-1 yr-1) and high N deposition level (10 to 13 kg 

N ha-1 yr-1) showed elevated tissue N concentrations, greater NPP and lower bulk 

density under the higher level of deposition 

Aldous 2002a 

Lindsey Brook Bog, Vanceboro Bog (ME)

2 to 4 wet 

deposition

Two bogs, Adirondacks (NY)

10 to 13 wet 

deposition 

Sphagnum spp. Main fi ndings of the 15N tracer experiments are 1) the two Adirondack sites had a 

greater N% translocation of N to the growing capitula than the Maine sites; 2) 11% of the tracer was retained in Maine and 80% in the Adirondacks; 3) atmospheric deposition is a portion of the Sphagnum N budget, the remaining N likely comes from N mineralization

Aldous 2002b

Lindsey Brook Bog, Vanceboro Bog (ME)

2 to 4 wet 

deposition

Pitcher plant:Hawley Bog (MA)Molly Bog (VT)

6.8 Pitcher plant

(Sarracenia purpurea)

Based on empirical demographics, an increase above 6.8 kg N ha-1 yr-1 would 

cause substantial increase in the extinction risk

Gotelli and Ellison 2002

26 bogs in Massachusetts and Vermont, US

9.76 to 16.42 Pitcher plant;

[NH4] of soil pore water 

Relative keel size of S. purpurea increased from 0.4 to 0.7 with soil pore water NH4

concentration increase from 0.1 to 0.5 mg L-1

Ellison and Gotelli 2002

Hawley Bog (MA)Molly Bog (VT)

10 to 14 Pitcher plant When populations occurring under low deposition (0-1.4 kg N ha-1 yr-1) are 

compared to those at higher deposition (10 to 14 kg N ha-1 yr-1), negative 

population growth rate is observed under higher deposition

Gotelli and Ellison 2006

Molly Bog (VT) 35 Pitcher plant Relative keel size of S. purpurea was compared in two nitrogen addition treatments 

(0 and 35 kg N ha-1 yr-1). Keel size increased from 0.2 to 0.8 under the higher 

loading

Ellison and Gotelli 2002

Prairie wetland:Controlled mesocoms at the University of Minnesota Horticultural Research Center (MN)

0,120, and 480 11 species in sedge 

meadow community 

common to prairies

N addition as nitrate-N 1) increased community biomass; 2) decreased the root:shoot ratio from 0.82 at 0 addition to 0.48 at 120 kg N ha-1 yr-1, no further 

reduction was observed at the highest N addition level; 3) decreased community diversity and evenness between the control and highest N addition.

Green and Galatowitsch 2002
Chapter 17—Wetlands GTR-NRS-80 197

Purple pitcher plant. Purple pitcher plant (Sarracenia 

purpurea), also known as northern pitcher plant, is long 

lived (30 to 50 years) and widely distributed in bogs, 

fens, and swamps across Canada and the eastern United 

States (Fig. 17.2; Ellison and Gotelli 2002). Purple 

pitcher plant has adapted to nutrient poor environments 

and is very sensitive to increasing N input. Purple 

pitcher plant was listed as vulnerable in New York, 

threatened in Florida, Maryland, and Michigan, and 

endangered in Illinois (USDA 2009). In a study of 

purple pitcher plant in Vermont and Massachusetts, 

Ellison and Gotelli (2002) conducted a series of N 

addition experiments by augmenting N availability to 

leaves with solution of 0, 0.1, and 1 mg L-1 nitrogen as 

ammonium chloride (NH4

Cl) for one growing season. 

Population growth rates, estimated by demographic 

survey, were positive for 0 and 0.1 mg N L-1 additions 

(equal to atmospheric deposition of 0 to 1.4 kg N 

ha-1 yr-1)

23 and negative for 1 mg N L-1 additions 

(equivalent to 14 kg N ha-1 yr-1)

23 (Gotelli and Ellison 

2006; Table 17.1). Based on the annual demographic 

information, a nonstationary matrix model forecasted 

that the extinction risk within the next 100 years was 

small if N deposition rate was maintained at 4.5 to 6.8 

kg N ha-1 yr-1.

23 However, a small annual increase (1 to 

4.7 percent) in N deposition rates would substantially 

increase the extinction risk (Gotelli and Ellison 2002).

Increasing N availability not only reduced population 

growth of purple pitcher plant, but also dramatically 

altered plant morphology. Purple pitcher plant produces 

carnivorous leaves (pitcher) and photosynthesis effi cient 

Figure 17.2—Distribution of purple pitcher plant in North America. Green denotes areas where 

purple pitcher plant is known to occur. Sources: USDA 2009.

23N treatments were selected to represent annual N 

deposition measured at the nearest monitoring sites of 

National Atmospheric Deposition Program (NADP). 

Th e unit of N treatments reported in the publication was 

precipitation-weighted mean concentrations (mg N L-1), 

from which we calculated the level of deposition (kg N ha-1 

yr-1) using the equation: deposition= precipitation-weighted 

mean concentrations × annual precipitation. More detailed 

information on N deposition is available on the NADP 

website: http://nadp.sws.uiuc.edu/sites/ntnmap.asp
198 Chapter 17—Wetlands GTR-NRS-80

leaves (phyllodia). Nitrogen enrichment was shown 

to stimulate the photosynthesis rate and increase the 

production of phyllodia relative to pitcher (Ellison 

and Gotelli 2002). Th e fi eld N deposition simulation 

experiment (treatments ranged from 0 to 35 kg N ha-1

yr-1)

23 revealed a positive linear relationship between N 

deposition level and relative keel size (keel width/total 

width, an indication of the relative size of the pitcher 

tube, in which prey are digested). Th is correlation 

was supported by the fi eld surveys of 26 sites across 

Massachusetts and Vermont (Ellison and Gotelli 2002), 

and 39 sites across Canada and the eastern United States 

(Ellison et al. 2004). Th e relative keel size of northern 

pitcher plant increased with increasing ammonium 

(NH4

+

) concentration in soil water, and may be used as 

a bioindicator (log [NH4

+

] = -1.57 + 1.78x relative keel 

size).

Greenhouse Gas Flux. A recent meta-analysis evaluated 

the eff ects of N enrichment on biogenic greenhouse 

gas (GHG) fl ux from terrestrial ecosystems (Liu and 

Greaver 2009). In wetlands, N enrichment increased 

methane (CH4) emission and nitrous oxide (N2

O) 

emission, but had no eff ect on net ecosystem exchange 

of C or CH4

 uptake. Th ese results are the average 

response for N addition experiments on wetlands 

located around the world. Data is limited for the United 

States; there are just four fi eld addition experiments 

on freshwater wetlands. Of these, three studies 

demonstrated a response to N addition. Th ese are 

summarized in Table 17.2. In the fi rst study, N addition 

of 20 and 60 kg ha-1 yr-1 in combination with 6.7 and 

20 kg ha-1 yr-1 phosphorus (P) had no eff ect on carbon 

dioxide (CO2

) or CH4 emissions from a Minnesota 

peatland fen (Keller et al. 2005). Th ere is evidence that 

N addition increases N2

O fl ux from swamp forest in 

Louisiana and mangrove forest in Florida. However, 

the lowest N addition level was 100 kg ha-1 yr-1, making 

it diffi cult to determine if N deposition would also 

increase N2

O fl ux (Lindau et al. 1994, Delaune et al. 

1998, Whigham et al. 2009). A 5-year addition study 

indicated that net ecosystem exchange of CO2

 with the 

atmosphere of bogs in Ottawa, Ontario, was reduced by 

45 percent with an addition of 64 kg N ha-1 yr-1 (Bubier 

et al. 2007).

17.3.2 Intertidal Wetlands

Nitrogen loading to intertidal wetlands tends to be 

much higher than loading to freshwater wetlands 

because, in addition to atmospheric deposition, N 

sources may include land runoff , ground water, marine 

water, and waste-water effl uent. Nitrogen load to one 

salt marsh in Narragansett Bay, Rhode Island, was 

calculated to be 10,253 kg N ha-1 yr-1 (Wigand et 

al. 2003), an extremely high value. Compared to N 

brought in by tidal water and ground water (typically 

565 to 668 kg N ha-1 yr-1), the N deposited directly 

to the surface of coastal marshes (~ up to 30 kg N 

ha-1 yr-1) is a small proportion of the total N budget 

(Morris 1991). However, atmospheric deposition 

can be the primary N source in some regions. For 

example, in southern New England estuaries, where N 

loading (mean =167 kg N ha-1 yr-1) was much lower 

than other U.S. estuaries (mean =893 kg N ha-1 yr-1), 

direct atmospheric deposition to the estuary surface 

was the largest N source, accounting for 37 percent of 

the total N loading (Latimer and Charpentier 2010). 

Th is information suggests that the impact of direct 

atmospheric deposition of N may vary among regions. 

Indirect atmospheric deposition—N deposited to 

the watershed and transported via surface or ground 

water—could be another major source of the total N 

load to coastal marshes. For example, the modeled 

contribution from the atmosphere (36 million kg N yr-1) 

was about 21 to 30 percent of the total N loading of 

170 million kg N yr-1 (Castro et al. 2003, Driscoll et al. 

2003) for Chesapeake Bay waters, and 16 percent of the 

total N loading of 0.029 million kg N yr-1 for southern 

New England estuaries (Latimer and Charpentier 2010) 

Future studies are needed to determine the role of direct 

and indirect atmospheric N deposition on the nutrient 

budget of intertidal wetlands.

Regardless of its source, N enrichment is shown to 

alter the structure and function of intertidal wetland 

ecosystems (Table 17.3) by causing increased primary 

production (Darby and Turner 2008a, Mendelssohn 

1979, Tyler et al. 2007, Wigand et al. 2003); invasion 

of nonnative species (Tyler et al. 2007); altered 

competition between native species (Crain 2007, 

Mendelssohn 1979, Wigand et al. 2003); loss of 
Chapter 17—Wetlands GTR-NRS-80 199

Table 17.2—Greenhouse gas fl ux in response to N enrichment for freshwater wetlands.

Site N Input

kg N ha-1 yr-1

Indicator Responses Reference

Mer Bleue Bog, 

Ottawa, ON

64 +PK NEE 

(net ecosystem 

exchange of 

carbon)

After 5 years nutrient treatment, the rate of maximum 

NEE was signifi cantly reduced (45%) under the highest 

levels of nutrient addition (64 kg N ha-1 yr-1 +5.0 kg 

P ha-1 yr-1 + 63 kg K ha-1 yr-1 ), but was not changed 

under other nutrient levels, compared to control

Bubier et al. 

2007

Swamp forest,

St. James Parish, 

LA

100 N2O emission Addition of NO3 and NH4 to the columns increased N2O 

production 56% and 15% over the control, respectively

Delaune et al. 

1998

Swamp forest,

Spring Bayou 

Wildlife 

Management 

Area, LA

100, 300 NH4+

- N; 100, 300 

NO3- -N

N2O emission N2O emission was largely stimulated by NO3

-

addition 

but not NH4

+

;

Total N2O evolution was estimated to be 0.44,0.36,2.83 

and 10.45 kg ha-1 yr-1 for the 100 NH4

+

, 300 NH4+, 

100 NO3- and 300 kg NO3-N ha -1 yr-1 treatments, 

respectively

Lindau et al. 

1994

Mangroves forest, 

Indian River 

Lagoon, FL

100 N2O emission Averaged across the three experimental years (2001 

-2003), N2O emission was almost six times higher in 

fertilized plots (0.028 mg N2O-N m2

 hr-1), compared to 

that in control plots (0.005 mg N2O- N m2

 hr-1)

Whigham et al. 

2009

sensitive species (Latimer and Rego 2010); increased 

herbivore damage to plants (Bertness et al. 2008); 

stimulated evapotranspiration (Howes et al. 1986); 

altered microbial community and pore water chemistry 

(Caff rey et al. 2007); and altered carbon allocation 

between roots and shoots (Darby and Turner 2008b). 

It is unclear if these eff ects could occur in response to 

current levels of deposition alone because most studies 

apply N additions that are several orders of magnitude 

greater than atmospheric deposition.

Spartina. Spartina is an aggressive perennial cordgrass 

native to the coasts of the Atlantic Ocean. Th e dense 

root mats of Spartina trap sediment, elevating the 

substrate, and creating new shoreline. Nitrogen 

enrichment has been shown to alter the ecophysiological 

processes and community composition of Spartina salt 

marshes. Th is work is based on N addition levels that 

simulate total loading, not deposition alone. Evidence 

from Mendelssohn et al. (1979) indicates 280 kg N ha-1

yr -1 alters growth and biomass of Spartina. Howes et al. 

(1986) found N addition (416 kg N ha-1 yr -1) increased 

live aboveground biomass, leaf area coverage, and 

evapotranspiration of smooth cordgrass (S. alternifl ora).

Under N-enriched condition, saltmeadow cordgrass 

(S. patens) is often out-competed and replaced by 

smooth cordgrass (Emery et al. 2001). Wigand et al. 

(2003) examined plant community structure of 10 salt 

marshes over a 2-year period, with N ranging from 2 

to 10,253 kg N ha-1 yr -1, in Narragansett Bay. Several 

dose response curves based on their fi ndings are given 

in Figure 17.3. Species richness decreased linearly with 

increasing N loading. With N loading higher than 63 to 

400 kg N ha-1 yr -1, the density of saltmeadow cordgrass 

and short smooth cordgrass rapidly decreased and the 

density of tall smooth cordgrass increased.

Eelgrass. Eelgrass, genus Zostera, is submerged aquatic 

vegetation that provides important habitat for many 

wildlife species, such as fi sh, shellfi sh, sea turtles, and 

waterfowl. Eelgrass beds are also important for sediment 

deposition and substrate stabilization. Th e massive 

decline of eelgrass has been widely documented in 

estuaries across Atlantic and Pacifi c coast (Latimer and 

Rego 2010, Ward et al. 2003). Excess N loading is one 

of the primary disturbances causing the rapid loss of 

eelgrass habitat. Latimer and Rego (2010) estimated 

the relationship between eelgrass extent and predicted 

watershed-derived N loading for 62 shallow estuaries 

along the coasts of Connecticut, Rhode Island, and 

Massachusetts. Th ey found that N loading greater than 

50 kg N ha-1 yr-1 lead to a signifi cant decline of eelgrass 

coverage, and almost all eelgrass habitats were lost at 

loading levels higher than 100 kg N ha-1 yr-1.
200 Chapter 17—Wetlands GTR-NRS-80 Table 17.3—Responses of coastal marsh ecosystems to N fertilization. Note an asterisk (*) indicates that N addition units were converted to kg N ha-1 yr-1 when different units were given in the publication. Site N Input Indicator Responses Reference Walden Creek, NC 0, 280, 560, and 1120 kg N ha-1 yr-1 Smooth cordgrass (Spartina alternifl ora) (short and tall) 1) increased the growth of short smooth cordgrass, but had no effect on tall smooth cordgrass; 2) NH4+ showed higher growth stimulation on short smooth cordgrass than NO3- 3) biomass production of short smooth cordgrass increased linearly with N addition Mendelssohn et al. 1979 Great Sippewissett Marsh, MA 26 and 76 kg N ha-1wk-1 from June to Sept (~416 and 1216 kg N ha-1 yr-1*) Smooth cordgrass N addition 1)increased live aboveground biomass 100 and 270% in the low and high addition, respectively; 2) increased leaf area coverage; 3) doubled the rate of 

evapotranspiration 

Howes et al. 1986

San Diego Bay, CA

15 g N m-2·wk-1 from March 

to Aug 1995 and 1996 (~3600 kg N ha-1 yr-1*)

N addition 1) increased plant tissue N concentration; 2) did not increase soil N concentration

Boyer et al. 2000

Narragansett Bay, RI

N gradient from 2 to 10,253 kg N ha-1 yr-1

Saltmeadow cordgrass (Spartina patens); smooth cordgrass (short and tall)

N addition 1) reduced species richness; 2) decreased the density and extent of saltmeadow cordgrass; 3) increased the density and extent of tall smooth cordgrass

Wigand et al. 2003

Tomales Bay, CA

N addition of 15 g N m- 2

urea fertilizer every two weeks from March 2000-July 2001 (~2100 to 3000 kg N ha-1 yr-1*)

Saltgrass (Distichlis spicata), Virginia glasswort (Salicornia virginica), marsh rosemary (Limonium californicum), Canadian sandspurry (Spergularia caadensis), arrowgrass (Triglochin sp).

N addition 1) increased soil nitrogen; 2)increased height, plant tissue N concentration in all species; 3)increased biomass of saltgrass, marsh rosemary, Canadian sandspurry, arrowgrass; 4) increased percent cover of all species at one or more times during the experiment

Traut 2005
Chapter 17—Wetlands GTR-NRS-80 201

San Francisco Bay (CA); Willapa Bay (WA)

Added 800 kg N ha-1 yr-1 Smooth cordgrass;

Hybrid Spartina (S.alternifl ora x 

S. foliosa); 

Virginia glasswort

1) increased biomass, stem density and the rate of habitat invasion; 2) promoted the spread of invaded species (hybrid Spartina); 3) decreased the density of native species (Virginia glasswort)

Tyler et al. 2007

Nags Creek (RI) 2 g N m- 2 added once per 

month from May-Aug for 3 years (~80 kg N ha-1 yr-1*)

Saltmeadow cordgrass N addition 1) increased bacterial production; 2) increased 

pore water NH4+ and NO3

- concentrations; 3) increased the 

uptake of dissolved inorganic phosphorous; 4) inconclusive effects on mineralization and denitrifi cation rates

Caffrey et al. 2007

Cocodrie (LA) N additions made once per 

month from April- Aug 2004 (~0, 230, 465, 930, 1860 and 3720 kg N ha-1 yr-1*) 

Smooth cordgrass N addition 1) increased the aboveground biomass and stem 

density; 2) did not change live belowground biomass

Darby and Turner 2008b

York River (ME); Nonesuch River (ME)

N addition of 1630 kg N ha-1

yr-1 for 3 years

Salt, brackish and oligohaline marshes

N addition 1) increased the productivity of all three types of marshes by 100% after 3 years; 2) lowered C:N and increased N:P in plant tissue in all marsh-types; 3) had no signifi cant effect on species composition of salt and brackish marshes, however a combination of N and P fertilization changed the species composition of oligohaline marshes

Crain 2007

Narragansett Bay (RI)

Added 30% N by weight of 60 g m-2 per month from 

May to August 2005 and 2006 (~720 kg N ha-1 yr-1*)

Smooth cordgrass N was added with phosphorous and potassium. The 

herbivore consumption of primary production was small in pristine marshes, but increased to 50-75% in eutrophic marshes 

Bertness et al. 2008

62 New England estuaries

N gradient from 24.2 to 928 kg N ha-1 yr-1

Eelgrass N loading greater than 50 kg N ha-1 yr-1 led to a signifi cant 

decline of eelgrass coverage, and almost all eelgrass habitats were lost at loading levels higher than 100 kg N ha-1 yr-1.

Latimer and Rego 2010
202 Chapter 17—Wetlands GTR-NRS-80

17.4 Critical load estimates

17.4.1 Freshwater Wetlands

Th ere is correlative evidence that Sphagnum-dominated 

freshwater wetlands in the United States are altered 

by N deposition. Th e critical load for altered peat 

accumulation and NPP is between 2.7 and 13 kg N ha-1

yr -1, based on observations from Aldous (2002a), Moore 

et al. (2004), Rochefort et al. (1990) and Vitt et al. 

(2003). Th e upper end of this critical load range is based 

on measurements of wet deposition only (10 to 13 kg N 

ha-1 yr -1) ( Aldous 2002a, 2002b) and therefore does not 

refl ect total N loading. Th ere is evidence showing that N 

deposition alters both the morphology and population 

dynamics of purple pitcher plant. Th e empirical evidence 

suggests a critical load to protect the population of 

purple pitchers of 10 to 14 kg N ha-1 yr-1 (Gotelli and 

Ellison 2006), while matrix modeling to forecast longterm population sustainability based on observations of 

population demographics suggests a lower value of 6.8 

kg N ha-1 yr -1 (Gotelli and Ellison 2002).

17.4.2 Intertidal Wetlands

Th e critical load range for atmospheric deposition is 

diffi cult to establish for intertidal wetlands because 

they have open nutrient cycles which are often strongly 

aff ected by N loading sources other than atmospheric 

deposition. Typically, the amount of N added in 

experimental treatments simulates total N input and 

therefore far exceeds the amount that U.S. intertidal 

wetlands would receive by atmospheric deposition. Only 

two studies (Table 17.3) have addition levels below 

100 kg N ha-1 yr -1. Based on the results of Wigand et 

al. (2003), a critical load to protect the community 

structure of salt marshes is likely to be 63 to 400 kg 

Figure 17.3—The effect of N loading on plant community structure of 10 salt marshes in Narragansett Bay (graphs plotted 

from data presented in Wigand et al. 2003)

















      

marsh N load (kg N/ha/yr)

Species richness

















      

marsh N load (kg N/ha/yr)

Density S. patens shoots (m2

)















      

marsh N load (kg N/ha/yr)

Density (tall) S. alterniflora shoots (m2

)













      

marsh N load (kg N/ha/yr)

Density (short) S. alterniflora shoot (m2) 
Chapter 17—Wetlands GTR-NRS-80 203

N ha-1 yr -1. Caff rey et al. (2007) provide additional 

evidence that 80 kg N ha-1 yr -1 alters microbial activity 

and biogeochemistry. Latimer and Rego (2010) found 

that eelgrass coverage started to decrease rapidly at N 

loading higher than 50 kg N ha-1 yr-1, with no eelgrass 

at loading levels higher than 100 kg N ha-1 yr-1. Note 

that these values are the total N loading to salt marshes, 

including N deposition directly to the marsh surface, 

as well as N deposited indirectly to the watershed, 

surface or ground water, and runoff from agriculture, 

urban areas, and other sources. Additional experimental 

evidence on ecosystem response to N loads that are 

similar to the amount of loading due to N deposition 

is needed to improve the critical load calculation for 

intertidal wetlands in the United States.

17.5 Comparison to Critical Loads 

from Europe

Th e 2003 assessment of N critical loads in Europe 

designated critical loads for multiple types of wetlands, 

including raised and blanket bogs, poor fens, rich fens, 

mountain rich fens, and intertidal wetlands (Bobbink et 

al. 2003). Th ere are numerous publications on N eff ects 

to wetlands in Europe compared to the United States. 

In general, documented responses include eff ects on 

growth and species composition, competition between 

species, peat and peat water chemistry, decomposition, 

and nutrient cycling. A brief summary of the European 

critical loads for wetlands is presented here and 

summarized in Table 17.4.

Bobbink et al. (2003) assigned a critical load of 5 to 

10 kg N ha-1 yr-1 for bog ecosystems, based on plant 

community and species responses to N deposition, 

and indicated that precipitation and P limitation 

should be used to assign critical loads to individual 

sites. Th e observed changes in the plant communities 

of ombrotrophic bogs included the replacement of 

Sphagnum-forming species with nitrophilous moss 

species (20 to 40 kg N ha-1 yr-1 in Dutch bogs) (Greven 

1992); the absence of characteristic Sphagnum species 

in British bogs (30 kg N ha-1 yr-1) ( Lee and Studholme 

1992); and reduction in the growth survivorship of 

characteristic bog species roundleaf sundew (Drosera 

rotundifolia) (10 kg N ha-1 yr-1 in Swedish bogs) 

(Bobbink et al 2003, Redbo-Torstensson 1994). Th e 

European critical load is similar to the range of critical 

loads suggested for freshwater wetlands in the United 

States (Table 17.5).

It is unclear how European critical loads for poor fens, 

rich fens, and mountain rich fens relate to the critical 

loads for similar wetlands in North America because 

the impacts of N loads have not been studied in the 

United States. European poor fens have a critical load 

of 10 to 20 kg N ha-1 yr-1 based on increased sedge and 

vascular plants and negative eff ects on peat mosses. Th e 

critical load for rich fens in Europe is 15 to 35 kg N ha-1

yr-1 based on increased tall graminoids and decreased 

diversity. Th e critical load for montane rich fens in 

Europe was 15 to 25 kg N ha-1 yr-1 based on increased 

vascular plants and decreased bryophytes. Note that 

changes in the vegetation composition and structure 

likely aff ect fauna species assemblages, such as spiders 

and beetles living in the originally open bog vegetation, 

and ground-breeding birds. Increased nutrient 

Table 17.4—European empirical critical loads of nutrient N for mire, bog and fen habitats 

(EUNIS class D). Reliability rating: ## reliable, # quite reliable and (#) expert judgment.

Ecosystem type kg N ha-1 yr-1 Reliability Indication of exceedance

Raised bogs and 

blanket fens

5-10 ## Increased vascular plants, decrease 

bryophytes, altered growth and species 

composition of mosses, increased N in peat 

and peat water

Poor fens 10-20 # Increases in sedges and vascular plants, 

negative effects on peat mosses

Rich fens 15-35 (#) Increases on tall graminoids, decrease diversity

Montane rich fens 15-25 (#) Increase vascular plants, decrease bryophytes

From Bobbink et al. 2003; Table 6.1
204 Chapter 17—Wetlands GTR-NRS-80

availability results in an increase of the nutrient content 

of plant material (Limpens et al. 2003a, Tomassen 

et al. 2003) and algal growth (Gulati and DeMott 

1997, Limpens et al. 2003b), and aff ects herbivorous, 

detritivorous, and carnivorous invertebrates (Van 

Duinen et al. 2004).

Th e European critical load for salt marshes, based on 

expert judgment, is 30 to 40 kg N ha-1 yr-1 (Bobbink 

et al. 2003), but studies of European salt marshes are 

limited. High levels of N input (65 to 70 kg N ha-1

yr-1) signifi cantly increased biomass production in 

the Netherlands (Van Wijnen and Bakker 1999); no 

changes in species composition and in diversity have 

been observed for the current deposition of 15 to 25 

kg N ha-1 yr-1 at sites in the Netherlands and Germany 

(Bobbink et al. 2003). Th e critical load for North 

American intertidal ecosystems may be closer to these 

values than to the high levels of N input previously 

studied (section 17.3.2).

17.6 Future Research Directions and 

Gaps in Data

An great diversity of wetlands ecosystems are found 

across Northern America, but experimental studies on 

the impacts of N enrichment in these wetland systems 

are scarce. Some data are present for bog systems and 

intertidal wetland (salt marshes), but even in these 

ecosystems fi eld additions studies with realistic N loads 

are rare. Realistic levels of long-term (5 to 10 years) 

experimental N addition in low background regions, 

with a high resolution of treatments, are needed to 

refi ne and establish reliable critical loads for wetlands. 

In addition, these studies should be done across a range 

of wetlands types in the United States, especially in 

oligotrophic to mesotrophic wetland habitat.

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# fairly reliable; (#) expert judgment

Ecosystem Critical loads for N 

deposition

kg N ha-1 yr-1

Reliability Response Study

Freshwater 

wetlands

2.7-13 # Peat accumulation and 

NPP 

Aldous 2002a, Moore et 

al. 2004, Rochefort et al. 

1990, Vitt et al 2003

Freshwater 

wetlands

6.8-14 (#) Pitcher plant community 

change

Gotelli and Ellison 2002, 

2006

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structure, microbial 

activity and 

biogeochemistry

Caffrey et al. 2007, 

Wigand et al. 2003 

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C.M. 2005. Limited eff ects of six years of 

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1197-1204.

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Natural nitrogen fi lter fails in polluted raised bogs.

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157(2): 339-347.

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F.; Zijlstra, J.D. 2003b. Th e interaction between 

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Lindau, C.W.; Delaune, R.D.; Pardue, J.H. 1994. 

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CO2 sink may be largely off set by stimulated N2O 

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wetland ecosystems with particular reference to 

atmospheric deposition. Annual Review of Ecology 

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Richard, P.J.H. 2004. Patterns of nitrogen 

and sulfur accumulation and retention in 

ombrotrophic bogs, eastern Canada. Global 

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Redbo-Torstensson, P. 1994. Th e demographic 

consequences of nitrogen fertilization of a 

population of sundew, Drosera rotundifolia. Acta 

Botanica Neerlandica. 43: 175-188.

Rochefort, L; Vitt,D.H.; Bayley, S.E. 1990. Growth, 

production and decomposition dynamics of 

Sphagnum under natural and experimentally 

acidifi ed conditions. Ecology. 71(5):1986-2000.

Shaver, G.R.; Melillo, J.M. 1984. Nutrient budgets of 

marsh plants - effi ciency concepts and relation to 

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Tomassen, H.B.M.; Smolders, A.J.P.; Lamers, L.P.M.; 

Roelofs, J.G.M. 2003. Stimulated growth of Betula 

pubescens and Molinia caerulea on ombrotrophic 

bogs: role of high levels of atmospheric nitrogen 

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Traut, B.H. 2005. Eff ects of nitrogen addition and 

salt grass Distichlis spicata. upon high salt marsh 

vegetation in Northern California, USA. Estuaries. 

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Tyler, A.C.; Lambrinos, J.G.; Grosholz, E.D. 2007. 

Nitrogen inputs promote the spread of an invasive 

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Assessment, Environmental Criteria and Assessment 

Offi ce; 3v. Available from: NTIS, Springfi eld, VA; 

PB95-124533, PB95-124525, and PB95-124517.

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nitrogen and sulfur- ecological criteria. Rep. No. 

EPA/600/R-08/082F. Research Triangle Park, NC: 

U.S. Environmental Protection Agency, National 

Center for Environmental Assessment, Environmental 

Media Assessment Group.

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Peeters, T.M.J.; Esselink, H. 2004. Do raised bog 

restoration measures rehabilitate aquatic fauna 

diversity? A comparative study between pristine, 

degraded, and rewetted raised bogs. In: Päivänen, 

J., ed. Wise use of peatlands. Proceedings of the 

12th International Peat Congress; 2004 June 6-11; 

Tampere, Finland. Jyväskylä, Finland: International 

Peat Society.

Van Wijnen, H.J.; Bakker, J.P. 1999. Nitrogen and 

phosphorus limitation in a coastal barrier salt 

marsh: the implications for vegetation succession.

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Vitt, D.H. 1994. An overview of factors that infl uence 

the development of canadian peatlands. Memoirs 

of the Entomological Society of Canada. 169: 7-20.

Vitt, D.H.; Wieder, K.; Halsey, L.A.; Turetsky, M. 

2003. Response of Sphagnum fuscum to nitrogen 

deposition: A case study of ombrogenous peatlands 

in Alberta, Canada. Th e Bryologist. 1062: 235-245.
208 Chapter 17—Wetlands GTR-NRS-80

Ward, D.H.; Morton, A.; Tibbitts, T.L.; Douglas, 

D.C.; Carrera-Gonzalez, E. 2003. Long-term change 

in eelgrass distribution at Bahia San Quintin, 

Baja California, Mexico, using satellite imagery.

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Nilsson, M.B.; Ericson, L. 2007. Global change 

shifts vegetation and plant-parasite interactions in 

a boreal mire. Ecology. 88: 454-464.

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Megonigal, P.J. 2009. Responses of Avicennia 

germinans (black mangrove) and the soil microbial 

community to nitrogen addition in a hypersaline 

wetland. Estuaries and Coasts. 32: 926-936.

Wigand, C.; McKinney, R.A.; Charpentier, M.A.; 

Chintala, M.M.; Th ursby, G.B. 2003. Relationships 

of nitrogen loadings, residential development, and 

physical characteristics with plant structure in New 

England salt marshes. Estuaries. 26: 1494-1504.
Chapter 18—Inland Surface Water GTR-NRS-80 209

18 INLAND SURFACE WATERS

J.S. Baron, C.T. Driscoll, J.L. Stoddard

nonpoint sources. Th e focus of this chapter is on surface 

waters that primarily receive N inputs from atmospheric 

deposition directly or indirectly through the watershed.

Freshwater aquatic ecosystems in which N has been 

observed to infl uence ecological processes either receive 

chronic or seasonally high inputs. Chemical responses 

range from long-term or temporary acidifi cation 

(Dumont et al. 2005), to measurable N concentrations 

in the water column (Stoddard 1994). Nitrogen-limited 

waters respond rapidly to inputs of N with increased 

productivity, or eutrophication (Baron et al. 2000, 

Bergström and Jansson 2006).

Th e number of surface waters that are sensitive to 

or impacted by acidic deposition is relatively small 

compared to the total number of surface waters in the 

United States (Charles 1991). Th e areas of the United 

States with low values of acid neutralizing capacity 

(ANC <400 μeq/L) are shown in Figure 18.1. Of the 

areas shown, only those with ANC values <100 μeq/L 

would be considered sensitive to acidifi cation by N 

deposition. Th ese areas include the northern, central, 

and southern Appalachian Mountain regions, portions 

of southern New England, the Pine Barrens of New 

Jersey, portions of Georgia and Florida, portions of the 

upper Midwest, portions of the Rockies, the Cascades 

and the Sierra Nevada Mountains of the West. Note 

18.1 Ecosystems Description

Freshwater aquatic ecosystems include rivers and 

streams, large and small lakes, reservoirs, and ephemeral 

ponds. Wetlands are defi ned and discussed in Chapter 

17 of this report. It is estimated that there are 123,400 

lakes with a surface area greater than 4 ha in the United 

States. Most lakes, however, are smaller than 4 ha; small 

lakes account for the majority of lake surface area both 

globally and in the United States (Table 18.1; Downing 

et al. 2006). Th e density of lakes varies greatly by region 

of the country, from 8.4 lakes per 100 km2

 in the 

upper Midwest and 7.8 lakes per 100 km2

 in Florida, 

to much lower values in other areas of the country (e.g., 

mid-Atlantic, Southeast, and West <1.0 lakes per 100 

km2

) ( Eilers and Selle 1991). Th e cumulative surface 

area of these lakes is approximately 9.5 million ha. Th e 

U.S. Geologic Survey's National Hydrographic Dataset 

(NHD) estimates that there are approximately 1.1 

million km of perennial fl owing streams in the United 

States. Of these about 91 percent are fi rst through 

fourth order (“wadeable”) (US EPA 2006).

18.2 Ecosystem Responses to N 

Deposition

A cascade of environmental eff ects occurs in response 

to elevated inputs of reactive N (Chapter 1; Galloway 

et al. 2003), but in freshwater ecosystems eff ects can 

be grouped under the categories of eutrophication 

(excess nutrient N) and acidifi cation. Reactive N from 

anthropogenic activities enters lakes and streams from 

a number of sources, including wastewater treatment 

point sources, agricultural and other nonpoint sources, 

and atmospheric deposition (Driscoll et al. 2003b). Th e 

relative importance of these reactive N sources varies 

by the position of lakes and streams on the landscape. 

In protected headwater areas (such as national or state 

forests or parks), atmospheric deposition is the only 

source of anthropogenic reactive N. Surface waters 

with drainage from agricultural, industrial, or urban 

areas may receive proportionally less reactive N from 

atmospheric deposition and more from point and other 

Table 18.1—Number and distribution of lakes (>4 ha in 

surface area) in the United States (after US EPA 2010)

Lake size 

category

 ha

Total number of 

lakes in the U.S.

Cumulative Surface 

Area

ha

1-4 155,201 310,402

4-10 68,559 433,605

10-20 24,902 352,167

20-50 16,488 521,396

50-100 6,134 433,739

>100 7,356 7,356,000

Total 278,640 9,407,310
210 Chapter 18—Inland Surface Water GTR-NRS-80 Total Alkalinity of Surface Waters by James M. Omernik , Glenn E. Griffith , Jeffrey T. Irish , and Colleen B. Johnson 

Environmental Research Laboratory U.S. Environmental Protection AgencyCorvallis, Oregon 97333

12 NSI, Technology Services Corp.

Corvallis, Oregon 97333

This map provides a synoptic illustration of the national patterns of surface water alkalinity in the conterminous United States. Alkalinity is the most readily available measure of the acid-neutralizing capacity of surface waters and provides a reasonable estimate of the relative potential sensitivity of lakes and streams to acidic deposition. Although the actual sensitivity of a water body depends on many watershed characteristics and processes, the low-alkalinity areas on the map indicate where sensitive surface waters are most likely to be found.The map is based on alkalinity data from approximately 39,000 lake and stream sites and the associations of the data values with factors such as land use, physiography, geology, and soils. Data were acquired from a variety of sources including federal and state agencies, university researchers. and private corporations. In many of the areas represented by a specific alkalinity range, an even greater range was observed in the water quality data. The shading on the map indicates the range of alkalinity within which the mean annual values of most of the surface waters of the area fall. Earlier alkalinity maps (1,2) depicted more generalized patterns of surface water alkalinity for the United States because of data base limitations and compilation conducted on small-scale maps. This map was compiled using more data and larger scale maps to provide a more precise national picture of surface water alkalinity. 1. Omernik, J.M., and C.F.. Powers. 1983. Total alkalinity of surface waters-a national map. Annals of the Association of American Geographers 73 (1):133-136. 2. Omernik, J.M., G.E. Griffith, and A.J. Kinney. 1985. Total alkalinity of surface waters. Corvallis Environmental Research Laboratory, U.S. Environmental Protection Agency, Corvallis, Oregon. 

200 0

0

200

400

400

400 800

mi

km

Albers Equal Area Projection

Scale 1:7,500,000

Total 

Alkalinity 

(μeq/l)

 <50 50-100100-200200-400 >400

U.S. ENVIRONMENTAL PROTECTION AGENCY 

ENVIRONMENTAL RESEARCH LAB - CORVALLIS

1988

1 2

2 2

Figure 18.1—Map of surface water values of acid neutralizing capacity (ANC) across the U.S. Note that ANC values less than 100 μeq L-1 are more sensitive to the effects 

of acidic deposition (Omernik and Powers 1982).
Chapter 18—Inland Surface Water GTR-NRS-80 211

that surface water acidifi cation mostly results from 

the combined eff ects of atmospheric sulfur (S) and N 

deposition.

Episodic acidifi cation from atmospheric N deposition 

has been reported mostly from snow-dominated 

environments where winter N deposition accumulates in 

the snowpack. Snowmelt triggers a pulse of reactive N to 

streams and lakes (Sickman and Melack 1998, Sickman 

et al. 2001, Stoddard 1994, Williams and Tonnessen 

2000). Both preferential elution of solutes, including N, 

from the snowpack, and N-rich soil solutions produced 

by N mineralization and nitrifi cation during the winter 

season of vegetation dormancy, are sources of spring 

nitrate (NO3

-

) pulses to streams and lakes (Brooks et al. 

1998, Stoddard 1994). Episodic acidifi cation associated 

with nitrate leaching can also occur during summer and 

fall rain events (Wigington et al. 1996).

Th e extent of lakes and streams aff ected by the 

eutrophication eff ects from atmospheric N deposition is 

much greater than those impacted by acidifi cation (see 

below).

In the United States, atmospheric N deposition eff ects 

on surface waters have received less attention than 

eff ects from atmospheric S deposition. A number of 

factors have combined to draw the scientifi c and policy 

focus away from N. First, eff orts to study the eff ects of 

atmospheric deposition had been largely directed toward 

surface water acidifi cation, where the focus has been 

on chronic acidifi cation, and particularly on the eff ects 

of S deposition on chronic acidifi cation (Stoddard et 

al. 2003). Th e acidifi cation response to N deposition is 

largely associated with episodic acidifi cation, which has 

received less attention. Sulfur emissions and deposition 

have been declining since 1973, and with substantial 

decreases since 1995 following the 1990 amendments 

of the Clean Air Act (Dennis et al. 2007). In contrast, 

changes in atmospheric N deposition have been variable 

(Lehmann et al. 2005). Atmospheric NO3

-

 deposition 

has generally decreased in the East following controls 

on emissions from electric utilities, but has increased 

in the west and central United States. Increases in 

ammonium (NH4

+

) deposition have been relatively 

widespread, particularly in the central and northern 

Midwest (Lehmann et al. 2005, 2007). Th e relative 

importance of N deposition has increased in recent years 

as S deposition has decreased, creating an opportunity 

to re-examine the role of N deposition in surface water 

acidifi cation (Lehmann et al. 2005).

Second, aquatic biologists have been focused on the 

role of phosphorus (P) in eutrophication of freshwaters 

for the past 40 years, largely due to the demonstrated 

eff ect of P increasing algal productivity worldwide 

(Schindler 1977, 1978). Th e prevailing wisdom is 

that P, an essential nutrient, is often the limiting 

nutrient to aquatic organisms, while N is tightly 

cycled in terrestrial systems, and that little N leaching 

occurs in undisturbed environments (Vitousek and 

Howarth 1991). Only within the past 20 years have 

there been studies questioning the established science 

and showing N limitation in some freshwaters, N 

leaching in some forest and alpine ecosystems, and 

N-induced acidifi cation in lakes and streams with low 

ANC. A global meta-analysis of 990 freshwater fi eld 

experiments found that N limitation of stream benthos, 

lake benthos, and phytoplankton was as common as P 

limitation from sites in all biomes worldwide (Elser et al. 

2007); experimental additions of N to these ecosystems 

stimulated primary production. While P-amendment 

experiments in some lakes and mesocosms have shown 

an increase in compensatory N fi xation, giving rise 

to the paradigm that P, not N, limits eutrophication 

(Schindler 1977, Schindler et al. 2008), many studies 

show N and P co-limitation (Elser et al. 1990, Elser 

et al. 2007). Th ere presently is no evidence for a 

consistently signifi cant contribution of planktonic N 

fi xation to oligotrophic and mesotrophic lakes (Lewis 

and Wurtsbaugh 2008). Oligotrophic (nutrient poor) 

and ultra-oligotrophic waters are inherently sensitive 

to even low inputs of atmospheric N deposition 

(Bergström et al. 2005, Bergstrom and Jansson 2006, 

Elser et al. 2009, Lewis and Wurtsbaugh 2008).

Th e current extent of N limitation can be estimated 

from nutrient data collected as part of regional lake 

surveys in the eastern (Linthurst et al. 1986) and 

western (Landers et al. 1987) United States. Table 18.2 

presents ratios of dissolved inorganic nitrogen (DIN; 

sum of NO3

-

 and NH4

+

) to total phosphorus (TP) 
212 Chapter 18—Inland Surface Water GTR-NRS-80

as an indicator of nutrient status of lakes. Th ese are 

based on water column analyses from the regional lake 

surveys. In Colorado lakes, Morris and Lewis (1988) 

found DIN:TP to be the most reliable predictor of 

nutrient limitation when compared against nutrient 

algal bioassays; they suggested using DIN:TP thresholds 

to identify probable N limitation (DIN:TP <4 by 

mass), P limitation (DIN:TP >12) and co-limitation 

(intermediate ratios). Percentages of N limited lakes vary 

from about 16 percent in the Adirondack Mountains, 

to 45 percent in the Rocky Mountains (Table 18.2). 

In general, N limitation is more common in regions 

with low N deposition (e.g., Sierra Nevada and Rocky 

Mountain ranges) and P limitation is more common in 

regions with higher N deposition (Adirondacks, New 

England, the Southeastern United States).

While the results in Table 18.2 suggest that N limitation 

is more widespread than is commonly understood, 

these values may actually underestimate historical 

rates of N limitation. Surveys of the literature report 

that oligotrophic waters are commonly N limited, 

especially undisturbed northern temperate or boreal 

lakes that receive low levels of atmospheric N deposition 

(Bergström et al. 2005, Elser et al. 1990, Elser et al. 

2007, Elser et al. 2009). A comprehensive study of 

available data from the northern hemisphere found 

unequivocal evidence of N limitation in lakes with low 

inputs of N, and increased N concentrations in lakes 

receiving N solely from atmospheric N deposition 

(Bergström and Jansson 2006). Studies of the response 

of phytoplankton to N or P amendments similarly 

found strong growth responses to N in lakes with low 

atmospheric N deposition, but no response to N in lakes 

with high N deposition (Elser et al. 2009). Th ese authors 

suggested that the majority of lakes in the northern 

hemisphere may have originally been N limited, and 

that long-term atmospheric N deposition has changed 

the balance of N and P in lakes such that P limitation is 

observed today.

18.2.1 Abiotic Responses to N Deposition

Th e primary abiotic eff ect of excess N deposition for 

inland surface waters is acidifi cation. Acidifi cation occurs 

when acidic deposition (primarily sulfuric (H2

SO4) and 

nitric (HNO3

) acids) exceeds the ability of watershed 

Table 18.2—Proportion of lakes in three nutrient limitation classes (N limited, P limited, and 

N and P co-limited) based on ratios of dissolved inorganic nitrogen (DIN; NO3

-

 + NH4

+

) to total 

phosphorus (TP). Data are from the Eastern Lake Survey (Linthurst et al. 1986), conducted in the 

fall of 1984, and the Western Lake Survey (Landers et al. 1987), conducted in the fall of 1985.

Region Number of Lakesa

Number of N 

limited lakes 

(%)b

Number of P 

limited lakes 

(%)c

Number of lakes with 

co-limited N:P 

(%)d

Adirondacks 1,290 208 (16%) 777 (60%) 3.5 (23%)

New England 4,361 1,470 (34%) 1,363 (31%) 1,529 (35%)

Poconos/Catskills 1,506 607 (40%) 559 (37%) 341 (22%)

Southeast 2,424 735 (30%) 1,148 (47%) 541 (22%)

Upper Midwest 8,755 3,142 (37%) 2,787 (32%) 2,646 (31%)

Rockies 6,666 2,998 (45%) 2,259 (34%) 1,409 (21%)

Sierra/Cascades 4,155 1,468 (35%) 1,958 (47%) 729 (18%)

Total 28,976 10,627 (36%) 10,851 (37%) 7,499 (26%)

a

The eastern and western Lake surveys were stratifi ed random samples of lakes; estimates of the number of lakes 

in each region are based on the target population sizes for each survey.

b

Lakes with DIN:TP ratios (by weight) less than four were characterized as N limited, based on the work of Morris 

and Lewis (1988).

c

Lakes with DIN:TP ratios (by weight) greater than 12 were characterized as P limited, based on the work of Morris 

and Lewis (1988).

d

Lakes with DIN:TP ratios (by weight) between 4 and 12 could not be assigned to nutrient limitation class, and are 

characterized as either co-limited, or limited by something other than N or P.
Chapter 18—Inland Surface Water GTR-NRS-80 213

soils, vegetation, and surface waters to neutralize these 

inputs. Catchment sensitivity to acidifi cation is largely 

governed by bedrock and surfi cial geology, and much 

of the United States is insensitive to acidic deposition 

(Charles 1991). Th e regions of most concern for 

acidifi cation are areas underlain by intrusive igneous 

rocks, typical of the western mountains, upper Midwest 

and parts of the Northeast, and base-poor sandstones 

typical of the Northeast, the mid-Atlantic, and the 

central and southern Appalachian Mountains (Figure 

18.1; Charles 1991). Catchment sensitivity is also 

infl uenced by past acidic deposition, which can diminish 

the ability of soil to neutralize ongoing acidic deposition, 

due to long-term declines in exchangeable base cations 

(Driscoll et al. 2001, Huntington et al. 2000).

Surface water sensitivity to acidifi cation is commonly 

measured in terms of pH and ANC. Th ese indicators 

decline as lakes and streams acidify, and waters are 

classifi ed as acidic when ANC values decrease to less 

than 0 μeq L-1. Acidifi cation can be further subdivided 

into chronic acidifi cation (when lakes and streams are 

acidic year-round) and episodic acidifi cation (shortterm decreases in pH and ANC, lasting on the order of 

hours to weeks). In water bodies with low ANC, strong 

hydrologic events can temporarily decrease ANC. While 

N deposition can infl uence both chronic and episodic 

acidity, it has the most marked eff ect on episodic 

acidifi cation (Murdoch and Stoddard 1992, Schaefer et 

al. 1990, Wigington et al. 1996).

Severe episodes can produce conditions that are as 

deleterious to biota as chronic acidifi cation (Baker et al. 

1996). Th e mechanisms that produce acidic episodes 

include dilution, and fl ushing of NO3

-

, sulfate (SO4

2-), 

and/or organic acids from forest soils or snowpack 

(Baker et al. 1996, Kahl et al. 1992, Lawrence 2002, 

Wigington et al. 1996, Williams and Tonnessen 

2000). Acidic deposition may contribute to episodic 

acidifi cation by supplying N to sensitive watersheds 

(producing pulses of NO3

-

 during high fl ow events), 

producing hydrologically mobile SO4

2-, and by lowering 

baseline pH and ANC so that even modest episodes 

are suffi cient to produce acidic (i.e., ANC <0 μeq L-1) 

or near-acidic conditions (i.e., low but positive values 

of ANC). A generally accepted conceptual model of 

episodic acidifi cation is that hydrologic events dilute 

base cation concentrations and trigger a pulsed increase 

in NO3

-

, causing short-term decreases in ANC (Figure 

18.2). In low ANC waters, hydrologic events can 

result in decreases in ANC to near or below 0 μeq L-1, 

causing adverse ecological eff ects (Baker et al. 1996). 

Th e deposition and associated leaching of SO4

2- lowers 

the ANC of surface waters during basefl ow, so that 

short-term pulses of NO3

-

 (and, to a lesser extent, pulses 

of stored SO4

2- and naturally produced organic acids) 

Flow

Q

NO3

-

SO4

2-

ANC

0

acidic epsiode

chronically acidic

Figure 18.2—Conceptual diagram of changes in surface 

water chemistry of two contrasting upland streams impacted 

by atmospheric deposition in response to a hydrologic 

event. The upper two panels are discharge (Q) and nitrate 

(NO3

-

) during the stream event; the pattern is identical 

for both streams. The two streams have two different 

concentrations of SO4

2- in the third panel; the solid line 

represents a stream with high SO4

2- and the dashed line is a 

stream with low SO4

2-. The fourth panel is acid neutralizing 

capacity (ANC). The stream with the higher SO4

2- (solid line) 

has a lower ANC during basefl ow, and is chronically acidic. 

During the episodic event, ANC values decrease somewhat 

from low basefl ow values, due to the pulsed input of NO3

-

. 

The low SO4

2- stream (dashed line) has low, but positive 

ANC during basefl ow. At this site the pulse of NO3

-

 during 

the event causes a short-term decrease in ANC to acidic 

conditions (ANC < 0 μeq/L).
214 Chapter 18—Inland Surface Water GTR-NRS-80

increase the severity of episodic or seasonal acidifi cation 

(Driscoll et al. 2001, Stoddard et al. 2003, Wigington et 

al. 1996).

18.2.2 Biotic Responses to N Deposition

Th e biotic eff ects of excess N deposition for inland 

surface waters can be caused by both acidifi cation and 

eutrophication. Acidifi cation can have toxic eff ects on 

animals (zooplankton, macroinvertebrates, and fi sh), 

often as a result of the mobilization of aluminum and 

low pH conditions (Driscoll et al. 2001). Nutrient N 

primarily aff ects lower trophic levels (phytoplankton, 

periphyton) by increasing productivity in N limited 

systems (Wetzel 2001). Increased productivity can lead 

to secondary eff ects, including changes in the species 

composition or biodiversity of algae. Whether low or 

modest N enrichment infl uences animal populations is 

poorly understood.

Although chronically high acid levels stress aquatic 

life, acidic episodes are particularly harmful because 

abrupt, large changes in water chemistry allow fi sh and 

other organisms few areas of refuge (Baker et al. 1996). 

High concentrations of dissolved inorganic aluminum 

(Al) are directly toxic to fi sh, and pulses of Al during 

acidic episodes are a primary cause of fi sh mortality 

(Baker et al. 1996, MacAvoy and Bulger 1995, Van 

Sickle et al. 1996). Lethal eff ects are brought about by 

a combination of high acidity and Al concentrations, 

which disrupt the salt and water balance of blood in a 

fi sh (Driscoll et al. 2001), and respiratory stress caused 

by gill damage and mucous clogging of gills (Baker and 

Schofi eld 1982).

For the analysis of nutrient N eff ects, we relied on 

papers and studies that linked aquatic biological and 

ecological response to atmospheric deposition, but the 

results are consistent with laboratory or in situ doseresponse studies and even land-use change studies. 

Essentially, the productivity of minimally disturbed 

aquatic ecosystems is often limited by the availability 

of N, and slight increases in available N trigger a rapid 

biological response that increases productivity and 

rearranges algal species assemblages (Lafrancois et al. 

2004, Michel et al. 2006, Nydick et al. 2004, Saros et 

al. 2005). Th e mechanism for change is alteration of 

N:P ratios, which can increase productivity of some 

species at the expense of others (Elser et al. 2009). 

Higher trophic levels (zooplankton, macroinvertebrates) 

may be secondarily aff ected by N, but further increases 

in primary, or autotrophic, production will be limited 

by other nutrients such as P or silica (Si).

18.3 Range of Responses Observed

18.3.1 Observed Abiotic Responses to N Deposition

Episodic acidifi cation is more widespread than chronic 

acidifi cation and can be important in acid-sensitive 

regions. For example, roughly 10 percent of Adirondack 

lakes were chronically acidifi ed in the early 1990s, 

but more than three times that number (31 percent) 

underwent episodic acidifi cation (Lawrence 2002). In 

a study of episodic acidifi cation in the northeastern 

United States, Wigington et al. (1996) ranked the 

importance of various mechanisms in producing shortterm acidity. Pulses of NO3

-

 were ranked as the second 

most important process in the Adirondack and Catskill 

mountains; only dilution of base cations, which occurs 

in virtually all surface waters under increased fl ow, 

ranked higher.

In the western United States, there are no lakes that 

are chronically acidic, but episodic acidity has been 

observed in both the Colorado Front Range and Sierra 

Nevada Mountains, where episodic NO3

-

 leaching 

during snowmelt decreases ANC to 0 μeq L-1 or below 

(Leydecker et al. 1999, Stoddard 1995, Williams and 

Tonnessen 2000).

18.3.2 Observed Biotic Responses to N Deposition

Algae are the aquatic biota most likely to show 

rapid eutrophication eff ects from N deposition. 

Eutrophication eff ects are most likely to occur in fresh 

waters that historically received low nutrient inputs. 

Especially in the western United States, these ecosystems 

are mainly located at high elevations. Studies have 

shown an increase in phytoplankton biomass in lakes 

with increasing N deposition in the Snowy Range in 

Wyoming, the Colorado Front Range, Sierra Nevada, 

Sweden, and across Europe (Table 18.3; Bergström 

et al. 2005, Bergström and Jansson 2006, Elser et 

al. 2009, Lafrancois et al. 2004, Nydick et al. 2004, 
Chapter 18—Inland Surface Water GTR-NRS-80 215

Sickman et al. 2003). Paleolimnological reconstructions 

show a stronger eutrophication response in lakes where 

N deposition is higher than in lakes with lower N 

deposition (Das et al. 2005; Enders et al. 2008; Saros et 

al. 2003; Wolfe et al. 2001, 2003; Table 18.4).

Community shifts in phytoplankton were observed in 

the Snowy Range, with chrysophytes (golden algae) 

favored in lakes having lower N and cyanophytes 

(photosynthetic bacteria, also called blue-green algae) 

and chlorophytes (green algae) favored in lakes having 

higher N (Lafrancois et al. 2004). Sediment cores from 

lakes in the Colorado Front Range showed increasing 

representation of mesotrophic diatoms in recent times, 

as compared with predevelopment conditions (Enders 

et al. 2008; Wolfe et al. 2001, 2003). In Lake Tahoe, 

California, there has been a sharp increase in the ratio 

of araphidinate pennate (characteristic of nutrient 

rich conditions) to centric (characteristics of nutrient 

poor conditions) diatoms since about 1950 (largely 

due increases in Fragilaria crotenensis), associated with 

increased N loading to the lake. Jassby et al. (1994) 

found that atmospheric deposition supplies most of the 

N to Lake Tahoe. While runoff from catchment streams 

supplies organic N to Lake Tahoe (Coats and Goldman 

2001), a recent meta-analysis fi nds little relation 

between DON runoff and atmospheric N deposition 

(Pellerin et al. 2006).

Two species of diatom, Asterionella formosa and 

Fragilaria crotonensis, now dominate the fl ora of at 

least several alpine and montane Rocky Mountain 

lakes (Baron et al. 2000; Interlandi and Kilham 1998; 

Saros et al. 2003, 2005; Wolfe et al. 2001, 2003). 

Th e growth of A. formosa has been stimulated with N 

amendments during in situ incubations, using bioassays 

and mesocosms (6.4 to 1616.0 μmol L-1 N, McKnight 

et al. 1990; 76.0 μmol L-1, Lafrancois et al. 2004; 18.0 

μmol L-1, Saros et al. 2005). In situ incubations in large 

lakes of Yellowstone National Park also stimulated F. 

crotonensis (Michel et al. 2006). Th e N requirements 

for A. formosa and F. crotonensis were determined to 

be 0.041 μmol L-1 and 0.006 μmol L-1, respectively; 

higher concentrations stimulated growth (Michel et 

Table 18.3—Nitrate concentrations at which phytoplankton response has been observed

Site N concentration 

mg NO3

-

-N L-1

Response Study

Colorado Front Range 1.1 Diatom spp. growth 

stimulation

Lafrancois et al. 2004

Snowy Range, WY 1.0 Shift in species composition Nydick et al. 2004

Beartooth Mountains WY 0.9 Stimulation of A. formosa, 

F. crotonensis

Saros et al. 2005

Sweden 0.9 Increasing biomass Bergström et al. 2005

Europe and No. Amer. Lakes 0.9 Increasing biomass Bergström and Jansson 2006

Table 18.4—Biological response observed in paleolimnological studies

Site Year(s) Response (method) Study

Colorado Rocky 

Mountains

1950 Increased microbial activity, greater primary production, 

change in species composition (isotopic analysis)

Enders et al. 2008

1950 Greater primary production, stimulation of A. formosa, 

F. crotonensis, shift in species assemblages (pigment 

analysis). 

Das et al. 2005

1950 Greater primary production, Stimulation of A. formosa, 

F. crotonensis, shift in species assemblages (taxonomy 

and cell counts)

Wolfe et al. 2001, 2003

1950-1964 Threshold modeled at 1.5 kg N ha-1 yr-1 Baron 2006

Beartooth Mountain 

Range, WY

1990 Greater primary production, stimulation of A. formosa, 

F. crotonensis, shift in species assemblages

Saros et al. 2003
216 Chapter 18—Inland Surface Water GTR-NRS-80

al. 2006). Other species, including Tetracyclus glans, 

and Staurosirella pinnata, have shown negative growth 

responses to increased N (Michel et al. 2006). In studies 

of lake sediment diatom remains, typical oligotrophic 

species such as Aulacoseria perglabra, Cyclotella steligera, 

and Achnanthes spp. declined coincident with the rise in 

dominance of A. formosa and F. crotonensis (Wolfe et al. 

2003)

Interlandi and Kilham (1998) demonstrated that 

maximum species diversity was maintained when N 

concentrations were low (<3.0 μmol L-1) in lakes in 

the Yellowstone National Park (Wyoming, Montana) 

region. Th e implication is that species diversity 

declines with increasing availability of N. Th is fi nding 

complements the results of terrestrial studies that also 

showed a negative relationship between species diversity 

and N availability (Gilliam 2006, Suding et al. 2005).

Studies of Lake Erie found N to be seasonally, but 

only moderately, limiting to phytoplankton growth. 

Phytoplankton were far more responsive to additions 

of P, iron (Fe), or Si than N during diff erent seasons 

(Moon and Carrick 2007, North et al. 2007). Studies of 

Lake Superior have documented a century-long increase 

in lake NO3

-

 concentrations, which has produced a 

stoichiometric imbalance and strong P limitation (Finlay 

et al. 2007, Sterner et al. 2007).

Th ere have been few studies on the eff ects of nutrient 

enrichment from N deposition on aquatic animals. 

Nydick et al. (2004) found that zooplankton density 

and biomass were altered by changes in nutrients in 

enclosure experiments, probably as a result of poor 

food quality. A whole-ecosystem experiment at the Bear 

Brook watershed, Maine simulated the eff ects of N and 

S deposition by means of experimental ammonium 

sulfate ((NH4)2

SO4) addition over a 10-year period. 

Researchers found that elevated N inputs had minimal 

eff ect on stream detritus processing (Chadwick and 

Huryn 2005). Th ey also found that N additions had no 

signifi cant eff ect on stream macroinvertebrate secondary 

production or varying production by functional feeding 

groups. Th ey concluded that climate-related variables, 

such as fl ow duration and litter inputs, controlled 

secondary production when N was not limiting.

Nitrogen deposition appears to have stimulated 

productivity and altered algal species assemblages at 

deposition rates of 1.5 to 2.5 kg N ha-1yr-1 (Baron 

2006, Bergström and Jansson 2006). A meta-analysis 

of lakes from 42 regions of Europe and North 

America concluded that atmospheric N deposition 

was responsible for increased concentrations of N 

in lake water, and elevated phytoplankton biomass 

(Bergström and Jansson 2006). Bergström and 

Jansson (2006) found a consistent pattern of nutrient 

limitation, showing N limitation for deposition below 

approximately 2.5 kg N ha-1yr-1, co-limitation of N and 

P for deposition between ~2.5 and 5.0 kg N ha-1yr-1, and 

P limitation in areas with N deposition greater than 5.0 

kg N ha-1yr-1.

18.4 Critical Loads Estimates

18.4.1 Critical Loads for Abiotic Responses to N 

Deposition

Despite the diffi culty of predicting current and future 

N behavior in lakes and streams, as discussed below 

(see section 18.6), both mechanistic (Aber and Driscoll 

1997) and conceptual (Stoddard 1994) models of N 

saturation predict that elevated rates of N deposition 

will eventually produce elevated runoff concentrations 

of NO3

-

 in relatively undisturbed catchments. Empirical 

cross-site watershed studies show the relationship 

between watershed N concentrations or loss as a 

function of atmospheric N deposition (Aber et al. 2003, 

Baron et al. 2000, Elser et al. 2009, Fenn et al. 2003, 

Sickman et al. 2002).

Such cross-site analyses were originally conducted for 

Europe (Dise and Wright 1995, Wright et al. 2001), 

and typically show a “dog-leg” pattern—low surface 

water NO3

-

 concentrations or leaching with low levels 

of N deposition, and elevated leaching losses and 

concentrations above a “threshold” of deposition. For 

the European data, the threshold is roughly 10 kg N ha1

yr-1 (Wright et al. 2001).

We examined watershed NO3

-

 concentrations as a 

function of total atmospheric N deposition for the 

eastern (Aber et al. 2003) and western United States 

(Sickman et al. 2002). Relationships are presented 

for annual average NO3

-

 and peak springtime NO3

-
Chapter 18—Inland Surface Water GTR-NRS-80 217

in Figure 18.3. We consider the development of 

relationships between lake and stream NO3

-

 and 

atmospheric N deposition to be the fi rst step toward 

empirical critical loads for N in these aquatic 

ecosystems. Because both abiotic and biotic responses 

to N deposition in surface water begin only when excess 

NO3

-

 begins to leach from catchments, the thresholds 

suggested in Figure 18.3 imply the lower limits for N 

Figure 18.3—Relationship between mean and peak annual nitrate concentration (μM) and nitrogen 

deposition (kg N ha-1 yr-1) in (a) forested catchments in Europe; (b) forested catchments in the 

northeastern United States; and (c) high elevation catchments in the western United States. 

European data from (Wright et al. 2001). Northeastern United States data are from Aber et al. 

(2003). Western NO3

-

 data are from Sickman et al. (2002) and Burns (2004). Western deposition 

was estimated from modeled NO3 deposition (Nanus et al. 2003) and NH4

+

 deposition from the 

nearest NADP station. Lines are drawn by eye to estimate the x- intercept, the value of N deposition 

at which signifi cant NO3

-

 leaching begins to occur.
218 Chapter 18—Inland Surface Water GTR-NRS-80

deposition below which eff ects are unlikely. Th e data 

suggest preliminary N critical loads of about 2 kg N ha1

yr-1 for the western mountains, and 8 kg N ha-1yr-1 for 

forested catchments in the East based on NO3

-

 leaching 

(Table 18.5). Th is value for the Western lakes can be 

considered a critical load of N for eutrophication. Note 

that there is some uncertainty in these estimates due 

to inherent diffi culty in quantifying dry N deposition 

(see Chapter 3 this report). In addition, deposition 

varies by elevation within watersheds, adding additional 

uncertainty (Weathers et al. 2006).

We anticipate that the critical loads of N for acidity 

would be higher than the threshold values for 

eutrophication. Note that it is essential to establish the 

values of watershed ANC and SO4

2- and other major 

solutes in order to determine the critical load for acidity 

(Leydecker et al. 1999, Meixner et al. 2004). Th is is 

most eff ectively done with steady-state or dynamic 

acidifi cation models that consider all major solutes 

simultaneously. We estimated critical loads of N for 

acidity using the literature and expert judgment. For 

Eastern lakes, we used TIME lake data (US EPA 2006) 

to identify lakes with low ANC values where a 10 μeq 

L-1 decrease in ANC, might increase the proportion 

of lakes exhibiting episodic acidifi cation (ANC < 0 

μeq L-1). Th is change could result from an increase of 

approximately 20 μeq L-1 in NO3

-

 during snowmelt 

based on the stoichiometric response of changes in 

ANC with additions in strong acid anions during 

snowmelt (Schaefer et al. 1990). Comparing average 

NO3

-

 concentrations with snowmelt values for Eastern 

lakes (Figure 18.3), suggests an N critical load for 

acidity of approximately 9 kg N ha-1yr-1. Th is analysis 

assumes that SO4

2- concentrations remain constant and 

are not changing. Th is value is similar to that derived for 

Northern Forests (Chapter 7; Aber et al. 2003), as well 

as that derived for European forested streams (Dise and 

Wright 1995), thus we consider it to be reliable.

Williams and Tonnessen (2000) reported episodic 

acidifi cation in a Colorado Front Range lake during 

spring snowmelt with annual wet N deposition rates of 

4 kg N ha-1yr-1 (Table 18.6). An ecosystem modeling 

eff ort with MAGIC (Model of Acidifi cation of 

Groundwater in Catchments) proposed an acidifi cation 

critical load for the Colorado Front Range of 8 kg N 

ha-1 yr-1 (Table 18.6; Sullivan et al. 2005). Because the 

empirical value came from a high elevation lake that 

was highly responsive to changes in deposition values, 

while the modeled value was for a stream with at least 

some surrounding vegetation, we consider the measured 

value of 4 kg N ha-1yr-1 to be fairly reliable, although 

more observations from additional lakes would lend 

strength to the critical load. Note these measurements 

are extremely diffi cult to obtain.

18.4.2 Critical Loads for Biotic Responses to N 

Deposition

Our estimate for N critical loads for eutrophication in 

the mountain West (i.e., 2 kg N ha-1 yr-1) is similar to 

other estimates of aquatic critical loads for the region 

(Baron 2006, Fenn et al. 2003a). Th e estimated aquatic 

critical N load is lower than for terrestrial ecosystems, 

which is suggested to be 4 kg N ha-1 yr-1 for tundra 

based on experimental studies (Bowman et al. 2006), 3 

to 4 kg N ha-1 yr-1 for tundra based on modeled studies 

(Baron et al. 1994), and 4 kg N ha-1 yr-1 for subalpine 

forest based on regional surveys (Rueth and Baron 

2002). Nitrogen deposition appears to have stimulated 

productivity and altered algal species assemblages at 

deposition rates of 1.5 to 2.5 kg N ha-1 yr-1 (Table 18.7). 

A hindcasting exercise determined that the changes in 

algal communities in Rocky Mountain National Park 

Table 18.5—Summary of critical loads of N for eutrophication and acidifi cation for Eastern and 

Western high elevation lakes. Reliability rating: ## reliable; # fairly reliable; (#) expert judgement

Western lakes

kg N ha-1 yr-1

Reliability for 

Western lakes

Eastern lakes

kg N ha-1 yr-1

Reliability for 

Eastern lakes

Critical load of N for 

eutrophication

2 ## 8 #

Critical load of N for acidity 4-8 # 9 (##)
Chapter 18—Inland Surface Water GTR-NRS-80 219

that occurred between 1850 and 1964 were associated 

with an increase in wet N deposition that was 1.5 kg 

N ha-1 yr-1 (Baron 2006; Table 18.7). Similar changes 

inferred from lake sediment cores of the Beartooth 

Mountains of Wyoming also occurred at about 1.5 kg N 

ha-1 yr-1 deposition (Table 18.7). Pre-industrial inorganic 

N deposition is estimated to have been 0.1 to 0.7 kg N 

ha-1 yr-1, based on measurements from remote parts of 

the world (Galloway et al. 2004, Holland et al. 1999). In 

the western United States, pre-industrial, or background, 

inorganic N deposition was estimated by Holland et al. 

(1999) to range from 0.4 to 0.7 kg N ha-1 yr-1.

Th e determination of N critical loads for eutrophication 

of eastern lakes and the Great Lakes is problematic. Our 

analysis of N-limited lakes in lake districts of the United 

States suggests that, surprisingly, there are relatively 

large numbers of N-limited lakes in the East (Table 

18.2). For those lakes, inputs of N associated with 

elevated atmospheric deposition are possibly stimulating 

plant productivity and altering algal communities, 

although mesotrophic lakes are often limited by P, and 

large lakes may be limited by Fe (Moon and Carrick 

2007, Schindler et al. 2008, Sterner et al. 2007, Wetzel 

2001). In a study of hundreds of Swedish lakes along 

Table 18.6—Critical load for N for acidity for surface water in the Rocky Mountains. RMNP=Rocky Mountain 

National Park

Ecosystem Site Critical load for 

N deposition 

kg N ha-1 yr-1

Response Comments Study

Alpine lake Southern Rockies/Niwot 4.0 Episodic freshwater 

acidifi cation

Observed Williams and 

Tonnesson 2000

Alpine stream Southern Rockies/Loch Vale 

RMNP

5 Episodic freshwater 

acidifi cation

Modeled Baron and Hartman 

2004

Alpine stream Southern Rockies/Loch Vale 

RMNP

7.3 Chronic freshwater 

acidifi cation 

Modeled Baron and Hartman 

2004

Alpine stream Southern Rockies/Loch Vale 

RMNP

8 Freshwater acidifi cation; 

ANC = 20 μeq L-1

Modeled Sullivan et al. 2005

Subalpine 

stream/lake

Southern Rockies/Loch Vale 

RMNP

12 Freshwater acidifi cation; 

ANC = 20 μeq L-1

Modeled Sullivan et al. 2005

Alpine stream Southern Rockies/Loch Vale 

RMNP

14 Freshwater acidifi cation; 

ANC = 0 μeq L-1

Modeled Sullivan et al. 2005 

Subalpine 

stream/lake

Southern Rockies/Loch Vale 

RMNP

21 Freshwater acidifi cation; 

ANC = 0 μeq L-1

Modeled Sullivan et al. 2005 

Table 18.7—Critical load for N for eutrophication for surface water in the Rocky Mountains. RMNP=Rocky 

Mountain National Park

Ecosystem Site Critical load for N 

deposition 

kg N ha-1 yr-1

Response Comments Study

Alpine lake Southern Rockies/

Loch Vale RMNP

1.5 Eutrophication Paleolimnological Baron 2006

Alpine lake Northern Rockies/

Beartooth Mountains, 

WY

1.5 Eutrophication Paleolimnological Saros et al. 2003

Alpine stream Southern Rockies/

Loch Vale RMNP

2 Freshwater 

eutrophication

Modeled Baron et al. 1994

Alpine lakes Rocky Mountains 2.5 Eutrophication, N 

and P co-limitation

Surveys and 

references therein

Bergström and 

Jansson 2006
220 Chapter 18—Inland Surface Water GTR-NRS-80

an N deposition gradient, lakes receiving greater than 

4 kg N ha-1 yr-1 had greater ratios of chlorophyll a:total 

P, signifying a eutrophication response (Bergström et 

al. 2005). In the eastern United States, it is unknown 

whether a reduction in atmospheric N emissions and 

deposition would aff ect productivity in the N-limited 

lakes. Because these ecosystems are currently receiving 

elevated N deposition and have been receiving elevated 

deposition for decades, it is diffi cult to determine a value 

for critical loads of N. Moreover, evaluating the benefi ts 

associated with control on NO3

-

 or NH4

+

 inputs in these 

N limited lakes is not straightforward. Inputs of NO3

-

and NH4

+

 are partially removed by in-lake processing 

(Kelly et al. 1987), which alters ANC. Th is process will 

infl uence the change in ANC associated with acidic 

deposition to eastern lakes.

18.5 Comparison to Critical Loads for 

Other Regions

Th e threshold of atmospheric N deposition at which 

elevated NO3

-

 leaching may occur is highest in Europe, 

intermediate in eastern North America and lowest in the 

West (Figure 18.3). Note that elevated leaching losses 

do not necessarily occur at deposition levels above these 

thresholds. Th ese thresholds only represent conditions of 

deposition above which elevated leaching has been noted 

at some sites and below which leaching rarely occurs. 

Th e mechanism(s) responsible for regional diff erences in 

the thresholds is not clear. Possible factors may include 

diff erences in the relative deposition of NH4

+

 and NO3

-

; 

more intensive forest and land management practices 

in Europe (Dise and Wright 1995), and diff erences 

in climate and hydrology. For a given region, even 

under elevated atmospheric N deposition, there is also 

considerable variability across sites. Th is variability may 

be due to several factors, the most important of which 

are likely land disturbance history (e.g., past forest 

cutting and agricultural practices) and climatic factors.

18.6 Future Research Directions and 

Limitations in Understanding

Adverse direct eff ects of N deposition on most fauna due 

to nutrient enrichment are probably minimal. Nitrogen 

concentrations alone are not high enough to infl uence 

animal metabolism, and the extent of eutrophication is 

insuffi cient (due to induced P limitation in oligotrophic 

waters) to cause oxygen depletion. Eff ects of N on 

invertebrates that are essential parts of aquatic food 

webs, however, are not well understood. Individual algal 

taxa are stimulated at low N concentrations, and several, 

including Asterionella formosa and Fragilaria crotonensis, 

have come to dominate some lakes that have received 

elevated N deposition (Saros et al. 2005, Wolfe et al. 

2003). Other species have declined commensurate with 

increased available N, but we found only one study 

that addressed the question of eff ects of N deposition 

on freshwater aquatic diversity. Clearly, more work is 

needed to clarify these eff ects.

Research is needed to address two other questions. First, 

how do freshwater ecosystems respond to decreases 

in inputs of N? Alternative responses, which are not 

mutually exclusive, include decreases in productivity, 

reversion of species assemblages to former compositions, 

or irreversible alterations that develop into novel 

ecosystems (Hobbs et al. 2006, Schindler et al. 2008). 

Second, how does climate change interact with N 

deposition to infl uence aquatic productivity and 

assemblages?

One manifestation of limited understanding of 

ecosystem response to atmospheric N deposition is our 

inability to eff ectively explain the long-term temporal 

patterns of surface water N concentrations and loss in 

watershed ecosystems. Nitrate concentrations have been 

increasing in Lake Superior since at least the 1930s, 

but the cause remains unknown, and examination of 

stable N isotopes shows a more complex N-cycling story 

than can be explained by atmospheric N deposition 

alone (Finlay et al. 2007). In the mountain West, a 

few monitoring sites have shown increases in surface 

water NO3

-

 attributable to increases in atmospheric N 

emissions and deposition (Jassby et al. 1994, Williams 

et al. 1996). Other sites have reported measurably 

high N concentrations since monitoring began in the 

1980s (Baron et al. 2000, Fenn et al. 2003a). In the 

Northeast, however, investigators have reported declines 

in surface water concentrations and loss of NO3

-

 in 

watersheds (Driscoll et al. 2003a, 2007; Goodale et 

al. 2003; Stoddard et al. 2003), with little change 
Chapter 18—Inland Surface Water GTR-NRS-80 221

in atmospheric N deposition. Th e mechanism(s) 

contributing to this decline in surface water NO3

-

 in 

the Northeast is not clear but may be infl uenced by 

carbon dioxide-fertilization of forest vegetation, shifts in 

the composition of tree species, increases in in-stream 

N retention and/or changes in climate (Bernhardt et 

al. 2005, Driscoll et al. 2003b, Goodale et al. 2003, 

Th omas et al. 2010).

Ecosystem models, as mentioned above, are useful 

for augmenting empirical studies for deriving critical 

loads. Th is is especially true for addressing acidifi cation 

critical loads, but ecosystem models have value for 

aiding in selection of critical loads for eutrophication 

where measurements are scarce. Ecosystem models, 

while widely used to represent N-cycling and eff ects, 

are still unable to represent some responses, refl ecting 

an inadequate understanding of ecosystem processes 

and drivers. Although there are many ecosystem 

N-cycling models, there are fl aws in all of them. For 

example, PnET-CN was developed to simulate N 

dynamics in forest ecosystems (Aber et al. 1997). While 

PnET-CN depicts many N-cycling processes, it overpredicts recent surface water NO3

-

 in eastern forest 

watersheds and fails to capture the recent declines in 

NO3

-

 concentrations (e.g., Aber et al. 2002). DayCentChem is another promising nutrient-cycling and 

geochemical model, but it also is not able to reproduce 

New England stream NO3

-

 trends (Hartman et al. 

2007, Hartman et al. 2009). Hydrochemical models 

such as the Alpine Hydrochemical Model (AHM) 

suff er from similar problems in reproducing chemical 

weathering or specifi c events (Meixner et al. 2004). As 

with many active ecosystem and hydrochemical models, 

however, AHM is constantly undergoing revisions and 

improvements (Molotch et al. 2008). Without eff ective 

simulation models, it will be diffi cult to develop critical 

loads for N and conduct comprehensive critical loads 

for acidity.

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228
Chapter 19—Synthesis GTR-NRS-80 229

19 SYNTHESIS

L.H. Pardo, L.H. Geiser, M.E. Fenn, C.T. Driscoll, C.L. Goodale, E.B. Allen, J.S. Baron, R. Bobbink, W.D. 

Bowman, C.M. Clark, B. Emmett, F.S. Gilliam, T. Greaver, S.J. Hall, E.A. Lilleskov, L. Liu, J.A. Lynch, K. 

Nadelhoffer, S.S. Perakis, M.J. Robin-Abbott, J.L. Stoddard, K.C. Weathers

19.1 Background

19.1.1 Eff ects of Nitrogen Deposition on Ecosystems

Human activity in the last century has led to a 

substantial increase in nitrogen (N) emissions and 

deposition (Galloway et al. 2003). Because of past, 

and, in some regions, continuing increases in emissions 

(Lehmann et al. 2005, Nilles and Conley 2001), this N 

deposition has reached a level that has caused or is likely 

to cause alterations and damage in many ecosystems 

across the United States. In some ecoregions, the 

impact of N deposition has been severe and has changed 

the biotic community structure and composition of 

ecosystems. In the Mediterranean California ecoregion, 

for example (see Chapter 13), replacement of native 

by exotic invasive vegetation is accelerated because 

exotic species are often more productive under elevated 

N deposition than native species in some California 

grasslands, coastal sage scrub, and desert scrub (Fenn et 

al. 2010, Rao and Allen 2010, Rao et al. 2010, Weiss 

1999, Yoshida and Allen 2004). Such shifts in plant 

community composition and species richness can have 

consequences beyond changes in ecosystem structure: 

shifts may lead to overall losses in biodiversity and 

further impair particular threatened or endangered 

species (Stevens et al. 2004). Th e extirpation of the 

endangered checkerspot butterfl y (Euphydryas editha 

bayensis), because the host plant for the larval stage 

disappears in N-enriched ecosystems (Fenn et al. 2010, 

Weiss 1999), is just one example of the detrimental 

impacts of elevated N deposition.

In addition to altering ecosystem structure, N 

deposition can also aff ect ecosystem function, aff ecting 

N-cycle processes such as N mineralization, nitrifi cation 

rates, and nitrate (NO3

-

) leaching rates, as well as plant 

tissue N concentration. Th ese changes indicate early 

stages of N saturation. Nitrogen saturation is the series 

of ecosystem changes that occur as available N exceeds 

plant and microbial demand (Aber et al. 1989, 1998). In 

some cases, these early responses may lead to a cascade 

of alterations in the N cycle that ultimately aff ect the 

function or structure of the ecosystem (Galloway et 

al. 2003). For example, elevated N inputs may lead to 

plant nutrient imbalances, which then increase plant 

susceptibility to inciting stressors such as cold, drought, 

or pests (Bobbink et al. 1998, Schaberg et al. 2002). 

Th is series of responses was observed in a southern 

Vermont montane red spruce (Picea rubens) stand, 

where increased foliar N concentration was associated 

with reductions in foliar membrane-associated calcium 

(Ca) and decreased cold tolerance, which resulted 

in increased winter injury (Schaberg et al. 2002). 

Another example of the N cascade (Galloway et al. 

2003) is increased soil NO3

-

 leaching, which can result 

in episodic acidifi cation of surface waters, harming 

fi sh species (Baker et al. 1996). Other responses to 

low levels of elevated N deposition, such as increased 

plant growth and increased carbon (C) sequestration 

by trees (Th omas et al. 2010), may be perceived as 

benefi cial where forests are managed for tree growth. 

In other instances, it is not known whether the early 

indicators of N saturation will be followed by other 

eff ects. In these cases, the perceived extent of harm 

caused by N deposition depends, in part, on which 

ecosystem service is of particular value for diff erent 

stakeholders. For example, the level or type of change 

or harm that is unacceptable may vary according to 

resource management goals. In a conservation area, for 

example, any alteration in N cycling may be considered 

unacceptable, whereas for other land areas, changes 

of a certain magnitude or scope may be acceptable or 

even desirable based on resource use (such as timber 

harvesting) or other factors. Land and resources may 

be valued for a wide range of purposes, including 

biodiversity, food and wood production, clean water, 

and recreation. Quantifi cation and then valuation of 

these ecosystem services for each land area of interest is 

required to fully account for impacts of N deposition.
230 Chapter 19—Synthesis GTR-NRS-80

19.1.2 Critical Loads Defi nition and Previous Uses

One method for evaluating the potential impacts of air 

pollution on ecosystems is the critical loads approach. 

Th e critical load is defi ned as the level of a pollutant 

below which no detrimental ecological eff ect occurs over the 

long term according to present knowledge (UBA 2004). 

Critical loads have been used most broadly in Europe 

(Posch et al. 1995, 2001) as a tool in the process of 

negotiating decreases in air pollution. Critical loads 

have been more widely applied in Canada than in the 

United States; critical loads have been published for 

upland forests (Ouimet et al. 2006) and lakes (Dupont 

et al. 2005) in eastern Canada and included in European 

assessments (Hettelingh et al. 2008). In the United 

States, critical loads have been calculated for specifi c 

regions such as the northeastern United States (Dupont 

et al. 2005, NEG/ECP 2003), California (Fenn et al. 

2008, 2010), and Colorado (Baron 2006, Bowman et al. 

2006, Williams and Tonnessen 2000), and, at a coarse 

scale, the conterminous United States (McNulty et al. 

2007). Critical loads are of interest to policy makers 

for assessing emission control programs and to natural 

resource managers as a tool to evaluate the potential 

impact of new pollution sources (Burns et al. 2008, 

Environment Canada 2008, Lovett and Tear 2008, 

Lovett et al. 2009, Porter et al. 2005, US EPA 2007, 

US EPA 2008). Critical loads are also used by policy 

makers and resource managers to establish benchmarks 

for resource protection and to communicate the 

impacts of deposition on natural resource conditions. 

Th e development and use of critical loads provides 

a framework in which the research community 

collaborates with natural resource managers to quantify 

the eff ects of air pollution on ecosystems and help 

guide emission control programs. As a result, scientifi c 

progress, international collaboration, and interactions 

between researchers and policy makers are enhanced. 

Likewise, as a result of this focus on developing critical 

loads, the results of scientifi c studies are more broadly 

applied toward ecosystem protection by land managers, 

policy makers, and regulators.

Critical loads have been determined most frequently in 

the United States for eff ects of acidity (NEG/ECP 2003, 

Sullivan et al. 2005), but are also being increasingly 

used in evaluating impacts of excess N deposition 

on ecosystems (Fenn et al. 2008, 2010 ). Empirical 

critical loads are determined from observations of 

detrimental responses of an ecosystem or ecosystem 

component to a given, observed N deposition input 

(Pardo 2010). Th is level of N deposition is set as the 

critical load and extrapolated to similar ecosystems. 

Empirical critical loads for N, which are set based on 

fi eld evidence, have been used in Europe since the 

1990s (Bobbink et al. 1992, 2003, 2010). Empirical 

critical loads are particularly valuable because they are 

based on actual observations of detrimental eff ects to 

ecosystems by N deposition. Other approaches include 

the steady-state mass balance method (UBA 2004) and 

dynamic modeling (de Vries et al. 2010, Slootweg et 

al. 2007), both of which have been used broadly in 

Europe. Steady-state mass balance modeling is based on 

estimating the net loss or accumulation of N inputs and 

outputs over the long term under the assumption that 

the ecosystem is at steady-state with respect to N inputs. 

Dynamic models also use a mass balance approach, but 

consider time-dependent processes and require detailed 

data sets for parameterization and testing (Belyazid et al. 

2006, de Vries et al. 2007).

Exceedance of the critical load is defi ned as the current 

deposition minus critical load; when exceedance is 

greater than zero, the ecosystem is susceptible to 

harmful ecological eff ects. Th e exceedance is useful in 

communicating the extent of risk to ecosystems under 

current and future deposition scenarios. Th e target load 

is a level of deposition set by policy makers to protect 

sensitive ecosystem components. Th e target load can 

be set below the critical load in order to eliminate 

exceedance within a given time period.

19.1.3 Objectives

Th e objective of this project is to synthesize current 

research relating atmospheric N deposition to eff ects on 

terrestrial and aquatic ecosystems in the United States 

and to identify empirical critical loads for atmospheric 

N deposition where possible. In this fi nal chapter, 

we summarize the critical loads reported for all the 

ecoregions of the United States, discuss the abiotic and 

biotic factors that aff ect the critical load for N within 
Chapter 19—Synthesis GTR-NRS-80 231

each ecoregion, compare critical loads by life form or 

ecosystem compartment (mycorrhizal fungi, lichens, 

herbaceous species, and trees/forest ecosystems) across 

all ecoregions, and compare critical loads in the United 

States to those for similar ecoregions or ecosystems in 

Europe. Finally, we discuss the signifi cance of these 

fi ndings and the highest priorities for future research.

19.2 Approach

For this assessment, we report responses to N inputs for 

ecoregions that occur in the United States based on the 

Commission for Environmental Cooperation (CEC) 

for North America Level I map of ecoregions for North 

America (CEC 1997; Figures 2.1 and 2.2; see Chapter 

2). Th e approach we used was to identify the receptor 

of concern (organism or ecosystem compartment), the 

response of concern, the critical threshold value for 

that response, which studies could be utilized, and the 

criteria for setting the critical load and extrapolating it 

to other sites or regions. Th ese methods are described in 

detail in Chapter 4.

Th e receptors that we evaluated included freshwater 

diatoms, mycorrhizal fungi and other soil microbes, 

lichens, bryophytes, herbaceous plants, shrubs, and 

trees. We also considered biogeochemical processes 

within the ecosystem. Th e main responses reported 

fell into two categories: (1) biogeochemical; and 

(2) individual species, population, and community 

responses. Biogeochemical responses included increased 

N mineralization and nitrifi cation (and N availability 

for plant and microbial uptake), increased gaseous N 

losses (ammonia (NH3)) volatilization, nitric oxide 

(NO) and nitrous oxide (N2

O) from nitrifi cation and 

denitrifi cation), and increased N leaching. Individual 

species, population, and community responses included 

increased tissue N concentration, physiological and 

nutrient imbalances, altered growth, altered root:shoot 

ratios, increased susceptibility to secondary stresses, 

altered fi re regimes, changes in species abundance, 

shifts in competitive interactions and community 

composition, and changes in species richness and other 

measures of biodiversity, and increases in invasive 

species.

We considered experimental N-addition studies, 

N-deposition gradient studies, and long-term 

monitoring studies in order to evaluate ecosystem 

response to N-deposition inputs. Most of these studies 

were not designed to quantify critical loads, which 

presented some challenges. We aff orded greater weight 

to long-term fertilization studies (5 to 10 years) than 

to short-term studies, although short-term studies were 

also considered when other observations were scarce. 

Single-dose forest fertilization studies exceeding 50 kg 

N ha-1 were generally not considered. When studies 

were designed to determine critical loads, the addition 

levels generally depicted modest increases above ambient 

deposition, and were more likely to have smaller 

increments between the treatment levels, multiple (three 

or more) treatment levels, and, ideally, treatments 

spanning the critical load. In such cases, our estimates 

of the critical load are made with greater certainty 

than with other approaches. Nitrogen gradient studies 

implicitly include long-term exposure to pollutants and 

therefore are more likely than N manipulation studies to 

depict conditions that are near steady-state with respect 

to ambient N inputs. Long-term monitoring studies 

sometimes off er the opportunity to observe changes over 

time in response to increasing or elevated N-deposition 

inputs. We estimated critical loads based on data from 

>3200 sites (Figure 3.1).

Th e critical threshold of the response parameter is the 

threshold value at which an acceptable response is still 

observed. For example, when lichens are the receptors 

of concern and the response variable is thallus N 

concentration, the critical threshold is the highest value 

of thallus N concentration that represents a desirable 

condition: the critical threshold is reported as 1 percent 

for a recent study (Fenn et al. in 2008). Th e critical 

threshold is also referred to as the critical limit (UBA 

2004). Th e critical threshold is diff erent from the critical 

load; in this case, the critical load is the deposition at 

which the lichen thallus N concentration has reached or 

exceeded the critical threshold of 1 percent.

In general, we determined the critical load based on 

the observed response pattern. In some cases, there was 

a clear dose-response relationship where the response 
232 Chapter 19—Synthesis GTR-NRS-80

changed above a certain threshold. In other cases, 

when response to increasing N was more linear, we 

estimated the “pristine” state and the deposition level 

that corresponded to a departure from that state. Th e 

criteria for setting critical loads are discussed in detail in 

Chapter 4.

19.3 Deposition

Total N emissions have increased substantially since 

the 1950s (Galloway 1998, Galloway et al. 2003). As 

S deposition has declined in response to U.S. Federal 

emission control programs, the magnitude of N relative 

to S deposition has increased since the 1980s (Driscoll 

et al. 2003). More recently, the relative proportion of 

NHx

 (ammonium (NH4

+

) and ammonia (NH3)) to 

NOx

 (nitric oxide (NO) and nitrogen dioxide (NO2)) 

emissions has also increased for many areas of the 

United States (Kelly et al. 2005, Lehmann et al. 2005). 

In order to quantify the critical load, we generally used 

the deposition reported in the publication or, when that 

was not available, we used modeled deposition (e.g., 

Community Multiscale Air Quality [CMAQ] model, 

ClimCalc [Ollinger et al. 1993], National Atmospheric 

Deposition Program [NADP; NADP 2009] maps). 

Th e diff erent forms of N deposition included in this 

assessment for estimating critical loads are: wet, bulk, 

wet plus dry, throughfall, and total (wet plus dry plus 

cloud/fog) inorganic N deposition. Total N deposition 

is considered the most appropriate value to use in 

evaluating ecosystem responses, however, in many 

studies this information is not available. Th roughfall 

N is generally considered a good surrogate for total N 

deposition, because it typically does not underestimate 

total N inputs as much as wet or bulk deposition 

and it is a good estimate of N delivered to the forest 

fl oor (Weathers et al. 2006). However, because of 

the potential for canopy uptake and transformation 

of N, throughfall is usually considered as a lowerbound estimate of total N deposition. None of the 

studies include reported inputs of organic N, so this 

report focuses on responses to inputs of inorganic N. 

Deposition used to calculate exceedance at sites included 

in this analysis (Chapter 3) was quantifi ed by the 

CMAQ model v.4.3 (hereafter CMAQ 2001 model, 

which uses 2001 reported data) (Byun and Ching 1999, 

Byun and Schere 2006) simulations of wet plus dry 

deposition of N species (Figure 3.1).

We rarely had suffi cient data to distinguish plant or 

ecosystem response to reduced forms (NHx

) versus 

oxidized forms (NOy

: NO, NO2, nitric acid (HNO3), 

organic and inorganic nitrates) of N. Th ere is some 

evidence that for some species, reduced forms of N 

may have more substantial impacts than oxidized 

N (Bobbink et al. 2003, Kleijn et al. 2008). Th is 

diff erential response may be due to direct toxicity of 

gaseous NH3 (Krupa 2003) or the toxicity to some plant 

species (or their mycorrhizal fungi) of high levels of 

NH4

+

 in soil, but can also be a result of soil acidifi cation 

(van den Berg et al. 2005). Lichens in California’s 

Central Valley (Jovan and McCune 2005) have been 

shown to be particularly sensitive to total reduced N 

(i.e. NH4

+

 plus NH3). Across Europe, lichen responded 

to NH3 and to a lesser extent NH4

+

 (Cape et al. 2009, 

Sutton et al. 2009). Much of the research on NH3

eff ects evaluates the response to concentration of NH3, 

which would be used for determining the critical level of 

NH3 rather than the critical load. Th is is an important 

distinction: the critical level is the atmospheric 

concentration above which adverse eff ects to sensitive 

vegetation may occur (UBA 2004). Diff erences in 

uptake rates and preference for NH4

+

 versus NO3

-

 across 

diff erent plant taxa (Falkengren-Grerup 1995, McKane 

et al. 2002, Miller and Bowman 2002, Nordin et al. 

2006) lead to diff erences in sensitivity to NHx

 (Krupa 

2003) and NOy

. Importantly, not all species are more 

sensitive to NHx

 than NOy: these responses vary by 

species and functional type. Some species are more 

sensitive to increases in NOy

, as was demonstrated for 

boreal forests by Nordin et al. (2006).

Th e accuracy of the atmospheric N-deposition values 

used directly infl uences the accuracy of critical load 

and exceedance estimates. Several factors contribute to 

uncertainty in N-deposition estimates, including sparse 

data for many ecosystem types, including arid, high 

elevation sites, and for sites with high inputs from snow 

or cloudwater/fog deposition, where N deposition tends 

to be underestimated. In addition, models of deposition 
Chapter 19—Synthesis GTR-NRS-80 233

often assume homogenous canopies or terrain, or the 

output (e.g., CMAQ) is at a spatial scale (grid size) too 

coarse to capture complex topography and other local 

infl uences on deposition (Weathers et al. 2006). Th ese 

issues are discussed in detail in Chapter 3. When more 

accurate and precise N-deposition estimates become 

available, the data presented in this study may be reevaluated to refi ne the critical loads estimates.

Note that CMAQ deposition data at a 36 km x 36 

km grid were used to calculate exceedances (current 

deposition—critical load) for this analysis. CMAQ 

deposition includes dry deposition of NH3 and trace 

NOy

 species which are not included in many estimates 

of deposition (including those used to estimate many 

of the N critical loads in this assessment). Although 

these N constituents generally make up a small fraction 

of total estimated N deposition, the use of CMAQ 

data has the potential to overestimate the area of 

exceedance. However, some studies suggest CMAQ 

actually underestimates N deposition (Fenn et al. 2010). 

Th e discrepancy between CMAQ estimates and actual 

deposition would be of greatest concern at sites where 

NH3, NO, and NO2 dry deposition or fog represent a 

high fraction of inputs. Th eir inclusion would be most 

signifi cant where the critical load is lowest. Fortunately, 

most of the lichen critical loads, which are typically 

the lowest reported within a given ecoregions (i.e., the 

most likely to be aff ected by the slightly higher CMAQ 

deposition), were determined based on CMAQ inputs. 

In other instances, for example in the arid West when 

emissions are high, CMAQ may underestimate total 

deposition (Fenn et al. 2010).

19.4 Sources of Uncertainty in 

Empirical Critical Loads Estimates

Th ere are several other sources of uncertainty in our 

assessment of empirical critical loads, beyond those 

associated with the measurement of atmospheric 

deposition. In general, there is a dearth of observations 

on ecosystem response to inputs near the critical load. 

To addesss these data gaps, we suggest priorities for 

future research below. In some ecoregions, a single 

study or very few studies are available. If the variability 

of ecosystem response to N deposition across an 

ecoregion is not depicted by the studies available, the 

estimated critical load for N may be relevant for only a 

single ecosystem type or a single sub-region within the 

ecoregion. Without extensive data, it is not possible to 

know whether a study site is more or less sensitive than 

other sites in the ecoregion. It is most eff ective to have 

a large number of studies which demonstrate the range 

of responses observed to better defi ne the threshold 

value (or constrain the deposition range over which the 

response occurs).

Other sources of uncertainty include time lags in the 

response to N deposition and the eff ects of multiple 

stressors, both of which are artifacts of the empirical 

approach, and, as such, are diffi cult to address. 

However, with more long-term studies and more 

response data, confi dence in these empirical critical 

load estimates will improve, as has been demonstrated 

in Europe (Bobbink et al. 1992, 2010). Because 

ecosystems do not respond instantaneously to changes 

in N inputs, there are inherent time lags associated 

especially with N addition studies. Th ese time lags 

become more important with increasing lifespan 

or size of organism; a tree will respond more slowly 

than an herbaceous annual, for example. Time lags 

also depend on the rate of N input, with lower rates 

of input typically leading to longer time lags before 

an initial response (Clark and Tilman 2008). Some 

species adapted to low nutrient supply also tend to 

respond slowly to N additions (Th eodose and Bowman 

1997). (Note that, although the time lag may be 

longer for these low-N-adapted species, they may still 

be amongst the most sensitive to small N additions). 

Th ere may be large diff erences in responses among 

species within the same ecosystem, with many species 

being relatively nonresponsive and a few opportunistic 

species transforming community structure and function. 

Ecosystems with inherently large N pools or capacity 

to absorb N will exhibit longer time lags with respect 

to changes in N inputs than ecosystems with smaller N 

storage pools. Th us, it can be diffi cult to extrapolate the 

response at a higher dose over a short study to a lower 

input over the long term (Clark and Tilman 2008). Th e 

absence of low N addition studies further complicates 

interpolation because often N additions are far greater 
234 Chapter 19—Synthesis GTR-NRS-80

than the critical load. For both N addition and gradient 

studies, the reference plot or low end of the deposition 

gradient may already have been altered from a “pristine” 

condition. Even for gradient studies, the time lag in 

response to N deposition must be considered. As the 

ecosystems in gradient studies are typically still being 

exposed to elevated N inputs, they are continuously 

responding to those N inputs, although they have had 

more time to approach steady-state with the N inputs 

compared with N addition studies. For N gradient 

studies, it can also be diffi cult to sort out the eff ects of 

other factors that may also vary along the deposition 

gradient, such as climate, interannual variation in 

weather, soils, vegetation, disturbances, and other 

pollutants. On the other hand, because these variations 

represent “real-world” conditions—in most locations, 

multiple stressors co-occur—the critical loads estimated 

in the presence of these stressors may better protect the 

ecosystems under the current conditions (Fenn et al. 

2008).

Another source of uncertainty in empirical relationships 

is that they are simply fi eld observations of responses 

to N inputs, and are often lacking mechanistic 

explanations. Th ey also can only refl ect research done 

to date, and with the exception of N addition studies, 

only conditions (N deposition rates) observed to date. 

Because these factors aff ect the accuracy of the critical 

load, we indicate the level of uncertainty in our critical 

loads tables.

19.5 Advantages of the Empirical 

Critical Loads Approach

In spite of some of the challenges discussed above, an 

important advantage of empirical critical loads is that 

they are based on measurable, physical evidence of 

ecosystem responses to N inputs. Conceptually, steadystate mass balance models have an advantage over 

empirical critical loads in terms of estimating long-term 

sustainability, because they are calculated over the long 

term. Th is means that steady-state models are less likely 

to overestimate the critical load, which can happen with 

empirical critical loads determined based on a rapidly 

occurring ecosystem response at a given deposition, 

although lower levels of atmospheric deposition over 

a longer time period will actually generate the same 

response. Currently in the United States, the uncertainty 

associated with the steady-state mass balance method is 

high because data are not available to refi ne the terms in 

the equations. In fact, the data assembled for empirical 

critical loads may be useful in defi ning the acceptable 

critical thresholds used in steady-state mass balance 

critical loads calculations. For example, provisional 

descriptions of the relationships between soil solution 

NO3

-

 concentration and changes in species composition 

in Th e Netherlands (Posch et al. 1993, de Vries et al. 

2007), have allowed determination of the critical NO3

-

concentration term used in steady-state calculations 

of critical loads. Dynamic models for critical loads of 

N in the United States have been applied on a limited 

basis (Fenn et al. 2008, Wu and Driscoll 2010). For 

dynamic modeling of nutrient N critical loads, empirical 

critical loads and other response data are essential: the 

current understanding of ecosystem response to N 

deposition in the United States has not been suffi cient 

to develop dynamic models that characterize the range 

of eff ects (for example, changes in biodiversity) such as 

those utilized in Europe (de Vries et al. 2010, Emmett 

and Reynolds 2003). Dynamic models must be based 

on a systematic understanding of the responses and 

mechanisms for those responses. Dynamic models are 

necessary to adequately characterize the complexity of N 

cycling at the ecosystem scale, but the models can only 

be as good as the data upon which they are based. Th us, 

empirical critical loads currently provide a uniquely 

valuable approach for assessing the risk of harm to 

ecosystems in the United States. Th is report represents a 

fi rst step toward that understanding by indicating which 

data are available for key ecosystems and where dynamic 

modeling could most profi tably be applied in the United 

States after further data collection.

19.6 Overview of Critical Loads across 

U.S. Ecoregions

Th e range of critical loads for nutrient N reported 

for the U.S. ecoregions, inland surface waters, and 

freshwater wetlands is 1 to 39 kg N ha-1 yr-1, while 

coastal wetlands are between 50 to 400 kg N ha-1 yr-1

(Table 19.1). Th is range, excluding coastal wetlands, 

spans N deposition observed over most of the 
Chapter 19—Synthesis GTR-NRS-80 235

country (see Chapter 3). Because N deposition varies 

considerably by region and the critical load varies both 

by region and receptor, we present the critical loads and 

likely risk of exceedance by receptor in section 19.8. 

Th e locations for which ecosystem response data were 

available (Figure 3.1) also vary in density of spatial 

distribution, which impacts the level of certainty of 

the empirical critical loads estimates. Th e basis for the 

critical loads values (Table 19.1) is discussed in detail in 

the preceding chapters.

Th e empirical critical loads for N tend to increase in the 

following sequence for diff erent life forms: freshwater 

diatoms, lichens and bryophytes, mycorrhizal fungi, 

herbaceous plants and shrubs, and trees (see section 

19.8). Low biomass ecosystem types (e.g., grasslands, 

coastal sage scrub, desert) are more sensitive to 

N-enhanced growth of invasive species (if invasive 

pressure occurs), leading to vegetation-type change. 

Th ese vegetation types sometimes occur because 

of warm and dry climatic conditions. As warmer 

temperatures often correspond to greater metabolic 

rates, longer periods of biological activity, greater 

biomass, and more rapid N cycling, one might expect 

that the critical load would increase with increasing 

temperature as has been suggested in Europe (Bobbink 

et al. 2003). We do not observe such a pattern across 

U.S. ecoregions in the critical loads reported in this 

study, but Europe does not have warm and dry 

deserts with low critical loads as in the United States. 

Note, however, that the reliability of the critical load 

estimates varies and is often fairly low, which may 

make it diffi cult to discern patterns in critical load 

values across regions. Moreover, a temperature pattern 

may be confounded by gradients in deposition form 

and quantity, moisture, and elevation. Critical loads 

seem to vary more by receptor and response type than 

by region. Th e western portion of the United States 

has generally similar critical loads values to the eastern 

United States for the same response for a given receptor. 

Th e apparent exception is that in forests the critical 

load for NO3

-

 leaching is approximately twice as high in 

Mediterranean California mixed conifers compared to 

northeastern forests (see section 19.8 and Figure 19.7). 

In contrast, the critical load for NO3

-

 leaching in high 

elevation catchments in the Colorado Front Range is 

lower than anywhere else in the United States, likely 

attributable to low biological N retention and storage 

capacity in these steep, rocky catchments (Baron et al. 

2000, Fenn et al. 2003a, Sickman et al. 2002, Williams 

and Tonnessen 2000).

In this synthesis, we found that higher N critical loads 

were often reported for regions with higher ambient 

N deposition, as has been observed outside the United 

States. One explanation for this pattern is that for 

ecosystems experiencing elevated N deposition, the 

current condition already represents a change from the 

condition prior to elevated N deposition (i.e., a pristine 

or near-pristine state). Th is pattern would explain why 

the empirical critical load is often above the ambient 

deposition even as that deposition increases in the same 

ecosystem type across a region. Empirical critical loads 

for N in Europe (see section 19.9) tend to be higher 

than those for the United States; in China, they are 

higher still (Duan 2009). Th is pattern suggests that 

sometimes the initial change in the ecosystem was not 

captured in ecosystem measurements, thus the critical 

load only prevents further change from the current 

state. Th is is even more likely to have occurred at 

sites in Europe where deposition has been very high. 

For example, European critical loads for lichens were 

infl uenced by a study in Scotland based on a deposition 

gradient from 10 to 22 kg N ha-1 yr-1 which set critical 

loads at 11 to 18 kg N ha-1 yr-1 (Mitchell et al. 2005). 

However, the species composition at all sites across the 

deposition gradient did not include any oligotrophic 

species, which were presumably present prior to elevated 

N deposition inputs. Th is critical load, which is higher 

than critical loads for lichens in the United States, may 

simply prevent further change from an already altered 

state. Similarly, in the Great Plains (see Chapter 11), 

it is not possible to determine whether the current 

condition of sites where deposition is lowest diff ers 

from the pristine condition (Clark and Tilman 2008). 

Th is further emphasizes the need to include “pristine” 

sites in gradient studies and for research experiments 

that remove N deposition in order to more accurately 

defi ne the baseline condition, which helps describe the 

ecosystem state prior to elevated N deposition.
236 Chapter 19—Synthesis GTR-NRS-80 Chapter Ecoregion Ecosystem Component Critical load for N deposition kg N ha-1 yr-1 Reliability Response Comments Study 5 Tundra Prostrate dwarf shrubs 1-3 ## Changes in CO2 exchange, cover, foliar N, and community composition of vascular plants N addition study, Greenland high arctic, P enhanced N effects. Arens et al. 2008a 5 Tundra Lichens 1-3 (#) Changes in lichen pigment production and ultrastructure, changes in lichen and bryophyte cover N addition studies, high and low arctic, P enhanced or moderated N effects. Arens et al. 2008a, Hyvärinen et al. 2003b, Makonen et al. 2007b 6 Taiga Lichen, moss, and algae 1-3 # Changes in alga, bryophyte, and lichen community composition, cover, tissue N or growth rates. Application of western OR and WA model for lichens Berryman et al. 2004c, Berryman and Straker 2008c, Geiser et al. 2010, Moore et al. 2004c, Poikolainen et al. 1998b, Strengbom et al. 2003d, Vitt et al. 2003c 6 Taiga Mycorrhizal fungi, spruce-fi r forests 5-7 (#) Ectomycorrhizal fungi, change in community structure Expert judgment extrapolated from Marine West coast spruce and northern spruce-fi r forest Lilleskov 1999; Lilleskov et al. 2001, 2002, 2008 6 Taiga Shrublands 6 ## Shrub and grass cover, increased parasitism of shrubs Long term, low N addition study: shrub cover decreased, grass cover increased Nordin et al. 2005d, Strengbom et al. 2003d 7 Northern Forests Hardwood and coniferous forests >3 # Tree growth and mortality Decreased growth of red pine, and decreased survivorship 

of yellow birch, scarlet and chestnut oak, quaking aspen, and basswood

Thomas et al. 2010

7 Northern 

Forests

Lichens 4-6 (#) Epiphytic lichen community 

change

Loss of oligotrophic species. Synergistic/confounding effects of acidic deposition not considered; assumes response threshold similar to Marine West Coast Forests

Geiser et al. 2010

7 Northern 

Forests

Ectomycorhizzal fungi 5-7 # Change in fungal community 

structure

Lilleskov et al. 2008

7 Northern 

Forests

Herbaceous cover species

>7 and <21 # Loss of prominent species Response observed in low-level 

fertilization experiment

Hurd et al. 1998

7 Northern 

Forests

Hardwood and coniferous forests

8 ## Increased surface water NO3-

leaching

Aber et al. 2003

Table 19.1—Summary of empirical critical loads of nutrient N for U.S. ecoregions. Reliability rating: ## reliable; # fairly reliable; (#) expert judgment
Chapter 19—Synthesis GTR-NRS-80 237

7 Northern 

Forests

Old-growth montane red spruce

>10 and <26 # Decreased growth and/or 

induced mortality

Response observed in low-level fertilization experiment

McNulty et al. 2005

7 Northern 

Forests

Arbuscular mycorrhizal fungi

<12 (#) Biomass decline and 

community composition change

van Diepen 2008, van Diepen et al. 2007

8 Northwestern 

Forested Mountains

Alpine lakes 1.5 ## Diatom assemblages As wet deposition only Baron 2006

8 Northwestern 

Forested Mountains

Lichens 1.2-3.7 (#) Epiphytic lichen community 

change in mixed-conifer forests, Alaska

Application of western Oregon and Washington model

Geiser et al. 2010

8 Northwestern 

Forested Mountains

Lichens 2.5-7.1 ## Epiphytic lichen community 

change, thallus N enrichment in mixed-conifer forests, nonAlaska

Fenn et al. 2008, Geiser et al. 2010

8 Northwestern 

Forested Mountains

Subalpine forest 4 ## Increase in organic horizon 

N, foliar N, potential net N mineralization, and soil solution N, initial increases in N leaching below the organic layer

Rueth and Baron 2002, Baron et al. 1994

8 Northwestern 

Forested Mountains

Alpine lakes 4.0 # Episodic freshwater 

acidifi cation

Williams and Tonnesson 2000

8 Northwestern 

Forested Mountains

Alpine grassland 4-10 ## Plant species composition Bowman et al. 2006

8 Northwestern 

Forested Mountains

Ectomycorrhizal fungi 5-10 (#) Ectomycorrhizal fungi 

community structure in white, black, and Engelmann spruce forests

Expert judgment extrapolated from Marine West Coast spruce and northern spruce-fi r forest

Lilleskov 1999; Lilleskov et al. 2001, 2002, 2008

8 Northwestern 

Forested Mountains

Mixed conifer forest 17 ##

#

NO3

- leaching, 

reduced fi ne root biomass

Fine root biomass in ponderosa pine is reduced by both ozone and elevated soil nitrogen

Fenn et al. 2008

9 Marine West 

Coast Forests

Western OR and WA forests

2.7-9.2 ## Epiphytic lichen community 

change

Loss of oligotrophic species, enhancement of eutrophic species. Critical load increases with regional range in mean annual precipitation from 45-450 cm

Geiser et al. 2010

continued
238 Chapter 19—Synthesis GTR-NRS-80 9 Marine West Coast Forests SE Alaska forests 5 (#) Fungal community change; declines in ectomycorrhizal fungal diversity Lilleskov 1999; Lilleskov et al. 2001, 2002; Whytemare et al. 1997 10 Eastern Temperate Forests Eastern hardwood forest >3 # Decreased growth of red pine, and decreased survivorship of yellow birch, scarlet and chestnut oak, quaking aspen, and basswood Thomas et al. 2010 10 Eastern Temperate Forests Lichens 4-8 (#) Epiphytic lichen community change Loss of oligotrophic species. Synergistic/confounding effects of acidic deposition not considered. Application of western OR and WA model Geiser et al. 2010 10 Eastern Temperate Forests Southeastern coastal plain 5-10 (#) Ectomycorrhizal fungi community response Dighton et al. 2004; Lilleskov et al. 2001, 2002, 2008 10 Eastern Temperate Forests Eastern hardwood forests 8 ## Increased surface water loading of NO3- Aber et al. 2003 10 Eastern Temperate 

Forests

Michigan deposition gradient

<12 (#) Arbuscular mycorrhizal 

fungal biomass decline and community composition change

van Diepen 2008, van Diepen et al. 2007

10 Eastern 

Temperate Forests

Herbaceous species <17.5 (#) Increases in nitrophilic species, 

declines in species-rich genera (e.g., Viola)

Gilliam 2006,2007; Gilliam et al. 2006

11 Great Plains Tallgrass prairie 5-15 # Biogeochemical N cycling, 

plant and insect community shifts

Clark et al. 2009, Clark and Tilman 2008; Tilman 1993, 1987; Wedin and Tilman 1996 

11 Great Plains Mixed-grass prairie 10-25 # Soil NO3- pools, leaching, plant 

community shifts

Clark et al. 2003, 2005; Jorgenson et al. 2005

11 Great Plains Shortgrass prairie 10-25 (#) Inferred from mixed-grass 

prairie

Epstein 2001, Barret and Burke 2002

11 Great Plains Mycorrhizal fungi 12 (#) Decline in arbuscular 

mycorrhizal fungal activity

Egerton-Warburtone

Table 19.1—continuedChapter Ecoregion Ecosystem Component Critical load for 

N deposition 

kg N ha-1 yr-1

Reliability Response Comments Study
Chapter 19—Synthesis GTR-NRS-80 239

12 North 

American Deserts

Lichens 3 (#) Lichen community shifts, 

thallus N concentration

Uncertainty regarding modeled estimates

Geiser et al. 2008, Porter 2007

12 North 

American Deserts

Shrubland, woodland, desert grassland

3-8.4 # Vegetation response, vascular 

plant community change

Allen et al. 2009; Inouye 2006; Rao et al. 2010 

13 Mediterranean 

California

Coastal sage scrub 7.8-10 # Invasive grass cover, native 

forb richness, arbuscular mycorrhizal fungi richness

Modeled and inferential N deposition estimates and published data for mycorrhizae, unpublished data for vegetation survey

Allenf, EgertonWarburton and Allen 

2000, Fenn et al. 2010, Tonnesen et al. 2007

13 Mediterranean 

California

Chaparral; lichens 3-6 # Epiphytic lichen community 

change

Lichen critical load is from modeled N deposition data and published data for lichens

Fenn et al. 2010, Geiser et al. 2010, Jovan 2008, Jovan and McCune 2005

13 Mediterranean 

California

Chaparral, oak woodlands, Central Valley

10-14 # NO3

- leaching; stimulated N 

cycling

Critical load for NO3

-leaching 

of 10 kg N ha-1 yr-1 is based 

on one year of throughfall data in Chamise Creek and an additional year of throughfall data from adjacent Ash Mountain, both in Sequoia National Park 

Fenn et al. 2003a, 2003b, 2003c; Fenn and Poth 1999; Meixner and Fenn 2004

13 Mediterranean 

California

Mixed conifer forest; lichens

3.1-5.2 ## Lichen chemistry and 

community changes

The lowest critical load is based on lichen tissue chemistry above the clean site threshold

Fenn et al. 2008, 2010

13 Mediterranean 

California

Mixed conifer forest; plant physiology

17 # Reduced fi ne root biomass Fine root biomass in ponderosa 

pine is reduced by both ozone and elevated soil nitrogen

Fenn et al. 2008, 2010; Grulke et al. 1998

13 Mediterranean 

California

Mixed conifer forest; soil processes

17-25.9 # NO3

- leaching; soil acidifi cation Fine root biomass in ponderosa 

pine is reduced by both ozone and elevated soil nitrogen

Breiner et al. 2007; Fenn et al. 2008, 2010

13 Mediterranean 

California

Mixed conifer forest; forest sustainability

24-39 (#) Understory biodiversity; forest 

sustainability

N deposition from Fenn et al. 2008

Allen et al. 2007; Grulke and Balduman 1999; Grulke et al. 1998, 2009; Jones et al. 2004

13 Mediterranean 

California

Serpentine grassland 6 ## Annual grass invasion, 

replacing native herbs

Critical load based on a local roadside gradient; serpentine grassland site is actually west of the Central Valley

Fenn et al. 2010, Weiss 1999

continued
240 Chapter 19—Synthesis GTR-NRS-80 15 Temperate Sierras Lichens 4-7 (#) Epiphytic lichen community change Increase in proportion of eutrophic species. Application of western OR and WA model, response threshold allows ~60% eutrophs due to dry, hot climate, hardwood infl uence Geiser et al. 2010 15 Temperate Sierras Las Cruces and Chichinautzin Ranges S/SW of Mexico City 15 # Elevated NO3- in stream and spring waters Data are from Mexican mountain pine (Pinus hartwegii) sites in the Desierto de los Leones National Park and Ajusco, Mexico Fenn et al. 1999, 2002 16 Tropical and Subtropical Humid Forests N-rich forests <5-10 (#) NO3- leaching, N trace gas emissions Critical load for N-rich forests should be lower than for N-poor forests based on possibility of N losses No direct studiesg 16 Tropical and Subtropical Humid Forests N-poor forests 5-10 (#) Changes in community composition; NO3- leaching, N trace gas emissions Critical load for N-poor forests based on estimates for Southeastern Coastal Plain forests No direct studiesg 17 Wetlands Freshwater wetlands 2.7-13 # Peat accumulation and net primary productivity (NPP) Critical load for wetlands in the northeastern U.S. and southeastern Canada Aldous 2002c, Moore et al. 2004c, Rochefort et al. 1990c, Vitt et al. 2003c

17 Wetlands Freshwater wetlands 6.8-14 (#) Pitcher plant community 

change

Critical load based on northeastern populations

Gotelli and Ellison 2002, 2006

17 Wetlands Intertidal wetlands 50-100 ## Loss of eelgrass Latimer and Rego 201017 Wetlands Intertidal salt marshes 63-400 (#) Salt marsh community 

structure, microbial activity and biogeochemistry

Caffrey et al. 2007, Wigand et al. 2003 

18 Freshwaters Western Lakes 2 ## Freshwater eutrophication Baron 200618 Freshwaters Eastern Lakes 8 # NO3- leaching Aber et al. 2003

Table 19.1—continued

Chapter Ecoregion Ecosystem Component Critical load for 

N deposition 

kg N ha-1 yr-1

Reliability Response Comments Study

a based on data from Greenland; b based on data from Finland; c based on data from Canada; d based on data from Sweden, e see footnote 25 on page 19-11; f Allen, E.B. Unpublished data. 

Professor and Natural Resources Extension Specialist, Department of Botany and Plant Sciences and Center for Conservation Biology, University of California, Riverside, CA 92521; g The 

critical load is based on expert judgment and knowledge of ecosystems which may function similarly.
Chapter 19—Synthesis GTR-NRS-80 241

One would expect that for an ecosystem that is 

nearer N saturation than another, it would take less 

additional N to reach a “tipping point”. However, it 

is important not to confuse that issue with the actual 

level of deposition—an ecosystem may be near N 

saturation when the ambient N deposition is low or 

when it is high. Th us, while prior exposure to elevated N 

deposition does push an ecosystem toward N saturation, 

high ambient deposition does not indicate the 

ecosystems most sensitive to further inputs—that status 

is a function of the ecosystem and the receptor.

19.7 Factors that Affect the Critical 

Load

One of the objectives of this assessment was to lay 

the groundwork for further refi ning and improving 

estimates of N critical loads. To that end, in this section, 

we discuss some of the factors that aff ect where the value 

of the critical load falls within the reported range (Table 

19.2). We present the factors that were reported in the 

studies included in this report. Th ese factors may be 

useful in setting empirical critical loads for specifi c sites 

as well as in the development of more complex models 

to assess ecosystem response to N inputs.

Abiotic factors that may aff ect the N critical load 

include elevation, latitude, topographic location, climate 

(temperature, precipitation, extent and rate of climate 

change), catchment size, soil type, extent of soil cover in 

high elevation systems, parent material, and hydrologic 

fl owpaths and processes. Disturbance may also play a 

substantial role, for example, the type of fi re regime or 

historical forest cutting can impact the critical load. A 

geographical region within the ecoregion may be more 

sensitive. Biological factors also likely contribute to 

lower N critical loads, including particularly sensitive 

species (diatoms, lichens, mycorrhizal fungi, certain 

plants), single species versus community responses, low 

biomass and low productivity ecosystems, short lifespan 

of receptor of concern, presence of invasive grasses, and 

presence of ozone-sensitive species.

Th e factors discussed above may provide some guidance 

in applying critical loads. To set a critical load for a 

given site using this report, one would fi rst determine 

whether the site was similar to the site/or sites on which 

the critical load for that ecosystem type is based (found 

in the tables presented in each ecoregion chapter). If 

the site diff ers from the sites in the ecoregions tables, 

one would then refer to Table 19.2 to determine how 

to adjust the N critical load for a given site based on 

the range reported for the ecoregion. Finally, one 

would consider the general factors discussed above and 

adjust the critical load within the range reported for the 

ecoregion based on these factors.

With better identifi cation of factors that aff ect the 

N critical load, we will move toward a mechanistic 

understanding of the responses and improve our ability 

to extrapolate observations across ecoregions or across 

diff erent ecosystems within an ecoregion. In some cases, 

it may be possible to develop simple relationships as a 

function of one or several variables that would allow 

refi nement of critical load estimates. For example, for 

lichens, Geiser et al. (2010) developed simple regression 

relationships that included precipitation volume that 

explain much of the variability in lichen community 

composition in response to N deposition. Th ese 

regression models can be used to estimate N critical 

loads in other regions and also can provide an estimate 

of the uncertainty associated with the critical load.

Future research could evaluate which of these factors 

are most important in aff ecting where the critical load 

lies within the range for an ecoregion and determining 

how the critical load varies as the key parameters change 

across the ecoregions.

19.8 Comparison of Critical Load by 

Receptor across Ecoregions

Because much of the variation in critical loads occurs as 

a function of the receptor of concern, we compare the 

critical loads for the key receptors across ecoregions.

19.8.1 Mycorrhizal fungi

Background. Mycorrhizal fungi reside at the interface 

between host plants and soils, exchanging soil resources, 

especially nutrients, with host plants in exchange for 

photosynthates (carbon compounds). Due to this 

important and unique ecological niche, mycorrhizal 
242 Chapter 19—Synthesis GTR-NRS-80 Table 19.2—Assessment and interpretation of empirical critical loads of nutrient N for U.S. ecoregions Chapter Ecoregion Factors affecting the range of critical loadsa Comparison within Ecoregionb 5 Tundra moisture competition between vascular plants and cryptogams P-limitation temperature pH The critical load is higher in wet and P-limited tundra; acidic tundra may be more sensitive to N deposition than non-acidic tundra. Increased N deposition may be more detrimental to lichens in the presence of graminoids and shrubs in the low and mid arctic than to lichens with less competition in the high arctic. Response time increases with latitude due to colder temperatures, less light, and poorer N and P mobilization. 6 Taiga soil depth vegetation type and species composition latitude Morphological damage to lichens has been observed at a lower deposition in forests and woodlands than in shrublands or bogs and fens; cryptogam dominated mats on thin soils become N saturated faster than forest islands. 7 Northern Forests receptor tree species stand age site history pre-existing N status Critical loads for lichens are generally lowest, followed by critical loads for ectomycorrhizal fungi and NO3- leaching. Critical loads for herbaceous species and forests are generally higher than for other responses. 8 Northwestern Forested Mountains biotic receptor accumulated load of N ecosystem region In alpine regions, diatom changes in lakes are seen at the lowest critical load. Changes in individual plants are seen next, followed by vegetation community change, then soil responses. In subalpine forests, the critical load of 4 kg ha-1 yr-1 for foliar and soil chemistry changes is similar to the lichen critical load of 3.1 - 5.2 for lichen community change. 9 Marine West 

Coast Forests

background N statussoil typespecies compositionfi re historyclimate

The midrange of responses reported for lichens (2.7 - 9.2 kg ha-1

yr-1) is broadly comparable to that for plant, soil, and mycorrhizal 

responses (5 kg ha-1 yr-1), despite limited studies for non-lichen 

responses.

10 Eastern 

Forests

precipitationsoil cation fertility and weatheringbiotic receptors

The critical load for NO3- leaching, lichen community change, 

and ectomycorrhizal fungal response are within the same range. Arbuscular mycorrhizal fungal and herbaceous critical loads are higher.

11 Great Plains N status

receptorprecipitation

Critical loads are lower in the tall grass prairie than in the mixedand short-grass prairies. Critical loads in tall- and mixed-grass prairie is lower on N poor sites and sites with very N responsive plant species. Critical loads in the short-grass prairie is likely lower in wet years than in dry years.
Chapter 19—Synthesis GTR-NRS-80 243

aThis explains what factors cause the critical load (CL) to be at the low or high end of the range reported.bComparison of values and causes for differences if multiple critical loads are reported for an ecoregion.12 North 

American Deserts

receptorinteraction of annual grasses with native forb coverprecipitation

The lichen critical load is lowest, at 3 kg N ha-1 yr-1; vegetation 

critical load varies from 3 to 8.4 kg N ha-1 yr-1

13 Mediterranean 

California

presence of invasive exotic annual grasses interacting with a highly diverse native forb community N-sensitivity of mycorrhizal fungiN-sensitivity of lichensN retention capacity of catchments, catchment sizeco-occurrence of ozone and ozone-sensitive tree species.

The lowest critical loads in Mediterranean California are for sensitive lichen in chaparral and oak woodlands and mixed conifer forests. The critical load for plant and mycorrhizal fungal community change in coastal sage scrub is higher, at 7.8 to 10 kg ha-1 yr-1.

Critical load for NO3

- leaching is lower in chaparral and oak 

woodlands (10 t0 14 kg ha-1 yr-1) than in mixed conifer forests 

(17 kg ha-1 yr-1). Critical loads are highest for mixed conifer forest 

plant community change and sustainability.

17 Wetlands vegetation species

fraction of rainfall in the total water budgetdegree of openness of N cycling

Critical load is much higher for intertidal wetlands (50 to 400 kg ha-1 y-1) than for freshwater wetlands (2.7 to 14 kg ha-1 y-1), which 

have relatively closed water and N cycles.
244 Chapter 19—Synthesis GTR-NRS-80

fungi are at particular risk due to changes in either the 

soil environment or host carbon allocation.

Response to N inputs. Nitrogen deposition adversely 

aff ects mycorrhizal fungi primarily in two ways: (1) by 

causing decreased belowground C allocation by hosts 

and increased N uptake and associated metabolic costs 

(Wallander 1995); and (2) via soil chemical changes 

associated with eutrophication and acidifi cation. Th ere 

are two major groups of mycorrhizal fungi that are 

evolutionarily and ecologically distinct: arbuscular 

mycorrhizal fungi (AMF) and ectomycorrhizal 

fungi (EMF). Under suffi ciently high N inputs, the 

progressive eff ect of elevated N is an early decline of 

sporocarp (reproductive structure) production for EMF 

and spore production for AMF, and subsequent decline 

in biological diversity and loss of taxa adapted to N-poor 

environments or that are sensitive to acidifi cation 

(Lilleskov 2005). Sporocarp and spore production 

appears to be especially sensitive to N deposition, often 

declining before the communities on root tips have been 

substantially altered, presumably because sporocarps and 

spores are at the end of the carbon fl ux pathway from 

hosts.

Of the two plant-fungal symbioses examined in this 

report, mycorrhizal fungi appear to be less sensitive 

to N deposition than lichens (Tables 19.3 and 19.4), 

presumably because the soil environment buff ers these 

soil fungi from some of the immediate impacts of N 

deposition, whereas lichens are directly exposed to 

atmospheric N pollution. (Lichens have an advantage 

as indicators when compared with mycorrhizal fungi 

because they can be relatively easily inventoried.) 

However, the essential role of mycorrhizal fungi as root 

symbionts central to plant nutrition and belowground 

production, as repositories of a large part of the 

eukaryote diversity in forests, as major components of 

forest food webs, and as nontimber forest products of 

high economic value (edible sporocarps) (Amaranthus 

1998) provides suffi cient justifi cation to improve our 

understanding of their response to N deposition.

Critical loads of N for mycorrhizal fungi. We reviewed 

empirical studies on mycorrhizal fungal response to N 

inputs as the basis for determining empirical critical 

loads for the United States (Table 19.3, Figure 19.1). 

Despite the sparse data, it is clear that N deposition 

suffi cient to elevate inorganic N, and especially NO3

-

, 

availability in soils can have measurable eff ects on 

mycorrhizal fungi. Th e data for EMF indicate that N 

deposition to N-limited conifer forests in the range of 

5 to10 kg ha-1 yr-1 can signifi cantly alter community 

structure and composition and decrease species richness 

(Dighton et al. 2004; Lilleskov 1999; Lilleskov et al. 

2001, 2002, 2008). Similarly, the data for AMF suggest 

N deposition levels of 7.8 to12 kg ha-1 yr-1 can lead to 

community changes, declines in spore abundance and 

root colonization, and changes in community function, 

based on reanalysis of data from Egerton-Warburton 

et al. (2000) combined with N deposition data and 

decreases in fungal abundance (van Diepen et al. 2007, 

Van Diepen 2008) and declines in fungal activity.24 Th e 

actual threshold for N eff ects on AMF could be even 

lower because high background deposition precludes 

consideration of sites receiving deposition at or near 

pre-industrial levels. Th erefore, our provisional expert 

judgment is that critical loads for mycorrhizal diversity 

for sensitive ecosystem types are 5 to10 kg ha-1 yr-1. 

Th ere is high uncertainty in this estimate because few 

studies have been conducted at low N deposition to 

further refi ne the critical load.

Th e critical load of N for mycorrhizal fungi, when 

community change occurs, is often on the order of 

current N deposition and thus is exceeded across 

much of the Eastern Temperate and Northern Forests 

ecoregions. Exceedance of N critical loads is more 

variable in the western United States. Th e critical 

load is exceeded in regions downwind of agricultural 

and urban emissions in the West (Figure 19.2). In 

the Northwestern Forested Mountains and Marine 

West Coast Forests, N deposition is generally below 

the critical load, although the lower end of the critical 

load range is exceeded in the Cascade Mountains. 

A similar pattern can be seen in Mediterranean 

California; the N critical load for mycorrhizal fungi is 

exceeded in the vicinity of the Sierra Nevada and in the 

24 Egerton-Warburton, L.M. Unpublished data. Chicago 

Botanic Garden, 1000 Lake Cook Road, Glencoe, IL, 60022
Chapter 19—Synthesis GTR-NRS-80 245

Table 19.3—Empirical critical loads of nutrient N for mycorrhizal fungi in U.S. ecoregions. Reliability rating: ## reliable; # fairly reliable; (#) expert judgmentEcoregion Ecosystem (Site) Critical load for 

N deposition 

kg ha-1 yr-1

Reliability Response Comments Study

Taiga Spruce forests 5-7 (#) Ectomycorrhizal fungi, 

change in community structure

Expert judgment extrapolated from Marine West Coast spruce and northern spruce-fi r forest

Lilleskov 1999; Lilleskov et al. 2001, 2002, 2008

Northern Forests

Spruce-fi r forest (Northeastern U.S. deposition gradient)

5-7 # Ectomycorrhizal fungi, 

change in community structure

Wet deposition estimated from Ollinger et al. (1993) model

Lilleskov et al. 2008

Northern Forests

Northern hardwood forests; sugar maple dominated (Michigan gradient)

<12 (#) Arbuscular mycorrhizal fungi, 

decrease in abundance in roots, soil, community change

N fertilization experiment van Diepen et al. 

2007, van Diepen 2008

Northwestern Forested Mountains

Engelmann spruce forests

5-10 (#) Ectomycorrhizal fungi, 

change in community structure

Expert judgment extrapolated from Marine West coast spruce and northern spruce-fi r forest

Lilleskov 1999; Lilleskov et al. 2001, 2002, 2008

Marine West Coast Forests

White spruce forest (Kenai Peninsula, Alaska)

5 (#) Ectomycorrhizal fungi, 

change in community structure, decrease in species richness

Bulk deposition. Historic N deposition was higher but not quantifi ed. Critical load estimated from regression

Lilleskov 1999, Lilleskov et al. 2001, 2002, Whytemare et al. 1997

Eastern Temperate Forests

Southeastern Coastal Plain 

5-10 (#) Ectomycorrhizal fungi, 

change in community structure

From one study in pine barrens plus extrapolation from other oligotrophic conifer forests

Dighton et al. 2004; Lilleskov et al. 2001, 2002, 2008

Eastern Temperate Forests

Pine Barrens (New Jersey; Southeastern Coastal Plain) 

<8 (#) Ectomycorrhizal fungal 

morphotype community change

Bulk deposition. Gradient study with three sample points

Dighton et al. 2004

Eastern Temperate Forests

Eastern hardwoods; sugar maple dominated (Michigan gradient)

<12 (#) Arbuscular mycorrhizal fungi; 

decrease in abundance in roots, soil, community change

Long-term (12 yr) N fertilization experiment in sugar maple

van Diepen et al. 2007, van Diepen 2008 

Great Plains Chicago grassland 12 (#) Arbuscular mycorrhizal fungi; 

decrease in % colonization, spore density

Critical load estimated from logarithmic curve of soil N vs. arbuscular mycorrhizal fungal activity. No low N baseline, so critical load may be lower

Egerton-Warburtona

Mediterranean California

Coastal sage scrub (southern California)

7.8-9.2 # Arbuscular mycorrhizal fungi, 

decrease in % colonization, spore density, spore richness

Critical load estimated from logarithmic curve fi tted to data from this study compared to modeled and inferential N deposition data

Egerton-Warburton and Allen 2000, Fenn et al. 2010, Tonnesen et al. 2007

asee footnote 24 on page 244
246 Chapter 19—Synthesis GTR-NRS-80

Empirical CL of N (kg ha-1 yr

-1)

 5 Marine West Coast Forests

 5 - 7 Northern Forests; Taiga

 5 - 10 Northwest Forested Mountains

 5 - 12 Eastern Temperate Forests

7.8 - 9.2 Mediterranean California

 12 Great Plains

Uncertainty

Reliable

Fairly Reliable

Expert Judgment

Figure 19.1—Map of critical loads (CL) for mycorrhizal fungi by ecoregion in the United States. The range of 

critical loads reported for mycorrhizal fungi is shown for each ecoregion. The hatch marks indicate increasing 

level of uncertainty: no hatch marks for the most certain “reliable” category, single hatching for the “fairly 

reliable” category, and double hatching for the “expert judgment” category. The color sequence moves from red 

toward blue and violet as the critical load increases. As the range of the critical load gets broader, the saturation 

of the color decreases. White areas lack data for critical loads determination for mycorrhizal fungi.
Chapter 19—Synthesis GTR-NRS-80 247

Exceedance of Critical Loads of N

Below CL min

At CL min

Above CL min

Above CL max

Uncertainty

Reliable

Fairly Reliable

Expert Judgement

Figure 19.2—Map of exceedance of critical loads (CL) for mycorrhizal fungi by ecoregion in the continental United States. 

Exceedance was calculated by subtracting critical loads from CMAQ nitrogen deposition. Exceedance (critical load - 

deposition) is shown for several categories: (1) No exceedance (Below CL), when deposition is lower than the CL range, 

(2) At CL, when deposition is within +/-1 of the CL range, (3) Above CLmin, when deposition is above the lower end of the 

CL range, but lower than the upper end of the range, (4) Above CLmax, when deposition is above the upper end of the CL 

range. CMAQ deposition data were not available for Alaska, so we were not able to calculate exceedance for Alaska. 

White areas lack data for critical loads determination for mycorrhizal fungi.

Transverse Mountain ranges of southern California. Th e 

uncertainty associated with the exceedance, like that for 

the critical load, is high.

19.8.2 Lichens and Bryophytes

Background. Lichens and bryophytes make substantial 

contributions to biodiversity. About 4100 lichen and 

2300 bryophyte species are known from North America 

north of Mexico—as about one-fourth of vascular plant 

diversity, which is about 26,600 species (NRCS 2009). 

Th erefore, N critical loads protective of the sensitive 

lichens and bryophytes help protect biological diversity. 

Lichens are symbiotic organisms consisting of a fungus, 

for which the organism is named, and a green algal 

and/or a blue-green bacterial partner. Bryophytes are 

small, thin-leaved, nonvascular plants encompassing 

the mosses, liverworts, and hornworts. Neither lichens 

nor bryophytes have true roots or other specialized 

conductive tissues. Individual species are adapted 

to specifi c nutrient availability regimes, therefore 

oligotrophic environments will be characterized by 

diff erent species than eutrophic environments.
248 Chapter 19—Synthesis GTR-NRS-80

Responses to N inputs. Lichens and bryophytes are among 

the most sensitive bioindicators of N in terrestrial 

ecosystems (Blett et al. 2003, Bobbink et al. 2003, Fenn 

et al. 2003a, Glavich and Geiser 2008). Unlike vascular 

plants, lichens and bryophytes lack specialized tissues 

to mediate the entry or loss of water and gases (e.g., 

waxy epidermis, guard cells, root steele). Th us, they 

rapidly hydrate and absorb gases, water, and dissolved 

nutrients during high humidity or precipitation events. 

However, they quickly dehydrate to a metabolically 

inactive state as well, making them slow-growing and 

vulnerable to contaminant accumulation. Consequently, 

the implementation of lichen or bryophyte-derived 

critical loads may prevent undesired impacts to 

much of the broader forest ecosystem (McCune et al. 

2007). In some cases, alteration of lichen community 

composition may signal the beginning of a cascade of 

changes in ecosystem N cycling, which may markedly 

alter the structure or function of the ecosystem as a 

whole. In many cases, changes in lichens may have 

implications in portions of the ecosystem beyond 

the lichen community. In other cases, alterations in 

the lichen community may have little impact on the 

overall structure and function of the ecosystem. It can 

be diffi cult to know at the outset whether the ultimate 

consequences of changes indicated by alterations to the 

lichen community will be large or small for the overall 

ecosystem over the long term.

Lichens and bryophytes can play important roles in 

ecosystems. Species of epiphytic lichens in wet and 

mesic forests that are most sensitive to N (i.e., the large 

pendant and foliose species) play important ecological 

roles that are not duplicated by the eutrophic (i.e., 

nitrogen tolerant) species that may replace them. 

Dominant regional oligotrophs (e.g., Alectoria, Bryoria, 

Lobaria, Ramalina, Usnea) comprise the bulk of lichen 

biomass in old-growth forests, contribute to nutrient 

cycling through N2

 fi xation, and are used for nesting 

material, essential winter forage for rodents and 

ungulates, and invertebrate habitat (McCune and Geiser 

2009). Storage of water and atmospheric nutrients by 

these lichen genera and epiphytic bryophytes moderates 

humidity and provides a slow release system of essential 

plant nutrients to the soil (Boonpragob et al. 1989, 

Cornelissen et al. 2007, Knops et al. 1991, Pypker 

2004). In the tundra, lichens and bryophytes represent a 

signifi cant portion of the biomass, and reindeer lichens 

are a vital link in the short arctic food chain (Kytöviita 

and Crittenden 2007). Mosses comprise the bulk of the 

biomass of the extensive boreal peatlands. In the desert, 

lichens and bryophytes, together with other microbiota, 

form cryptogamic mats important to soil stabilization 

and fertility.

A hypothetical example of sensitive species diminishing 

to the extent where they cannot fulfi ll their ecological 

roles for the northern fl ying squirrel of Pacifi c 

Northwest forests follows. Over 90 percent of the 

squirrel’s diet consists of hypogeous and epipgeous 

mycorrhizal fungi in summer and the N-sensitive 

horsehair lichen (Bryoria fremontii) in winter. Th e 

squirrel is both an important dispersal agent for 

mycorrhizal fungi (obligate symbionts with the roots of 

conifer trees) and a primary prey of the northern spotted 

owl (Strix occidentalis), a threatened and endangered 

species (Maser et al. 1985). If N deposition extirpated 

the horsehair lichen, as it nearly has in the Netherlands 

(van Herk et al. 2003), all dependent species would 

presumably decline regardless of whether they are 

themselves sensitive to N deposition.

Critical loads of N for lichens. Th e N critical loads 

estimated in this report for lichens range from 1 to 9 

kg N ha-1 yr-1 (Table 19.4, Figure 19.3). Although the 

reported range of N critical loads is not as large as the 

ranges for forests or herbs, the certainty associated with 

these estimates varies considerably. Th is is partially 

because of diff erences in sampling scheme and intensity. 

For example in the Pacifi c Northwest and California, 

lichen communities were assessed intensively across 

wide environmental gradients spanning low to high 

N deposition on a fi ne grid over time (Geiser and 

Neitlich 2007, Jovan 2008), yielding highly reliable 

critical N load estimates. Assessments in the eastern 

United States are more problematic, due to historical 

and contemporary S emissions and acid deposition. 

In such cases, where historical information necessary 

to identify a “pristine” or “clean” state is lacking, it is 

more diffi cult to determine the N critical load, and the 

resulting confi dence associated with the critical load is 

low. Th e critical load of N for lichens, based on the shift 
Chapter 19—Synthesis GTR-NRS-80 249

in community composition when eutrophs dominate 

at the expense of oligotrophs, is on the order of current 

N deposition and thus is exceeded across much of the 

Eastern Temperate and Northern Forest ecoregions and 

in many areas (e.g., high deposition) in the West (Figure 

19.4). Th e uncertainty associated with the exceedance, 

like that for the critical load, is low for the Marine 

West Coast and Northwestern Forested Mountains 

ecoregions and Mediterranean California forests, but 

high elsewhere.

Studies in the Pacifi c Northwest demonstrate that 

increasing precipitation allows lichens to tolerate higher 

N deposition (Geiser and Neitlich 2007, Geiser et al. 

2010, Chapter 4). Th e importance of precipitation 

volume in the N critical load for lichens is likely due to 

the direct infl uence of N concentration on lichens, that 

is, the concentrations of N compounds to which lichens 

are exposed are more important than total N loading 

(Geiser et al. 2010). If such simple models could be 

tested and confi rmed in other regions of the country, the 

confi dence in the critical loads in those regions would 

improve.

19.8.3 Herbaceous Species and Shrubs

Background. Herbaceous species and shrubs (Table 19.5, 

Figure 19.5) are found in grasslands, shrublands, forests, 

deserts, and wetlands, and comprise the majority of the 

roughly 26,600 vascular plant species found in North 

America north of Mexico (NRCS 2009).

Response to N inputs. Herbaceous species and some 

shrubs appear intermediate between cryptogram and 

tree species in their sensitivity to N deposition, due 

to specialized tissues that mediate the entry or loss of 

water and gases compared with cryptograms, and rapid 

growth rates, shallow rooting systems, and often shorter 

lifespan compared with trees. Th us, herbaceous species 

in a forest understory will likely respond more rapidly 

to changes in N deposition and to a greater degree 

than the trees with which they coexist. Herbaceous 

species in alpine or tundra environments will respond 

later and to a lesser degree than the cryptograms with 

which they coexist. Herbaceous plants obviously play 

an important role in those ecosystems in which they 

are the dominant primary producers (e.g., grasslands, 

shrublands). In forests, however, the role of the 

herbaceous community in ecosystem function has an 

importance that is disproportionate to its relatively low 

biomass. For example, although they represent only 

approximately 0.2 percent of standing aboveground 

biomass, herbaceous understory species produce more 

than15 percent of forest litter biomass and comprise 

up to 90 percent of forest plant biodiversity, including 

endangered or threatened species (Gilliam 2007).

Critical loads of N for herbaceous vegetation. Th e range 

of critical loads for N for herbaceous species and shrubs 

across all ecoregions is 3 to 33 kg N ha-1 yr-1 (Table 

19.5, Figure 19.5). Although this range is broader than 

those for lichens or mycorrhizal fungi, many of the 

critical loads for herbaceous species fall into the range 

of 5 to15 kg N ha-1 yr-1. Th ere is moderate uncertainty 

in these estimates. Th e shorter lifespan of some 

herbaceous species results in a more rapid response to N 

addition. Th is pattern is especially relevant for annuals 

and perennials with little N storage. In grasslands, for 

example, elevated N deposition often leads to a rapid 

(1 to 10 years) increase in herbaceous production and a 

shift in biomass allocation towards more aboveground 

tissue. Th is often decreases light levels at ground 

surface and decreases the numbers of plant species, 

primarily of perennials, legumes, and natives (Clark 

and Tilman 2008, Suding et al. 2004, Tilman 1993). 

Experimental studies of moderate to long duration (3 

to10 years) allow determination of the N critical load 

with reasonable certainty. Longer studies (>10 years) 

would decrease the uncertainty further. In some cases, 

it can be diffi cult to determine whether the condition 

in reference plots or at the low end of a deposition 

gradient represents a “pristine” condition or whether a 

site has already been altered by N deposition prior to 

or at the time of the study. For example, the watershed 

acidifi cation study at Fernow Experimental Forest, 

West Virginia, Adams et al. (2006) added 35 kg N 

ha-1 yr-1 via aerial N application in addition to ambient 

deposition of 15 to 20 kg N ha-1 yr-1, which has led to 

changes in understory species composition. Recently, 

similar changes in understory species composition have 

occurred on the adjacent reference watershed receiving 
250 Chapter 19—Synthesis GTR-NRS-80 Table 19.4—Empirical critical loads of nutrient N for lichens in U.S. ecoregions. Reliability rating: ## reliable; # fairly reliable; (#) expert judgment Ecoregion Ecosystem (Site) Critical load for N deposition kg ha-1 yr-1 Reliability Response of Comments Study Tundra Tundra 1-3 (#) Changes in lichen pigment production and ultrastructure, changes in lichen and bryophyte cover N addition studies, high and low Arctic, P enhanced or moderated N effects Arens et al. 2008a, Hyvärinen et al. 2003b, Makkonen et al. 2007b Taiga Taiga 1-3 # Changes in alga, bryophyte, and lichen community composition, cover, tissue N or growth rates Berryman et al. 2004c, Berryman and Straker 2008c, Geiser et al. 2010, Moore et al. 2004c, Poikolainen et al. 1998b, Strengbom et al. 2003d, Vitt et al. 2003c Northern Forests Northern Forests 4-6 (#) Changes in lichen physiology and community structure Application of western OR and WA model Geiser et al. 2010 Northwest Forested Mountains Coniferous forests, Alaska 1.2-3.7 (#) Lichen community composition change Application of western OR and WA model Geiser et al. 2010 Northwest Forested Mountains Coniferous Forests, non- Alaska 2.5-7.1 ## Lichen community composition change Application of western OR and WA model Geiser et al. 2010

Northwest Forested Mountains

Central Southern Sierras

3.1-5.2 ## Shifts in epiphytic lichen 

communities favoring eutrophs

Critical load based on exceedance of a N concentration threshold in the lichen Letharia vulpina

Fenn et al. 2008, 2010

Marine West Coast Forests

Western OR and WA forests

2.7-9.2 ## Shifts in epiphytic lichen 

communities favoring eutrophs

Critical load increases with increasing mean annual precipitation from 40 to 240 cm

Geiser et al. 2010

Eastern Forests

Eastern hardwoods and Southeast Coastal Plain

4-84-6

(#)(#)

Shifts in epiphytic lichen communities favoring eutrophs

Application of western OR and WA model

Geiser et al. 2010

North American Deserts 

Cold desert (Hells Canyon National Resource Area)

3 (#) Increased cover and 

abundance of nitrophilous lichens on tall shrubs, increased parasitism of lichens

Critical load estimated from overlay of course grid (36 km) CMAQ N; local N deposition from NH3 was likely higher

Geiser et al. 2008, Porter 2007
Chapter 19—Synthesis GTR-NRS-80 251

a based on data from Greenland; b based on data from Finland; c based on data from Canada; d based on data from SwedenMediterranean California 

Oak woodlands and chaparral (Central Valley: Sacramento Valley, Coast Ranges and Sierra foothills)

3-6 # Shifts in epiphytic lichen 

communities favoring eutrophs

Forest Health Monitoring (FHM) lichen survey of 118 forested sites. Eutrophs dominated communities when CMAQ 4 km N deposition estimates were >5.5 kg ha-1 yr-1

Fenn et al. 2010, Geiser et al. 2010, Jovan 2008, Jovan and McCune 2005 

Mediterranean California

Mixed Conifer forest (Sierra Nevada)

3.1-5.2 ## Shifts in epiphytic lichen 

communities favoring eutrophs

Extrapolated from Northwestern Forested Mountains Sierra Nevada study

Fenn et al. 2008, 2010

Temperate Sierras

Lichens 4-7 (#) Shifts in epiphytic lichen 

communities favoring eutrophs

Application of western OR and WA model

Geiser et al. 2010
252 Chapter 19—Synthesis GTR-NRS-80

Empirical CL of N (kg ha-1 yr-1)

 1 - 3 Tundra; Taiga

 1.2 - 3.7 Northwest Forested Mountains, Alaska

 2.5 - 7.1 Northwest Forested Mountains, non- Alaska

 2.7 - 9.2 Marine West Coast Forests

 3 North American Deserts

 3.1 - 6 Mediterranean California

 4 - 6 Northern Forests

 4 - 7 Temperate Sierras

 4 - 8 Eastern Temperate Forests

Uncertainty

Reliable

Fairly Reliable

Expert Judgment

Figure 19.3—Map of critical loads (CL) for lichens by ecoregion in the United States. The range of critical 

loads reported for lichens is shown for each ecoregion. The hatch marks indicate increasing level of 

uncertainty: no hatch marks for the most certain “reliable” category, single hatching for the “fairly reliable” 

category, and double hatching for the “expert judgment” category. The color sequence moves from red 

toward blue and violet as the critical load increases. As the range of the critical load gets broader, the 

saturation of the color decreases. White areas lack data for critical loads determination for lichens.
Chapter 19—Synthesis GTR-NRS-80 253

only ambient atmospheric deposition (Gilliam et 

al.1996).25 Th is pattern suggests that N deposition to 

the reference watershed currently exceeds the critical 

load. It is diffi cult to determine the empirical N critical 

load at sites where ambient deposition exceeds the 

critical load. Where deposition rates exceed the critical 

load, empirical measurement of the rate of change of 

an ecological metric (e.g., plant abundance, diversity, 

or community composition) over a range of N inputs 

provides an approach to estimate the N level at which 

that metric begins to change (Bowman et al. 2006).

Th e critical load of N for herbaceous species and herbs, 

when community change occurs (in some cases with 

exotic invasives replacing native species), is exceeded 

across much of the Great Plains ecoregion, in portions 

of the Southwest, and in high elevation and high 

deposition areas of the other ecoregions (Figure 19.6). 

Th e uncertainty associated with the exceedance, like that 

for the critical load, varies.

Exceedance of Critical Loads of N

Below CL min

At CL min

Above CL min

Above CL max

Uncertainty

Reliable

Fairly Reliable

Expert Judgement

Figure 19.4—Map of exceedance of critical loads (CL) for lichen by ecoregion in the continental United States. Exceedance 

was calculated by subtracting critical loads from CMAQ nitrogen deposition. Exceedance (critical load - deposition) is 

shown for several categories: (1) No exceedance (Below CL), when deposition is lower than the CL range, (2) At CL, when 

deposition is within +/-1 of the CL range, (3) Above CLmin, when deposition is above the lower end of the CL range, but 

lower than the upper end of the range, (4) Above CLmax, when deposition is above the upper end of the CL range. CMAQ 

deposition data were not available for Alaska, so we were not able to calculate exceedance for Alaska. White areas lack 

data for critical loads determination for lichens.

25Gilliam, F.S. Unpublished data. Professor, Department of 

Biological Sciences, Marshall University, Huntington, WV 

25755-2510.
254 Chapter 19—Synthesis GTR-NRS-80 Table 19.5—Empirical critical loads of nutrient N for herbaceous plants and shrubs in U.S. ecoregions. Reliability rating: ## reliable; # fairly reliable; (#) expert judgment Ecoregion Ecosystem (Site) Critical load for N kg ha-1 yr-1 Reliability Response Comments Study Tundra Prostrate dwarf shrub 1-3 ## Changes in CO2 exchange, cover, foliar N, and community composition of vascular plants N addition study, Greenland high arctic, P enhanced N effects Arens et al. 2008a Taiga Shrublands 6 ## Change in shrub and grass cover, increased parasitism of shrubs Long-term, low N addition study: shrub cover decreased, grass cover increased Nordin et al. 2005b, Strengbom et al 2003b Northern Forests Northern hardwood forests (Adirondacks) >7 and <21 # Alteration of herbaceous understory Hurd et al. 1998 Northwestern Forested Mountains Alpine grasslands 4-10 ## Plant species composition change Based on long-term experiment Bowman et al. 2006 Eastern Forests Eastern hardwood 

forests (Fernow Experimental Forest, WV)

<17.5 (#) Increases in nitrophilic 

species, declines in species-rich genera (e.g., Viola)

Gilliam 2006, Gilliam 2007, Gilliam et al. 2006

Great Plains Tallgrass prairie 5-15 # Biogeochemical N cycling, 

plant and insect community shifts

Long-term, low N addition study that also added other nutrients.

Clark et al. 2009; Clark and Tilman 2008; Tilman 1993, 1987; Wedin and Tilman 1996 

Great Plains Mixed-grass 

prairie

10-25 # Soil NO3- pools, leaching, 

plant community shifts

Short-term, low N addition study

Jorgenson et al. 2005, Clark et al 2003, 2005

Great Plains Shortgrass prairie 10-25 (#) Inferred from mixedgrass prairie Epstein 2001, Barrett and Burke 2002

North American Desert

Warm desert (Joshua Tree National Park, Mojave Desert)

3-8.4 # Increased biomass of 

invasive grasses; decrease of native forbs

Allen et al. 2009; Inouye 2006; Rao et al. 2009, 2010
Chapter 19—Synthesis GTR-NRS-80 255

a based on data from Greenland; b based on data from Sweden; c Allen, E.B. Unpublished data. Professor and Natural Resources Extension Specialist, 

Department of Botany and Plant Sciences and Center for Conservation Biology, University of California, Riverside, CA 92521; d Fenn, M.E. Unpublished data. 

Mediterranean California

Research plant pathologist, Forest Fire Laboratory, 4955 Canyon Crest Dr., Riverside, CA 92507.

Serpentine grassland

6 ## Annual grass invasion, 

replacing native herbs

Critical load based on a local roadside gradient; serpentine grassland site is actually west of the Central Valley

Fenn et al. 2010, Weiss 1999 

Mediterranean California

Coastal sage scrub

7.8-10 # Invasive grass cover, native 

forb richness

Modeled and inferential N deposition estimates and unpublished data for vegetation survey

Allen unpublishedc, 

Egerton-Warburton et al. 2001, Fenn et al. 2010, Tonnesen et al. 2007

Mediterranean California

Mixed conifer forests (San Bernardino Mountains)

24-33 (#) Biodiversity of understory: 

percent cover and no. of species/3 ha

Based on plant surveys in 1970s and 2003

Allen et al. 2007N deposition data: Fenn et al. 2008 and Fennd

Wetlands Freshwater 

wetlands

6.8-14 (#) Pitcher plant community 

change

Critical load based on northeastern populations

Gotelli and Ellison 2002, 2006

Wetlands Intertidal 

wetlands

50-100 ## Loss of eelgrass Latimer and Rego 2010

Wetlands Intertidal salt 

marsh

63-400 (#) Salt marsh community 

structure, microbial activity and biogeochemistry

Caffrey et al. 2007, Wigand et al. 2003 
256 Chapter 19—Synthesis GTR-NRS-80

Empirical CL of N (kg ha-1 yr

-1)

 1 - 3 Tundra

 3 - 8.4 North American Desert

 4 - 10 Northwest Forested Mountains

 5 - 25 Great Plains

 6 Taiga

 6 - 33 Mediterranean California

 >7 - <21 Northern Forests

 <17.5 Eastern Temperate Forests

Uncertainty

Reliable

Fairly Reliable

Expert Judgment

Figure 19.5—Map of critical loads (CL) for herbaceous plants and shrubs by ecoregion in the United 

States. The range of critical loads reported for herbaceous plants and shrubs is shown for each ecoregion. 

The hatch marks indicate increasing level of uncertainty: no hatch marks for the most certain “reliable” 

category, single hatching for the “fairly reliable” category, and double hatching for the “expert judgment” 

category. The color sequence moves from red toward blue and violet as the critical load increases. As 

the range of the critical load gets broader, the saturation of the color decreases. White areas lack data for 

critical loads determination for herbaceous species and shrubs.
Chapter 19—Synthesis GTR-NRS-80 257

Exceedance of Critical Loads of N

Below CL min

At CL min

Above CL min

Above CL max

Uncertainty

Reliable

Fairly Reliable

Expert Judgement

Figure 19.6—Map of exceedance of critical loads (CL) for herbaceous plants and shrubs by ecoregion in the continental 

United States. Exceedance was calculated by subtracting critical loads from CMAQ nitrogen deposition. Exceedance 

(critical load - deposition) is shown for several categories: (1) No exceedance (Below CL), when deposition is lower 

than the CL range, (2) At CL, when deposition is within +/-1 of the CL range, (3) Above CLmin, when deposition is above 

the lower end of the CL range, but lower than the upper end of the range, (4) Above CLmax, when deposition is above 

the upper end of the CL range. CMAQ deposition data were not available for Alaska, so we were not able to calculate 

exceedance for Alaska. White areas lack data for critical loads determination for herbaceous species and shrubs.

Comparisons of N critical load with current deposition 

indicates that the low end of the critical load range for 

herbaceous species is exceeded for most of the Great 

Plains and Mediterranean California ecoregions (Figure 

19.6). Th e low end of the critical load range is also 

exceeded in portions of the Eastern Temperate Forests, 

Northern Forests, Northeastern Forested Mountains, 

and North American Deserts ecoregions; uncertainty for 

exceedance varies. Portions of the Cascade Mountain 

range, the Rocky Mountains, the Adirondacks, and the 

Green and White Mountain ranges exceed the low end 

of the critical load range.

19.8.4 Trees/Forest Ecosystems

Background. In this section we discuss the responses 

of trees and the biogeochemical responses of forest 

ecosystems to N inputs (Table 19.6), excluding the 

specifi c responses of mycorrhizal fungi, lichens, or 

understory herbaceous plants. Forest ecosystems 

represent a third of landcover in the United States 
258 Chapter 19—Synthesis GTR-NRS-80

(USFS 2001) and are signifi cant in the Northern, 

Eastern Temperate, Tropical and Subtropical Humid, 

and Marine West Coast Forests; Northwestern Forest 

Mountains; and Mediterranean California ecoregions.

Response to N inputs. In northeastern forests, gradient 

studies demonstrate that N deposition enhances 

growth in some fast-growing tree species, including 

the hardwoods studied with arbuscular mycorrhizal 

associations, whereas it slows growth in some species (red 

spruce, red maple [Acer rubrum]), and has no detectable 

eff ect on still other species (Th omas et al. 2010). 

Similarly, N deposition enhances survivorship in a few 

species (black cherry [Prunus serotina], red maple, paper 

birch [Betula papyrifera]) and decreases survivorship in 

others (Th omas et al. 2010). Survivorship under chronic 

N deposition, and possibly other co-occurring pollutants 

such as ozone, is often dependent on interactions with 

other stressors such as pests, pathogens, climate change, 

or drought (Grulke et al. 2009, McNulty and Boggs 

2010). Over the long term, these diff erential eff ects of N 

deposition on tree growth and survivorship are likely to 

shift species composition, possibly to more nitrophilic 

species, similar to patterns seen for organisms with 

shorter lifespans.

We have few data that show a major structural or 

functional shift in forest ecosystems because of the long 

response time of trees and forest soils to changes in N 

inputs and N availability (Table 19.6). Th e relatively 

large pools of organic N in the forest fl oor, mineral 

soil, tree biomass, and detritus contribute to the 

relatively long lag time in forest ecosystem response to 

N inputs. Because of the long lag time in response to 

N treatments, it can be diffi cult to determine the actual 

critical N load for forest ecosystems based on shortterm fertilization studies. If a response is observed over 

a relatively short period of time (i.e., years), it is nearly 

certain that the critical load is below the total N input 

at the treatment site and it can be diffi cult to further 

constrain the critical load.

It is expected that the more complex and interconnected 

processes in forests will result in a higher N critical 

load, in part because large N storage pools give forest 

ecosystems a greater capacity to buff er N inputs. In 

herbaceous plants, too, responses in individual species 

tend to be observed at lower N inputs than changes in 

community composition, which are more complex and 

interconnected (Bowman et al. 2006).

Critical loads of N in forests. Th e range of critical loads 

reported for forest ecosystems is 4 to 39 kg N ha-1 yr-1

(Table 19.6, Figure 19.7). Th e threshold N deposition 

value which caused increased NO3

-

 leaching from forest 

ecosystems into surface water was 8 to 17 kg N ha-1 yr-1; 

the lower end of the range represents Northern and 

Eastern Temperate Forests, the upper end represents 

Mediterranean California mixed conifers (Table 19.6, 

Figure 19.7). At 4 kg N ha-1 yr-1, increasing [NO3

-

] was 

reported in the organic horizon in the Colorado Front 

Range, which suggests incipient N saturation (Rueth 

and Baron 2002).Th e highest critical loads were reported 

for Mediterranean California mixed-conifer forests for 

soil acidifi cation caused by increased N deposition and 

for forest sustainability. Th ese sites experience some of 

the highest N deposition reported in the United States, 

up to approximately 70 kg N ha-1 yr-1 (Fenn et al. 2008).

Th e N critical load is exceeded across much of the 

eastern forests (Eastern Temperate and Northern Forest 

ecoregions). Th e lower end of the N critical load range 

is exceeded for the remaining portions of the eastern 

forests, as well as portions of the Marine West Coast, 

Northwestern Forested Mountains, and Tropical and 

Subtropical Humid Forest ecoregions (Figure 19.8). Th e 

N critical load for forest ecosystems was not exceeded 

for much of area of the Mediterranean California 

ecoregion, in part because the critical load was very high 

and, in part, because the CMAQ 2001 deposition that 

we used was at a coarse resolution that underestimates 

deposition compared to the fi ner scale (4 km x 4 km 

grid) used by Fenn et al. (2010). Th e critical load for 

nitrate leaching (Figure 19.9) is exceeded in portions of 

the Mediterranean California and the lower end of the 

critical load range is exceeded for most eastern forest and 

part of the Great Plains (Figure 19.10).

19.8.5 Inland Surface Waters and Wetland 

Ecosystems

Background. Freshwater lakes and streams, and wetlands 

(freshwater and estuarine intertidal) are ecosystem 
Chapter 19—Synthesis GTR-NRS-80 259

Table 19.6—Empirical critical loads of nutrient N for forest ecosystems in U.S. ecoregions. Reliability rating: ## reliable; # fairly reliable; (#) expert judgmentEcoregion Ecosystem (Site) CL for N 

kg ha-1 yr-1

Reliability Response Comments Study

Northern Forests Northeastern 

gradient

>3 # Decline in survivorship of 

sensitive species

Based on study of gradient of N deposition from 3-11 kg N ha-1 yr-1

Thomas et al. 2010

Northern Forests Hardwood and 

coniferous forests

8 ## Increased surface water 

and NO3

- leaching

Aber et al. 2003

Northern Forests Montane spruce fi r 

(Mt. Ascutney, VT)

>10 and 

<26

# Declines in growth and 

increased mortality 

McNulty et al. 2005

Northwestern Forested Mountains

Subalpine forest 4 ## soil organic horizon and 

foliar N enrichment and higher potential net N mineralization rates

Baron et al. 1994, Rueth and Baron 2002

Northwestern Forested Mountains

Mixed conifer forest

17 ##

#

NO3

- leaching, 

reduced fi ne root biomass

Co-occurring ozone also affects fi ne root biomassi in ponderosa pine

Fenn et al. 2008

Marine West Coast Forests

Coastal white spruce forest (South-Central Alaska)

5 (#) Declines in tree health; 

changes in understory composition; foliar nutritient imbalances; elevated NO3

- in forest 

fl oor and mineral soil 

Lilleskov 1999, Lilleskov et al. 2001, 2002; Whytemare et al. 1997

Eastern Forests Eastern hardwood 

forests

>3 # Decline in survivorship of 

sensitive species

Based on study of gradient of N deposition from 3-11 kg N ha-1 yr-1

Thomas et al. 2010

Eastern Forests Eastern Hardwood 

Forests

8 ## Increased surface water 

loading of NO3

-

Aber et al. 2003

Mediterranean California

Mixed conifer forests (San Bernardino Mountains and southern Sierra Nevada range)

17 ## Streamwater [NO3

-] > 14 

μM

Based on regression of throughfall vs. peak streamwater NO3

- concentrations. Daycent 

simulations gave similar results

Fenn et al. 2008, 2010

Mediterranean California

Mixed conifer forests (San Bernardino Mountains)

17 # Reduced fi ne root 

biomass

Based on regression of throughfall N deposition and fi ne root biomass in ponderosa pine (also affected by co-occurring ozone)

Fenn et al. 2008; Grulke et al., 1998

continued
260 Chapter 19—Synthesis GTR-NRS-80

a

Mediterranean California

 The critical load is based on expert judgment and knowledge of ecosystems which may function similarly.

Mixed conifer forests (San Bernardino Mountains)

25.9 # Soil acidifi cation; pH 

≤ 4.6 Based on regression of throughfall 

N deposition and mineral soil H+

Breiner et al. 2007

Mediterranean California

Mixed conifer forests (San Bernardino mountains)

39 (#) Forest sustainability Based on shifts in plant phenology 

and C allocation. Caused by combined effects of ozone and N deposition. Leads to increased bark beetle mortality and wildfi re risk 

Grulke and Balduman 1999; Grulke et al. 1998, 2009; Jones et al. 2004N deposition data from Fenn et al. 2008.

Tropical and Subtropical Humid Forests

N-poor tropical and subtropical forests

5-10 (#) Changes in community 

composition; NO3

-

leaching, N trace gas emissions

Critical load for N-poor forests based on estimates for Southeastern Coastal Plain forests

No direct studiesa

Tropical and Subtropical Humid Forests

N-rich tropical and subtropical forests

<5-10 (#) NO3

- leaching, N trace 

gas emissions

Critical load for N-rich forests should be lower than for N-poor forests based on possibility of N losses

No direct studiesa

Ecoregion Ecosystem (Site) CL for N 

kg ha-1 yr-1

Reliability Response Comments Study

Table 19.6—continued
Chapter 19—Synthesis GTR-NRS-80 261

types that occur in most ecoregions in North America. 

In freshwater lakes and streams, phytoplankton, 

or algae that live in the water column, are sensitive 

to the chemical environment in which they reside. 

Many species can be used as indicators of the levels 

of nutrients or acidity because of individual species’ 

preference for specifi c chemical conditions. Diatoms, 

unicellular plants that form the base of freshwater food 

webs, are used in this discussion because there has 

been more work published on these algae than others, 

but other types of algae also respond to N deposition 

(Lafrancois et al. 2004, Michel et al. 2006). Of the 

wetlands which occur in the conterminous United 

States, 95 percent are freshwater wetlands and 5 percent 

are estuarine or marine wetlands (FWS 2005). Th e 

species composition tends to diff er between freshwater 

and intertidal wetlands, although together they support 

more than 4200 native plant species. Despite the high 

biodiversity, the eff ects of N loading are studied in just a 

few plant species.

Response to N inputs. For the analysis of nutrient N 

eff ects to freshwater lakes and streams, we relied on 

papers and studies that linked aquatic biological and 

ecological response to atmospheric deposition, but the 

results are consistent with laboratory or in situ doseresponse studies and even land-use change studies. 

Th e productivity of minimally disturbed aquatic 

ecosystems is often limited by the availability of N, and 

slight increases in available N trigger a rapid biological 

response that increases productivity and rearranges 

algal species assemblages (Nydick et al. 2004, Saros et 

al. 2005). Th e mechanism for change is alteration of 

nitrogen:phosphorus (N:P) ratios, which can increase 

productivity of some species at the expense of others 

(Elser et al. 2009). As with the terrestrial ecosystems 

described above, freshwater nutrient responses are most 

evident where land use change and acidic deposition 

have been limited, thus most evidence of exceedance 

of N critical loads comes from the western United 

States (Chapter 18). As with terrestrial plants, some 

diatoms respond rapidly to an increase in available N. 

An example is dominance of two diatoms (Asterionella 

formosa and Fragilaria crotonensis) in numerous Rocky 

Mountain lakes with higher N, in contrast with lakes 

with lower N deposition, where there is a more even 

distribution, thus high biodiversity, of diatoms. Higher 

trophic levels (zooplankton, macroinvertebrates) may 

be secondarily aff ected by N, but further increases in 

primary, or autotrophic, production will be limited by 

other nutrients such as P or silica (Si).

Both freshwater and estuarine intertidal wetlands tend to 

be N-limited ecosystems (LeBauer and Tresseder 2008, 

US EPA 1993). Known responses to N enrichment are 

generally derived from nutrient-addition studies in the 

fi eld and observations along gradients of N deposition. 

A variety of ecological endpoints are evaluated, 

such as altered soil biogeochemistry, increased peat 

accumulation, elevated primary production, changes 

in plant morphology, changes in plant population 

dynamics, and altered plant species composition (US 

EPA 2008). In general, the sensitivity of wetland 

ecosystems to N is related to the fraction of rainfall (a 

proxy for atmospheric N deposition) in its total water 

budget. Most freshwater wetlands, such as bogs, fens, 

marshes and swamps, have relatively closed water and 

N cycles, thus are more sensitive to N deposition than 

intertidal wetlands, such as salt marshes, and eelgrass 

(Zostera sp.) beds (Chapter 17).

Critical loads of N for freshwater ecosystems. In general, 

critical loads for freshwater lakes and streams tend to 

be low because the target organisms are unicellular 

algae that respond rapidly to changes in their 

chemical environment. Th e range of critical loads 

for eutrophication and acidity is 2 to 9 kg N ha-1 yr-1

(Chapter 18); the range reported for terrestrial and 

wetland ecosystems is much broader (Table 19.1). 

Critical loads for NO3

-

 leaching from terrestrial 

ecosystems ranged from 4 to 17 kg N ha-1 yr-1 (Figure 

19.9), but many sensitive freshwaters at high altitudes 

are found above the tree line where watershed N 

retention is limited due to little vegetation, poorly 

developed soils, short hydraulic residence time, and 

steep topography. Many hydrological factors including 

hydraulic residence time, N pool size, and conditions 

of water saturation aff ect N loss. Th ese factors infl uence 

how rapidly a system exhibits elevated N leaching 

in response to increased N deposition, and how this 

increased N availability subsequently infl uences biota. 

In general, lakes have relatively rapid N turnover times 
262 Chapter 19—Synthesis GTR-NRS-80

Empirical CL of N (kg ha-1 yr

-1)

 >3 - 8 Eastern Temperate Forests

 >3 - <26 Northern Forests

 4 - 17 Northwest Forested Mountains

 <5 - 10 Tropical and Subtropical Humid Forests

 5 Marine West Coast Forests

 17 - 39 Mediterranean California

Uncertainty

Reliable

Fairly Reliable

Expert Judgment

Figure 19.7—Map of critical loads (CL) for forest ecosystems by ecoregion in the United States. The range 

of critical loads reported for forest ecosystems is shown for each ecoregion; this map does not include the 

responses of mycorrhizal fungi, lichens, or understory herbaceous plants already represented. The hatch marks 

indicate increasing level of uncertainty: no hatch marks for the most certain “reliable” category, single hatching for 

the “fairly reliable” category, and double hatching for the “expert judgment” category. The color sequence moves 

from red toward blue and violet as the critical load increases. As the range of the critical load gets broader, the 

saturation of the color decreases. White areas lack data for critical loads determination for forest ecosystems.
Chapter 19—Synthesis GTR-NRS-80 263

Exceedance of Critical Loads of N

Below CL min

At CL min

Above CL min

Above CL max

Uncertainty

Reliable

Fairly Reliable

Expert Judgement

Figure 19.8—Map of exceedance of critical loads (CL) for forest ecosystems by ecoregion in the continental United 

States. Exceedance was calculated by subtracting critical loads from CMAQ nitrogen deposition. Exceedance (critical 

load – deposition) is shown for several categories: (1) No exceedance (Below CL), when deposition is lower than the 

CL range, (2) At CL, when deposition is within +/-1 of the CL range, (3) Above CLmin, when deposition is above the 

lower end of the CL range, but lower than the upper end of the range, (4) Above CLmax, when deposition is above the 

upper end of the CL range. CMAQ deposition data were not available for Alaska, so we were not able to calculate 

exceedance for Alaska. White areas lack data for critical loads determination for forest ecosystems.

compared to soil N pools and are at least seasonally well 

mixed. Th ey would, thus, be expected to have lower 

critical loads. Turnover times for N in mineral soil 

pools can be very long, and, as a result, buff er changes 

in soil solution that would aff ect terrestrial plants. Th us 

responses by terrestrial plants would not be expected to 

be as rapid as those of freshwater organisms.

Generally the most sensitive type of wetland to N 

deposition are freshwater wetlands, with critical loads 

that range from 2.7 to14 kg N ha-1 yr-1 (Chapter 

17). Th e nonvascular plant genus, Sphagnum, and 

the carnivorous pitcher plant (Sarracenia sp.) are the 

two species most commonly studied. Th e critical 

loads reported for freshwater wetlands (Chapter 

17) fall between those reported for inland surface 

waters (Chapter 18) and those reported for terrestrial 

ecosystems. Th e critical load tends to be higher for 

intertidal wetlands than other types of ecosystems 

because they have open nutrient cycles which are often 

strongly aff ected by N loading sources other than 

atmospheric deposition. Based on fi eld observations of 
264 Chapter 19—Synthesis GTR-NRS-80

Empirical CL of N (kg ha-1 y

-1)

 4 - 17 Northwest Forested Mountains

 8 Northern Forests; Eastern Temperate Forests

10 - 17 Mediterranean California

10 - 25 Great Plains

Uncertainty

Reliable

Fairly Reliable

Expert Judgment

Figure 19.9—Map of critical loads (CL) for NO3- leaching by ecoregion in the United States. The range of critical 

loads based on increased nitrate leaching for each ecoregion. The hatch marks indicate increasing level of 

uncertainty: no hatch marks for the most certain “reliable” category, single hatching for the “fairly reliable” category, 

and double hatching for the “expert judgment” category. The color sequence moves from red toward blue and violet 

as the critical load increases. As the range of the critical load gets broader, the saturation of the color decreases. 

White areas lack data for critical loads determination for nitrate leaching.
Chapter 19—Synthesis GTR-NRS-80 265

Exceedance of Critical Loads of N

Below CL min

At CL min

Above CL min

Above CL max

Uncertainty

Reliable

Fairly Reliable

Expert Judgement

Figure 19.10—Map of exceedance of critical loads (CL) for NO3

-

 leaching by ecoregion in the continental United 

States. Exceedance was calculated by subtracting critical loads from CMAQ nitrogen deposition. Exceedance (critical 

load - deposition) is shown for several categories: (1) No exceedance (Below CL), when deposition is lower than the 

CL range, (2) At CL, when deposition is within +/-1 of the CL range, (3) Above CLmin, when deposition is above the 

lower end of the CL range, but lower than the upper end of the range, (4) Above CLmax, when deposition is above the 

upper end of the CL range. CMAQ deposition data were not available for Alaska, so we were not able to calculate 

exceedance for Alaska. White areas lack data for critical loads determination for nitrate leaching.

N loading to plant growth and species composition on 

salt marsh and eelgrass habitat, the critical load ranges 

between 50 and 400 kg N ha-1 yr-1.

19.9 Comparison to Critical Loads in 

Europe

Th e critical loads for N deposition we report are 

consistently lower than those reported for Europe, with 

a few exceptions (Table 19.7). It is diffi cult to make 

the comparison between the United States and Europe 

because the ecosystem classifi cation systems used are 

not parallel. Empirical N critical loads for Europe 

(Bobbink et al. 2003) were reported at diff erent scales 

using the European University Information Systems 

(EUNIS) than the U.S. critical loads: coarser for forests, 

for example, and fi ner for nonforests. Furthermore, 

the response variables and thresholds values of those 

variables are not always the same.

Th ere are several reasons that N critical loads in Europe 

may be higher than in the United States. First, as 

discussed earlier, because N deposition in Europe has 
266 Chapter 19—Synthesis GTR-NRS-80

Table 19.7—Comparison of critical loads for Europe and the United States. Reliability rating: ## reliable; # fairly reliable; (#) expert. judgment. European data from Bobbink et al. 2003.North American ecoregion

Ecosystem component

Response Critical Load 

kg N ha-1 yr-1

Reliability European 

EUNIS 

codea

European ecosystem type or component

European indication of exceedance

European 

critical load 

kg N ha-1 yr-1

Reliability

Tundra Prostrate dwarf 

shrub

Changes in CO2 exchange, 

cover, foliar N, and community composition of vascular plants

1-3 ## F1 Tundra Changes in biomass, physiological 

effects, changes in species composition in moss layer, and decrease in lichens

5-10 #

Tundra Lichens Changes in lichen 

pigment production and ultrastructure, changes in lichen and bryophyte cover

1-3 (#) F1 Tundra Changes in biomass, physiological 

effects, changes in species composition in moss layer, and decrease in lichens

5-10 #

Taiga Lichens, algae, 

and bryophytes

Changes community composition, abundance, physiology, or ultrastructure.

1-3 # G Temperate and 

boreal forests: lichens and algae

Increase of algae, decrease of lichens

10-15 (#)

Taiga Mycorrhizal fungi Ectomycorrhizal fungal 

change in community structure

5-7 (#) G Temperate and 

boreal forests: mycorrhiza

Reduced sporocarp production, changed/reduced belowground species composition

10-20 (#)

Taiga Shrublands Change in shrub and grass 

cover, increased parasitism of shrubs

6 ## F2 Arctic, alpine and 

subalpine scrub habitats

Decline in lichens, mosses, and evergreen shrubs

5-15 (#)

Northern Forests

Lichens Change in lichen community 

composition

4-6 (#) G Temperate and 

boreal forests: lichens and algae

Increase of algae, decrease of lichens

10-15 (#)

Northern Forests

Mycorrhizal fungi Ectomycorrhizal fungal 

change in community structure

5-7 # G Temperate and 

boreal forests: mycorrhiza

Reduced sporocarp production, changed/reduced belowground species composition

10-20 (#)

Northern Forests

Herbaceous species

Loss of prominent species >7 and <21 # G Temperate and 

boreal forests: ground vegetation

Changed species composition, increase of nitrophilous species, increased susceptibility to parasites

10-15 #

Northern Forests

Coniferous forests 

Increased surface water NO3- leaching

8 ## G Coniferous forests Increased NO3- leaching 10-15 ##

Northern Forests

Hardwood forests Increased surface water 

NO3- leaching

8 ## G Deciduous forests Increased NO3- leaching 10-15 (#)

Northwestern Forested Mountains

Coniferous forests, Alaska: lichens

Lichen community change 1.2-3.7 (#) G Temperate and 

boreal forests: lichens and algae

Increase of algae, decrease of lichens

10-15 (#)

Northwestern Forested Mountains

Coniferous forests, nonAlaska: lichens

Lichen community change 2.5-7.1 ## G Temperate and 

boreal forests: lichens and algae

Increase of algae, decrease of lichens

10-15 (#)

Northwestern Forested Mountains

Subalpine forest Increase in organic horizon 

N; higher potential net N mineralization rates

4 ## G Deciduous and 

coniferous forests

Increased N mineralization and nitrifi cation

10-15 #
Chapter 19—Synthesis GTR-NRS-80 267

Northwestern Forested Mountains

Alpine vegetation: dry meadow tundra

Change in species composition

4-10 ## E4.3 Alpine or subalpine 

grasslands

Increase in tall graminoids, biodiversity change

10-15 (#)

Northwestern Forested Mountains

Mycorrhizal fungi Ectomycorrhizal fungal 

change in community structure

5-10 (#) G Temperate and 

boreal forests: mycorrhiza

Reduced sporocarp production, changed/reduced belowground species composition

10-20 (#)

Marine West Coast Forests

Lichens Epiphytic lichen community 

change

2.7-9.2 ## G Temperate and 

boreal forests: lichens and algae

Increase of algae, decrease of lichens

10-15 (#)

Eastern Forests

Lichens Epiphytic lichen community 

change

4-8 (#) G Temperate and 

boreal forests: lichens and algae

Increase of algae, decrease of lichens

10-15 (#)

Eastern Forests

Southeastern Coastal Plain: mycorrhizal fungi

Ectomycorrhizal fungal community response

5-10 (#) G Temperate and 

boreal forests: mycorrhiza

Reduced sporocarp production, changed/reduced belowground species composition

10-20 (#)

Eastern Forests

Eastern hardwood forests

Increased surface water loading of NO3

-

8 ## G Deciduous forests Increased NO3

- leaching 10-15 (#)

Eastern Forests

Eastern hardwood forest

Increased foliar percent N >9b G Deciduous and 

coniferous forests

Changed N/macro nutrient ratios, decreased P, K, Mg; increased N concentration in foliar tissue

15-20 #

Eastern Forests

Herbaceous species

Herb layer changes <17.5 (#) G Temperate and 

boreal forests: ground vegetation

Changed species composition, increase of nitrophilous species, increased susceptibility to parasites

10-15 #

Great Plains Tallgrass prairie Biogeochemical N cycling, 

plant and insect community shifts

5-15 # E1.7 Non-Mediterranean 

dry acid and neutral closed grassland

Increase in graminoids, decline typical species

10-20 #

Great Plains Mixed-grass 

prairie

Soil NO3

- pools, leaching, 

plant community shifts

10-25 # E1.26 Sub-Atlantic semidry calcareous 

grassland

Increased tall grasses, decline in diversity; increased mineralisation and N leaching

15-25 ##

Wetlands Bogs, fens, and 

swamps

Alterations in sphagnum accumulation and net primary productivity; alteration in pitcher plant community

2.7-14 # D1 Raised and blanket 

bogs

Increase vascular plants, decrease bryophytes, altered growth and species composition of mosses, increased N in peat and peat water

5-10 ##

Wetlands Intertidal salt 

marsh

Salt marsh community structure, microbial activity and biogeochemistry

63-400 (#) A2.64 and 

A2.65

Pioneer and lowmid salt marshes Increase late-successional species, increase productivity 30-40 (#)

Aquatic Western high 

elevation lakes

Eutrophication 2 ## C1.1 Soft-water lakes Isoetid species negatively affected 5-10 ##

Aquatic Eastern high 

elevation lakes

Eutrophication 8 # C1.1 Soft-water lakes Isoetid species negatively affected 5-10 ##

a EUNIS categories: A—Marine habitats; C—Inland surface water habitats; D—Mire, bog, and fen habitats; E—Grasslands and tall forb habitats; F—Heathland, scrub, and tundra habitats; G—Forest habitatsb This value, from Boggs et al. (2005) is discussed in Chapter 10 as a response to N deposition. We did not include this value as a critical load in other tables, as the response threshold for detrimental 

ecological effects was not clear.
268 Chapter 19—Synthesis GTR-NRS-80

been considerably higher than that in the United States 

over many decades, it is diffi cult to fi nd background or 

unimpacted sites as a basis of comparison. Th is means 

that, in some cases, the observed change in species 

composition does not refl ect the initial response of 

a community to increased N inputs, but rather the 

response of a previously impacted community to further 

increases in N deposition. In addition, NH4

+

 inputs 

tend to be higher and represent a greater proportion 

of total N inputs in Europe. Note also that when dry 

deposition is underestimated in the United States, the N 

critical loads will also be underestimated, which would 

contribute to them being lower than those in Europe. 

Finally, since a greater proportion of the landscape in 

Europe, especially forested land, is managed, this may 

contribute to European N critical loads being higher, 

as N removal in harvesting results in greater N demand 

and storage during re-establishment of the forest stand.

Another possible explanation is that the response 

thresholds utilized in Europe are sometimes higher. A 

key example is lichen community responses: when a shift 

in community composition is considered the threshold 

of change, the N critical loads will be low. Some earlier 

work in Europe, in contrast, used a diff erent biological 

threshold—the near extirpation of lichen species—

leading to a higher N critical load (Bobbink et al. 2003). 

Another example of higher response thresholds used for 

setting critical loads in Europe relates to responses at the 

forest ecosystem level. Using a more sensitive endpoint, 

such as changes in N biogeochemistry interpreted as 

incipient responses of N saturation, led to a critical load 

<4 kg N ha-1 yr-1 in the Colorado Front Range (Rueth 

et al. 2003). Th is input is a subtle initial N enrichment 

response when compared to the magnitude of change 

for the critical loads thresholds in Europe (10 to 15 kg 

ha-1 yr-1). 

19.10 Selecting Critical Loads for 

Natural Resource Decisions

Empirical critical loads may be determined for 

multiple receptors, responses, and response thresholds 

within an ecosystem. Th us, several critical loads may 

be determined for a given ecosystem type (various 

ecosystem components) when data are available. Th e 

determination of these critical loads should be based 

on the best available scientifi c information. In section 

19.7, we presented some of the factors that aff ect where 

the N critical load will fall within the reported range, 

which could be used to refi ne the critical load estimate 

for a given ecosystem. Several other issues need to be 

addressed in order to select empirical critical loads based 

on the values reported in this document. Th e objectives 

of diff erent policy makers and resource managers will 

drive the decisions in selecting the N critical load 

most appropriate for a given area. First, the receptor 

of concern needs to be defi ned. Th e receptor may be 

a keystone species, the dominant species, the most 

sensitive species, or a threatened or endangered species. 

Next the response and response threshold must be set. 

In some cases, the response itself may be a signifi cant 

change in the ecosystem (e.g., change in species 

composition) and in some cases, it may be an earlier 

stage response (e.g., increase in foliar N concentration). 

One challenge in determining the critical threshold 

for the response is that it can be diffi cult to defi ne a 

pristine or unimpacted condition when much, or all, 

of the ecosystem has already experienced elevated N 

deposition. At this point, the policy makers or resource 

managers may also consider the degree of harm caused 

by a particular response; the determination of what level 

of harm is considered unacceptable is ultimately a policy 

decision.

Policy and resource management goals will determine 

the geographic extent and level of response that are 

considered unacceptable. In a conservation (e.g., 

wilderness) area, for example, any alteration in N 

cycling may be considered unacceptable—in these 

cases, when a range of N critical load values for a 

suite of resources and responses is presented, the 

lowest value is usually selected. Federal land managers 

responsible for Class I areas are required to “err of the 

side of protecting” these areas in determining the level 

of pollution these lands can tolerate. Other resource 

managers may choose to protect certain keystone species, 

threatened or endangered species, or species of economic 

or cultural signifi cance. Finally, in some cases, the level 

of certainty about the critical load may determine which 

critical load is used.
Chapter 19—Synthesis GTR-NRS-80 269

Another approach for presenting maps of critical loads, 

which is used in Europe, is to map the critical load that 

would protect 95 percent of the habitat or ecosystem 

area (Hettelingh et al. 2008).

19.11 Use of Critical loads in the 

United States

In the United States, the critical loads approach has 

not been widely used as an approach for ecosystem 

protection. For example, the Clean Air Act does not 

specifi cally require development and use of critical 

loads in implementing and assessing environmental and 

natural resource management programs. Nevertheless, 

the critical loads approach is being explored at Federal, 

state, and international levels as an ecosystem assessment 

tool with great potential to simplify complex scientifi c 

information and eff ectively communicate with the 

policy community and the public. Th e critical loads 

approach can provide a useful lens through which to 

assess the results of current policies and programs and 

to evaluate the potential ecosystem-protection value of 

proposed policy options.

Recent developments in the United States indicate that 

critical loads might be emerging as a useful ecosystem 

protection and program assessment tool. In 2004, the 

National Research Council recommended that critical 

loads be examined as a tool for ecosystem protection 

(NRC 2004). Between 2002 and 2006, several Federal 

agencies convened conferences and workshops to review 

the experience with critical loads in other countries, 

discuss science and modeling eff orts related to critical 

loads, and explore the possible future role of a critical 

loads approach as an air pollution control policy tool 

in the United States. A growing number of scientists 

are conducting research related to critical loads and are 

using various approaches to estimate critical loads in the 

United States.

Federal and state agencies are now exploring critical 

loads approaches to protect and manage sensitive 

ecosystems on Federal lands (Burns et al. 2008, 

Porter et al. 2005). In 2005, the U.S. Environmental 

Protection Agency (EPA) included a provision in its 

Nitrogen Dioxide Increment Rule enabling states to 

propose the use of critical loads information as part 

of their air quality management approach, to satisfy 

requirements under Clean Air Act provisions regarding 

“prevention of signifi cant deterioration.” Th e National 

Park Service is working with the Colorado Department 

of Public Health and Environment and EPA to address 

harmful impacts to air quality and other natural 

resources occurring in Rocky Mountain National Park 

in Colorado, and to reverse a trend of increasing N 

deposition. Th e National Park Service has established a 

resource management goal, linked to a critical load for 

wet N deposition of 1.5 kg ha-1 yr-1 for high elevation 

aquatic ecosystems. Resource managers within the U.S. 

Forest Service use critical loads to serve as a practical 

guideline when considering the potential impacts from 

new sources of emissions on resources in Class I areas. 

Forest Service resource managers are also using critical 

loads in the national watershed condition assessment 

process to identify potential areas on national forests 

for mitigation of ecosystem impairment. Several states 

have developed Total Maximum Daily Load (TMDL) 

estimates for acidic deposition to alleviate surface waters 

that are impaired due to elevated acidity (i.e., low pH, 

low ANC).

Th e critical loads approach recently has been utilized for 

ecosystem assessments within the broader environmental 

policy context as well. In 2005, for example, the Clean 

Air Act Advisory Committee recommended that the 

EPA use critical loads as a means to evaluate progress 

in reducing ecological impacts of air pollution. As a 

result, EPA’s 2007 and 2008 Annual Acid Rain Program 

progress reports utilized critical loads as a means of 

assessing the extent to which implementation of Title IV 

of the 1990 Clean Air Act Amendments has decreased 

S and N suffi ciently to protect acid-sensitive ecosystems 

in the Adirondack Mountain and Central Appalachian 

regions. Th e critical loads approach also has entered 

the realm of bilateral environmental policy. In 2008, 

the ninth biennial progress report completed under the 

1991 United States-Canada Air Quality Agreement 

included estimates of critical loads in acid-sensitive lakes 

in the northeastern United States. While the Canadian 

government routinely reports critical load data, this was 

the fi rst time critical load data for the United States were 
270 Chapter 19—Synthesis GTR-NRS-80

reported in a progress report prepared by the bilateral 

U.S.-Canada Air Quality Committee (Environment 

Canada 2008).

19.12 Summary

In an analysis of the nine major environmental 

challenges facing humanity, only three have clearly 

exceeded safe operating boundaries: biodiversity loss, 

increased N, and climate change (Rockström et al. 

2009). Because most terrestrial and many aquatic 

ecosystems are N limited under unpolluted conditions, 

increases in N input to ecosystems are likely to have an 

impact. Increased N deposition can cause a shift in the 

processing and movement of N (function) and to the 

physical composition (structure) of the ecosystem as 

evidenced by the examples below.

Large parts of the eastern United States, as well 

as localized areas in the West, are experiencing N 

deposition that exceeds the critical load for sensitive 

ecosystem components. Th e resources most threatened 

by elevated N deposition include freshwater diatoms, 

lichens, bryophytes, and herbaceous plants. Th e most 

signifi cant changes that we are currently observing in 

the United States in response to elevated N deposition 

are changes in species composition: losses of N-sensitive 

species, shifts in dominance, and losses of native species 

in favor of exotic, invasive species. Shifts in diatom and 

lichen community composition away from N-intolerant 

(oligotrophic) species are observed across the country. 

Alterations in herbaceous species are broadly observed, 

but are not always clearly documentable because of 

the long-term pollution inputs and other disturbances 

(including land-use change) that caused changes prior to 

most current studies.

Numerous examples illustrate the signifi cance of these 

species- and community-level eff ects. In serpentine 

grasslands in California, it was clearly demonstrated 

that unless N inputs are decreased or N is removed 

in biomass, a larval host plant and numerous nectar 

source plants utilized by a threatened and endangered 

butterfl y will decrease to levels unable to sustain the 

butterfl y population (Fenn et al. 2010, Weiss 1999). 

In Joshua Tree National Park and adjacent deserts in 

southern California, elevated N deposition favors the 

production of suffi cient invasive grass biomass to sustain 

fi res that threaten the survival of the namesake species 

(Fenn et al. 2010, Rao et al. 2010). Other sensitive 

ecosystems include alpine meadows, where relatively 

low levels of N deposition have already changed species 

composition in this fragile community (Bowman et 

al. 2006). Changes in historical diatom community 

composition from N-limited to N-tolerant species have 

been observed in lake sediment cores at many locations 

in the western United States, providing early evidence 

of eutrophication of freshwater ecosystems (Saros et al. 

2010; Wolfe et al. 2001, 2003).

Changes in ecosystem structure are linked to changes 

in ecosystem function. For example, extirpation of 

lichens can alter food webs by reducing the availability 

of nesting material for birds, invertebrate habitat, and 

critical winter forage for mammals, and can also aff ect 

nutrient cycling (Cornelissen et al. 2007). In California, 

where elevated N deposition and arid low-biomass 

ecosystems coincide (e.g., coastal sage scrub, grassland, 

desert), N-enhanced growth of invasive species resulting 

in major alterations of plant communities, conversion 

of vegetation type, and increased fi re risk, even in areas 

where fi re is normally infrequent (Allen et al. 2009, 

Fenn et al. 2010, Rao et al. 2010).

Th ere is also evidence that N deposition contributes 

to multiple stress complexes and has decreased forest 

sustainability in California (Grulke et al. 2009) and 

in North Carolina (McNulty and Boggs 2010). In 

North Carolina, elevated N deposition predisposed a 

pine ecosystem to a pest outbreak following a drought 

(McNulty and Boggs 2010). Another example of N 

deposition interactions with other forest stressors 

is the observation that increased NO3

-

 leaching and 

nitrifi cation contribute to soil acidifi cation and 

depletion of available nutrient cations which have 

negative eff ects on tree growth, vigor, and cold tolerance 

in some forests. Elevated NO3

-

 concentrations in 

surface water and groundwater may diminish drinking 

water quality, although the drinking water standard 

is often only exceeded for brief periods, for example 

when N saturated watersheds are disturbed (e.g., fi re or 

harvesting). Th ese types of complex interactions may 
Chapter 19—Synthesis GTR-NRS-80 271

be diffi cult to predict, but may intensify the impact of 

elevated N deposition in concert with other stressors 

including consequences of climate change.

Further examples of changes in ecosystem structure and 

function are observed in coastal areas, where increased 

N export has led to toxic algal blooms (Rabalais 2002). 

As an example of N deposition eff ects on trace gas 

chemistry and climate change, N loading to ecosystems 

results in increased emissions of N trace gases, such as 

nitric oxide (NO), an ozone (O3) precursor; nitrous 

oxide (N2

O), a long-lived greenhouse gas; as well as 

declines in soil uptake of methane (CH4), another longlived greenhouse gas (Liu and Greaver 2009).

Th e above examples provide compelling evidence for 

signifi cant alteration of ecosystem structure and function 

in ecoregions across the United States due to elevated 

N deposition. To protect ecosystems from harm caused 

by N deposition, it is necessary to identify the level of N 

deposition which would lead to detrimental ecological 

eff ects. Empirical critical loads for N provide a valuable 

approach for evaluating the risk of harm to ecosystems. 

Th is approach has been used broadly in Europe 

(Bobbink et al. 2003, UBA 2004) and has the advantage 

of being scientifi cally based on observed responses. Th is 

link to actual ecosystem responses is especially benefi cial 

in resource management and policy contexts.

Th is report provides the fi rst comprehensive assessment 

of empirical critical loads of N for ecoregions across the 

United States. It represents an important step toward 

providing policy makers and resource managers with 

a tool for ecosystem protection as suggested by the 

National Research Council (NRC 2004).

19.13 Future Research Priorities

Th e principal knowledge gaps that limit our 

understanding of N impacts on ecosystems include 

poor quantifi cation of total N deposition (especially 

in deposition hotspots) and the paucity of long-term, 

low N-fertilization studies and adequate deposition 

gradient studies. A higher density of long-term, low 

N fertilization studies, as well as long-term and larger 

scale gradient studies across both a greater diversity of 

ecosystem types and regions of low N deposition, are 

necessary to develop dose response curves that would 

better defi ne critical loads and the associated uncertainty. 

In the United States, observations of ecosystem response 

to N inputs are particularly limited in the Tundra, 

Taiga, and North American Desert ecoregions.

Other important issues include: 

• Th e diff erential response to reduced (NHx

) 

versus oxidized (NOy

) N inputs. Because some 

plants are particularly sensitive to NHx

 inputs 

(Krupa 2003) while others are more sensitive 

to NOy

 (Nordin et al. 2006), assembling more 

comprehensive data about these species-specifi c 

responses would allow more accurate assessment 

of potential risks to ecosystems in relation to 

the major N emissions sources. Oxidized and 

reduced N forms also result in diff erent levels of 

acidifi cation.

• Impacts on plant biodiversity in forests have 

not been well described, in part because of 

the diffi culty of assessing such changes in 

ecosystems with longer-lived organisms, and in 

part, because in many of these ecosystems the 

herbaceous plants have already been altered by 

historical N deposition, other pollutants, or 

habitat alteration.

• Eff ects of N deposition on forest growth and 

sustainability. Insuffi cient data are available 

to determine critical loads for the eff ects 

of increasing N fertility on pest outbreaks, 

drought, cold tolerance, tree vigor, and multiple 

stress complexes in general.

• Identifi cation of mechanisms that control plant 

and ecosystem responses to N deposition. Th is 

is a necessary step in refi ning critical loads 

estimates, improving their reliability, and laying 

the groundwork for more complex dynamic 

models, which are necessary for broad scale 

assessments, including detailed national maps of 

empirical critical loads for N.

Th e objective of future critical-loads-driven research 

should be to fi ll in gaps in data and improve the 
272 Chapter 19—Synthesis GTR-NRS-80

reliability of estimates. When more data are available, 

especially data of higher reliability, it will be possible 

to make a map of empirical critical loads of N such as 

that developed for California (Figure 19.11; Fenn et 

al. 2010). Th is map also utilized N deposition modeled 

at a much fi ner grid (4 km x 4 km). One approach for 

assembling the data necessary to estimate critical loads 

is using systematically sampled large scale studies (for 

example the U.S. Forest Service FIA grid sampling) 

to ensure that a broad gradient in N deposition, 

climate, and other variables are included in the dataset 

generated (e.g., Geiser and Neitlich 2007, Jovan 2008). 

Th is approach would allow extrapolation of the N 

critical loads to a broad area with confi dence. Better 

understanding of when the “baseline” response has been 

altered by prior N deposition is necessary to identify 

empirical critical loads using N deposition gradient or N 

addition studies. Long-term monitoring is also necessary 

to evaluate the scope of particular responses and to 

assess future responses to reductions in N deposition. 

Th e accuracy of empirical N critical loads is limited 

by the accuracy of the N deposition values used. Th us, 

improving estimates of total N deposition is essential 

for improving empirical critical loads and exceedance 

estimates.

Th us, the highest research priorities should be:

(1) Lichens: systematic sampling of lichens in areas 

where there are few data in combination with 

analyzing existing FIA data could yield very 

useful results for this sensitive indicator.

(2) Diatoms in lake sediments and phytoplankton 

in lakes: diatoms preserved in lake sediments 

can be used to identify when and at what 

atmospheric N deposition amounts critical 

Figure 19.11—Composite showing 

critical load exceedance for seven 

vegetation types in California (Fenn 

et al. 2010).
Chapter 19—Synthesis GTR-NRS-80 273

thresholds were crossed in the past. Th e 

approach has applications far beyond the 

few locations where it has been applied. 

Phytoplankton in lakes respond rapidly to 

changes in limiting nutrients, and far more 

studies are needed to show how (and how 

much) N deposition is needed to aff ect the 

N:P stoichiometric ratio and subsequent 

shifts in food webs and ecological processes 

in oligotrophic lakes. Equally important are 

studies of whether the removal of N deposition 

allows the return of N-limited conditions and 

oligotrophic phytoplankton. A fi nal and key 

research priority for lakes are experiments 

to determine the eff ects of atmospheric N 

deposition on algal biodiversity.

(3) Herbaceous species: research is needed to 

identify the most responsive species across a 

variety of ecosystem types.

(4) Identifying indicator species in general: species 

which allow evaluation of the ecosystem 

condition are especially useful for empirical N 

critical loads estimates.

(5) Long-term low N addition experiments: more 

long-term low N fertilization studies will help 

make more accurate determinations of critical 

loads.

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levels/loads and geographical areas where they are 

exceeded. Berlin: Federal Environmental Agency 

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285

APPENDIX 1.—PLANT SPECIES COMMON AND SCIENTIFIC NAMES

Life form Common name Latin name Alternate name

Tree/shrub Acacia Acacia Mill.

Shrub/subshrub//Forb/

herb

Agave Agave L.

Tree/shrub Alder Alnus Mill.

Forb/herb Alpine chickweed Cerastium alpinum L.

Graminoid Alpine fescue Festuca brachyphylla Schult. ex Schult. & Schult. f.

Tree American chestnut Castanea dentata (Marsh.) Borkh.

Shrub American tarwort Flourensia cernua DC. Tarbrush

Forb/herb Annual ragweed Ambrosia artemisiifolia L.

Shrub Antelope bitterbrush Purshia tridentata (Pursh) DC. Antelope brush; 

buckbrush

Graminoid Arctic bluegrass Poa arctica R. Br.

Lichen Arctic kidney Nephroma arcticum (L.) Torss.

Subshrub/shrub Arctic willow Salix arctica Pall.

Graminoid Arrow grass Triglochin L.

Lichen Asahina’s cartilage lichend Ramalina asahinae (L.) Ach.

Tree Ash Fraxinus L.

Tree Balsam fi r Abies balsamea (L.) Mill.

Tree Balsam poplar Populus balsamifera L.

Tree Basswood Tilia L.

Shrub Bearberry Arctostaphylos alpina (L.) Spreng. alpine bearberry

Lichen Beard lichena Usnea Dill. ex Adans. 

Tree Beech Fagus L.

Graminoid Big bluestem Andropogon gerardii Vitman

Tree Big leaf maple Acer macrophyllum Pursh

Tree/Shrub Big sagebrush Artemisia tridentata Nutt.

Tree Bigtooth aspen Populus grandidentata Michx.

Tree/shrub Black cherry Prunus serotina, Ehrh.

Tree/shrub Black mangrove Avicennia germinans L.

Tree Black oak Quercus velutina, Lam.

Tree Black spruce Picea mariana (Mill.) Britton, Sterns & Poggenb.

Graminoid Blue grama Bouteloua gracilis (Willd. ex Kunth) Lag. ex Griffi ths

Tree Blue oak Quercus douglasii Hook. & Arn.

Graminoid Bluestem Andropogon L.

Tree Box elder Acer negundo L.

Forb/herb Bracken fern Pteridium aquilinum var. pubescens

Tree Bristlecone pine Pinus aristata Engelm.

Subshrub/Shrub Brittlebush Encelia Adans.

Shrub Buckbrush Ceanothus cuneatus (Hook.) Nutt.

Graminoid Buffalo grass Bouteloua dactyloides (Nutt.) J.T. Columbus Buchloe dactyloides

Tree/shrub Buttonwood Conocarpus erectus L.

Tree/shrub California ash Fraxinus dipetala Hook. & Arn.

Tree/shrub California black oak Quercus kelloggii Newberry

Tree/shrub California buckeye Aesculus californica (Spach) Nutt.

Tree California foothill pine Pinus sabiniana Douglas ex Douglas Digger pine

continued
286

Tree California redwood Sequoia sempervirens (Lamb. ex D. Don) Endl.

Forb/herb Canada goldenrod Solidago Canadensis L.

Forb/herb Canadian sandspurry Spergularia canadensis (Pers.) G. Don

Shrub Ceanothus (California 

lilac)

Ceanothus spp.

Lichen Cetraria lichena Cetraria ericetorum Opiz

Lichen Cetraria lichenb Cetrariella delisei (Bory ex Schaerer) Karnefelt & A. 

Thell 

Lichen Cetraria lichenb Flavocetraria cucullata (L.) Karnefelt & A. Thell 

Lichen Cetraria lichenb Flavocetraria nivalis (Bellardi) Karnefelt & A. Thell 

Shrub Chamise Adenostoma fasciculatum Hook. and Arn.

Graminoid Cheatgrass Bromus tectorum L.

Tree Chestnut oak Quercus prinus L.

Shrub (cactus) Cholla Cylindropuntia spp.

Forb/herb Clover Trifolium spp.

Tree Coast Douglas-fi r Pseudotsuga menziesii (Mirb.) Franco var. menziesii

Tree/shrub Coast live oak Quercus agrifolia Née var. oxyadenia (Torr.) J.T. 

Howell

Shrub Coastal sagebrush Artemisia californica Less. California sagebrush

Tree/subshrub/shrub Coco plum Chrysobalanus icaco L.

Graminoid Common Mediterranean 

grass

Schismus barbatus (Loefl . ex L.) Thell. Mediterranean 

splitgrass

Graminoid Cordgrass Spartina spp.

Graminoid Cotton grassa Eriophorum L.

Shrub Creosote bush Larrea tridentata (DC.) Coville

Lichen Cumberland 

xanthoparmelia lichen

Xanthoparmelia cumberlandia (Gyelnik) Hale 

Lichen Cup lichen Cladonia P. Browne Reindeer lichenc

Graminoid Curly sedge Carex rupestris All.

Bryophyte Dicranum mossa Dicranum polysetum Sw. 

Bryophyte Dicranum mossa Dicranum scoparium Hedw. 

Tree/shrub Dogwood Cornus L.

Tree Douglas-fi r Pseudotsuga menziesii (Mirb.) Franco

Shrub/subshrub Dwarf birch Betula nana L.

Tree Eastern hemlock Tsuga canadensis (L.) Carrière

Subshrub/shrub Eastern Mojave 

buckwheat 

Eriogonum fasciculatum Benth. California buckwheat

Tree Eastern red cedar Juniperus virginiana L.

Tree Eastern white pine Pinus strobus L. white pine

Forb/herb Eelgrass Zostera spp.

Tree Elm Ulmus L.

Tree Engelmann oak Quercus engelmannii Greene

Tree Engelmann spruce Picea engelmannii Parry ex Engelm.

Subshrub/shrub/forb/

herb

Entireleaf mountain avens Dryas integrifolia Vahl

Tree Eucalyptus Eucalyptus camaldulensis Dehnh.

Life form Common name Latin name Alternate name

Appendix 1.—continued

continued
287

Lichen Felt lichena Peltigera aphthosa (L.) Willd. 

Lichen Felt lichena Peltigera polydactylon (Necker) Hoffm.

Tree Florida slash pine Pinus elliottii Engelm.

Tree/shrub Flowering dogwood Cornus fl orida L.

Tree Fremont cottonwood Populus fremontii S. Watson

Graminoid Galleta grass Hilaria jamesii (Torr.) Benth. Pleuraphis jamesii

Torr.; James galleta

Tree Giant sequoia Sequoiadendron giganteum (Lindl.) J. Buchholz

Graminoid Grama grass Bouteloua spp.

Tree Grand fi r Abies grandis (Douglas ex D. Don) Lindl.

Shrub Greasewood Sarcobatus Nees Sarcobatus

vermiculatus (Hook.) 

Torr.

Subshrub/Forb/herb Great ragweed Ambrosia trifi da L.

Subshrub/Forb/herb Green pitcher plant Sarracenia oreophila (Kearney) Wherry 

Lichen Greygreen reindeer lichen Cladonia rangiferina (L.) F. H. Wigg.

Tree/shrub Gumbo limbo Bursera simaruba (L.) Sarg.

Tree/shrub Hackberry Celtis occidentalis L.

Tree/shrub Hawaiian holly Ilex anomala Hook. & Arnott

Shrub Heather Calluna vulgaris (L.) Hull

Tree Hickory Carya Nutt.

Lichen Horsehair lichen a Bryoria capillaris (Ach.) Brodo & D. Hawksw. 

Lichen Horsehair lichena Bryoria Brodo & D. Hawksw. 

Lichen Hydrothyria lichen Peltigera hydrothyria Miadl. & Lutzoni Hydrothyria venosa

Lichen Inactive tube lichen Hypogymnia inactiva (Krog) Ohlsson

Tree Incense cedar Calocedrus decurrens (Torr.) Florin

Graminoid Indian ricegrass Oryzopsis hymenoides (Roem. & Schult.) Ricker ex 

Piper

Achnatherum

hymenoides (Roem. & 

Schult.) Barkworth

Graminoid Indiangrass Sorghastrum nutans (L.) Nash

Tree Jack pine Pinus banksiana Lamb.

Tree Jeffrey pine Pinus jeffreyi Balf.

Tree/shrub Joshua tree Yucca brevifolia Engelm.

Tree/shrub Juniper Juniperus L.

Bryophyte Juniper polytrichum moss Polytrichum juniperinum Hedw. 

Graminoid Kentucky bluegrass Poa pratensis L.

Tree Koa Acacia koa A. Gray

Shrub Labrador tea Ledum L.

Shrub/subshrub Lingonberry vaccinium vitis-idaea L.

Graminoid Little bluestem Schizachyrium scoparium (Michx.) Nash 

Tree Live oak Quercus virginiana Mill. Virginia live oak

Tree Lodgepole pine Pinus contorta Douglas ex Louden

Bryophyte Magellan’s sphagnum Sphagnum magellanicum Brid.

Tree Mahogany Swietenia mahogani (L.) Jacq.

Shrub Manzanita Arctostaphylos Adans.

Tree Maple Acer L.

continued

Life form Common name Latin name Alternate name

Appendix 1.—continued
288

Subshrub/forb/herb Marsh rosemary Limonium californicum (Boiss.) A. Heller

Lichen Menzies’ cartilage lichen Ramalina menziesii Taylor 

Tree/shrub Mesquite Prosopis L.

Tree Mexican mountain pine Pinus hartwegii Lindl.

Subshrub/shrub Mountain heather Cassiope D. Don 

Tree Mountain hemlock Tsuga mertensiana (Bong.) Carrière

Shrub Mountain-avens Dryas L.

Graminoid Muhly grass Muhlenbergia Schreb.

Tree Nemoca cimarrona Ocotea spathulata Mez

Tree/shrub Netleaf hackberry Celtis laevigata Willd. var. reticulata (Torr.) L.D. 

Benson

Tree Nootka cypress Callitropsis nootkatensis (D. Spach) Alaska yellowcedar; Cupressus 

nootkatensis D. Don

Tree Northern red oak Quercus rubra L. Red oak

Tree Northern white-cedar Thuja occidentalis L.

Tree Norway spruce Picea abies (L.) Karst.

Tree Oak Quercus L.

Tree/shrub Ohi’a lehua Metrosideros polymorpha Gaud

Tree Olapa Cheirodendron

Lichen Orange wall lichena Xanthoria (Fr.) Th. Fr.

Lichen Orange wall lichene Xanthomendoza S. Kondr. & Kärnefelt

Lichen Orange wall lichene Xanthomendoza fallax (Hepp ex Arnold) Søchting, 

Kärnefelt & S. Kondr. 

Tree Oregon ash Fraxinus latifolia Benth

Tree/shrub Oregon white oak Quercus garryana Douglas ex Hook. Garry oak

Tree Pacifi c madrone Arbutus menziesii Pursh

Tree Pacifi c silver fi r Abies amabilis (Douglas ex Louden) Douglas ex 

Forbes

silver fi r

Tree/shrub Palo Colorado Cyrilla racemifl ora L. Swamp titi

Tree/shrub Paloverde Parkinsonia L.

Tree Paper birch Betula papyrifera Marsh. White birch

Lichen parmotrema lichena Parmotrema stuppeum (Taylor) Hale

Tree Persimmon Diospyros virginiana L. common persimmon

Tree Pignut hickory Carya glabra (Mill.) Sweet

Tree Pine Pinus L.

Bryophyte Polytrichum moss Polytrichum commune Hedw.

Tree Ponderosa pine Pinus ponderosa C. Lawson

Shrub (cactus) Prickly pear cactus Opuntia Mill.

Forb/herb Purple mountain saxifrage Saxifraga oppositifolia L.

Subshrub/forb/herb Purple pitcher plant Sarracenia purpurea L.

Graminoid Quackgrass Agropyron repens (L.) P. Beauv. Elymus repens (L.) 

Gould

Tree Quaking aspen Populus tremuloides Michx. Trembling aspen

Graminoid Quillworts Isoetes L.

Shrub Rabbit brush Chrysothamnus spp.

Life form Common name Latin name Alternate name

Appendix 1.—continued

continued
289

Lichen Ragged lichena Platismatia glauca (L.) W. L. Culb.. & C. F. Culb.

Tree Red alder Alnus rubra Bong.

Graminoid Red brome Bromus rubens L.

Tree/shrub Red mangrove Rhizophora mangle L.

Tree Red maple Acer rubrum L.

Tree Red pine Pinus resinosa Ait.

Tree Red spruce Picea rubens Sarg.

Lichen Reindeer lichenc Cladonia mitis Sandst.

Lichen Ring lichena Evernia mesomorpha Nyl. 

Tree River birch Betula nigra L. black birch

Tree Rocky Mountain Douglasfi rPseudotsuga menziesii (Mirb.) Franco var. glauca

(Beissn.) Franco 

Forb/herb Round head lespedeza Lespedeza capitata Michx. Round-head bushclover

Forb/herb Roundleaf sundew Drosera rotundifolia L.

Forb/herb Russian Thistle Salsola iberica (Sennen & Pau) Botsch. ex Czerep. Salsola tragus L.

Subshrub/shrub Sacahuista Nolina microcarpa S. Watson

Tree Sacred fi r Abies religiosa (H.B.K.) Schl. et Cham.

Forb/herb Sage Salvia L.

Subshrub/shrub Sagebrush Artemisia L.

Tree (cactus) Saguaro Cereus gigantea Engelm. Carnegiea gigantea

(Engelm.) Britton & 

Rose

Shrub Saltbush Atriplex L.

Graminoid Saltgrass Distichlis spicata (L.) Greene

Graminoid Saltmeadow cordgrass Spartina patens (Aiton) Muhl.

Tree/shrub Saw palmetto Serenoa repens (Bartram) Small

Tree Scarlet oak Quercus coccinea Münchh.

Bryophyte Schreber’s big red stem 

moss 

Pleurozium schreberi (Brid.) Mitt. 

Tree Scots pine Pinus sylvestris L. Scotch pine

graminoid Sheep fescue Festuca ovina L.

Lichen Shield lichena Parmelia sulcata 

Tree Sitka spruce Picea sitchensis (Bong.) Carrière

Lichen Skin lichena Leptogium rivale Tuck.

Graminoid Smooth cordgrass Spartina alternifl ora Loisel.

Lichen Snow lichena Stereocaulon Hoffm. 

Lichen Snow lichena Stereocaulon paschale (L.) Hoffm. 

Tree/shrub Soaptree yucca Yucca elata (Engelm.) Engelm.

Tree/shrub Southern California walnut Juglans californica S. Watson

Tree Southern oak Quercus falcata Michx. Southern red oak

Tree Spanish fi r Abies pinsapo Boiss.

Bryophyte Sphagnum Sphagnum fuscum (Schimp.) Klinggr. brown peatmoss

Bryophyte Sphagnuma Sphagnum L.

Bryophyte Sphagnuma Sphagnum fuscum (Schimp.) Klinggr. 

Bryophyte Splendid feather moss Hylocomium splendens (Hedw.) Schimp. 

Life form Common name Latin name Alternate name

Appendix 1.—continued

continued
290

Lichen Star reindeer lichen Cladonia stellaris (Opiz) Pouzar & Vězda

Forb/herb Stickywilly Galium aparine L. 

Forb/herb Stork’s bill Erodium cicutarium (L.) L’Hér. ex Aiton

Tree/shrub/vine Strangler fi g Ficus aurea Nutt.

Tree Subalpine fi r Abies lasiocarpa (Hook.) Nutt.

Tree Sugar maple Acer saccharum Marsh.

Tree Sugar pine Pinus lambertiana Douglas

Tree Sweet gum Liquidambar styracifl ua L.

Graminoid Switchgrass Panicum virgatum L.

Tree Tabonuco tree Dacryodes excels Vahl

Graminoid Tall fescue Schedonorus phoenix (Scop.) Holub Festuca arundinacea

Schreb.

Tree Tamarack Larix larcina (Du Roi) K. Koch American larch

Tree/shrub Texas live oak Quercus fusiformis Small

Graminoid Threeawn Aristida L.

Graminoid Tobosagrass Pleuraphis mutica Buckley Hilaria mutica

Tree/shrub Toyon Heteromeles arbutifolia (Lindl.) M. Roem.

Lichen Tube lichena Hypogymnia enteromorpha (Ach.) Nyl.

Lichen Tube lichena Hypogymnia physodes (L.) Nyl. 

Forb/herb Tule Schoenoplectus acutus (Muhl. ex Bigelow) A. Löve & 

D. Löve var. occidentalis (S. Watson) S.G. Sm. 

Graminoid Tussock cottongrass Eriophorum vaginatum L.

Subshrub/forb/herb Virginia glasswort Salicornia depressa Standl. Salicornia virginica L.

Graminoid Wavy hairgrass Deschampsia fl exuosa (L.) Trin.

Tree Western hemlock Tsuga heterophylla (Raf.) Sarg.

Tree Western juniper Juniperus occidentalis Hook.

Tree Western red cedar Thuja plicata Donn ex D. Don

Tree Western white pine Pinus monticola Douglas ex D. Don

Graminoid Wheatgrass Agropyron Gaertn.

Tree White ash Fraxinus americana L.

Tree White fi r Abies concolor (Gord. & Glend.) Lindl. ex Hildebr.

Tree/shrub White mangrove Laguncularia racemosa (L.) Gaertn. f.

Tree White spruce Picea glauca (Moench)Voss

Subshrub/shrub Whortleberry Vaccinium myrtillus L.

Tree/shrub Willows Salix L.

Lichen Witch’s hair lichena Alectoria nigricans (Ach.) Nyl.

Lichen Witch’s hair lichena Alectoria ochroleuca (Hoffm.) A. Massal.

Lichen Witch’s hair lichena Alectoria sarmentosa (Ach.) Ach.

Lichen Wolf lichen Letharia vulpina (L.) Hue

Tree Yellow birch Betula alleghaniensis Britt.

Tree Yellow-poplar Liriodendron tulipifera L. Tulip tree

a The same common name is used for all species in the genus.

b This common name has not been updated in the NRCS plants database (NRCS 2010) since the genus name changed.

c General common name for species that were formerly in the genus Cladina. d This common name is not listed in the NRCS plants database (NRCS 2010), but follows NRCS naming conventions.

e General common name for species that were formerly in the genus Xantharia; this common name has not been updated in the NRCS plants 

database (NRCS 2010) since the genus name changed.

NRCS (National Resources Conservation Service). 2010. PLANTS database. Baton Rouge, LA: U.S. Department of Agriculture, Natural 

Resources Conservation Service, National Plant Data Center. Available at: http://plants.usda.gov (Accessed February 9, 2011).

Life form Common name Latin name Alternate name

Appendix 1.—continued
291

APPENDIX 2.—CHEMICAL APPENDIX

Chemical formula or abbreviation Chemical name or terminology

Al Aluminum

ANH4I Aitken mode ammonium mass

ANH4J Accumulation mode ammonium mass

ANO3I Aitken mode aerosol nitrate mass

ANO3J Accumulation mode nitrate mass

C Carbon

C18H16O7 Usnic acid

CH4 Methane

CO2 Carbon dioxide

Ca Calcium

HNO3 Nitric acid

HONO Nitrous acid

H2S Hydrogen sulfi de

H2SO4 Sulfuric acid

Hg Mercury

K Potassium

KH2PO4 Potassium dihydrogen phosphate

MA-PAN PAN analogue formed from methacrolein (C4H6O)

Mg Magnesium

N Nitrogen

NH3 Ammonia 

NH4

- Ammonium

NH4Cl Ammonium chloride

NH4NO3 Ammonium nitrate

(NH4)2SO4 Ammonium sulfate

NHx NH3 and NH4

-

NO Nitric oxide

NO2 Nitrogen dioxide

NO3

- Nitrate

NOx Nitric oxides

NOy NO, NO2, nitric acid (HNO3), organic and inorganic nitrates

N2 Nitrogen gas

N2O Nitrous oxide

N2O5 Dinitrogen pentoxide

NaNO3 Sodium nitrate

O3 Ozone

P Phosphorus

PAN (C2H3NO5) Peroxyacetyl nitrate

PAN2 (RCO5N) Higher peroxyacetyl nitrate (based on peroxypropyl nitrate)

PBZN (C7H5NO5) Peroxybenzoyl nitrate

RNO3 Organic nitrates

S Sulfur

SO2

- Sulfur dioxide

SO4

2- Sulfate

SOx Sulfur oxides
Printed on Recycled Paper

Pardo, L.H.; Robin-Abbott, M.J.; Driscoll, C.T., eds. 2011. Assessment of Nitrogen 

deposition effects and empirical critical loads of Nitrogen for ecoregions of the 

United States. Gen. Tech. Rep. NRS-80. Newtown Square, PA: U.S. Department of 

Agriculture, Forest Service, Northern Research Station. 291 p.

This report synthesizes current research relating atmospheric nitrogen (N) deposition 

to effects on terrestrial and aquatic ecosystems in the United States and to identify 

empirical critical loads for atmospheric N deposition. The report evaluates the following 

receptors: freshwater diatoms, mycorrhizal fungi and other soil microbes, lichens, 

herbaceous plants, shrubs, and trees. The main responses reported fell into two 

categories: (1) biogeochemical; and (2) individual species, population, and community 

responses. The range of critical loads for nutrient N reported for U.S. ecoregions, inland 

surface waters, and freshwater wetlands is 1 to 39 kg N ha-1 y-1. This range spans the 

range of N deposition observed over most of the country. The empirical critical loads 

for N tend to increase in the following sequence for different life forms: diatoms, lichens 

and bryophytes, mycorrhizal fungi, herbaceous plants and shrubs, trees.

KEY WORDS: air pollution, atmospheric N deposition, critical nitrogen loads, 

biodiversity, nitrate leaching, nitrogen saturation, plant nitrogen concentration, 

natural resource protection, vegetation type conversion

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