EPA/600/R-05/087
October 2005
Ecosystem Stress from Chronic Exposure
to Low Levels of Nitrate
Eric E. Jorgensen, Scott M. Holub, Paul M. Mayer, & Mary E. Gonsoulin
U.S. Environmental Protection Agency
Office of Research and Development
National Risk Management Research Laboratory
Ada, Oklahoma 74820
Rendahandi G. Silva & Ann E. West
Oak Ridge Institute for Science and Education
Oak Ridge, Tennessee 37831
Susan J. Tunnell, Jay E. Clark, Jennifer L Parsons, David M. Engle, & Eric C. Hellgren
Oklahoma State University
Stillwater, Oklahoma 74078
Julie D.H. Spears & Clyde E. Butler
Shaw Environmental, Inc.
Ada, Oklahoma 74820
D.M. Leslie, Jr.
United States Geological Survey
Oklahoma Cooperative Fish & Wildlife Research Unit
Stillwater, Oklahoma 74078
Project Officer
Eric E. Jorgensen
Ground Water and Ecosystems Restoration Division
National Risk Management Research Laboratory
Ada, Oklahoma 74820
National Risk Management Research Laboratory
Office of Research and Development
U.S. Environmental Protection Agency
Cincinnati, Ohio 45268
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Notice
The U.S. Environmental Protection Agency through its Office of Research and
Development funded, and managed the research described herein. It has been
subjected to the Agency's peer and administrative review and has been approved
for publication as an EPA document. Mention of trade names or commercial prod-
ucts does not constitute endorsement or recommendation for use.
All research projects making conclusions or recommendations based on environ-
mentally related measurements and funded by the U.S. Environmental Protection
Agency are required to participate in the Protection Agency Quality Assurance
Program. This project was conducted under an approved Quality Assurance Project
Plan. The procedures specified in this plan were used without exception. Informa-
tion on the plan and documentation of the quality assurance activities and results
are available from the Principal Investigator.
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Foreword
The U.S. Environmental Protection Agency is charged by Congress with protecting the Nation's land, air,
and water resources. Under a mandate of national environmental laws, the Agency strives to formulate
and implement actions leading to a compatible balance between human activities and the ability of natural
systems to support and nurture life. To meet this mandate, EPA's research program is providing data
and technical support for solving environmental problems today and building a science knowledge base
necessary to manage our ecological resources wisely, understand how pollutants affect our health, and
prevent or reduce environmental risks in the future.
The National Risk Management Research Laboratory is the Agency's center for investigation of technologi-
cal and management approaches for preventing and reducing risks from pollution that threatens human
health and the environment. The focus of the Laboratory's research program is on methods and their
cost-effectiveness for prevention and control of pollution to air, land, water, and subsurface resources;
protection of water quality in public water systems; remediation of contaminated sites, sediments and
ground water; prevention and control of indoor air pollution; and restoration of ecosystems. NRMRL
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problems by: developing and promoting technologies that protect and improve the environment; advanc-
ing scientific and engineering information to support regulatory and policy decisions; and providing the
technical support and information transfer to ensure implementation of environmental regulations and
strategies at the national, state, and community levels.
This publication has been produced as part of the Laboratory's strategic long-term research plan. It is
published and made available by EPA's Office of Research and Development to assist the user com-
munity and to link researchers with their clients..
In 1998 we initiated an integrated multi-disciplinary study investigating the effects of chronic exposure
of ecosystems to low doses of bioavailable nitrogen. We investigated several aspects of ecosystem
response to chronic exposure to low doses of bioavailable nitrogen on sixteen 40x40-m study plots in
south-central Oklahoma in conjunction with complementary short-term field and laboratory studies. In this
nitrogen-limited system, the ability of the soil system to adapt to new nitrogen inputs was compromised
after 1 year of exposure. Concentrations of nitrate-N in the soil peaked at 1169% more than expected
and averaged 254% greater than expected. Our experiments demonstrate that even the relatively small
amounts of bioavailable nitrogen that are deposited in precipitation have the capacity to change multiple
aspects of ecosystem nitrogen retention, sequestration, and processing. The changes observed are al-
ways deleterious in that they lead to greater concentrations of nitrate-N and thereby make more available
for leaching to surface and groundwater. As outputs of nitrogen to the atmosphere can reasonably be
expected to increase in the foreseeable decades, it is prudent to identify and develop management op-
tions now to both restore ecosystems that are already compromised and to buffer affects to ecosystems
that are at risk from new nitrogen inputs.
Stephen G. Schmelling, Directc
Ground Water and Ecosystenrts Restoration Division
National Risk Management Research Laboratory
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Abstract
Throughout the eastern United States, from the Front Range of the Rocky Mountains to the Atlantic Ocean,
bioavailable nitrogen has been falling in the rain since the industrial revolution. Bioavailable nitrogen is a
limiting nutrient throughout this region. While long-term research conclusively demonstrates that exposure
of soil ecosystems to large doses of bioavailable nitrogen leads to deleterious environmental impacts (i.e.,
eutrophication, toxic algae blooms, hypoxia, toxicity, acid rain, global climate change) that can compro-
mise people's health and the economic vigor of communities, the potential effects of chronic exposure
to lower doses of bioavailable nitrogen are relatively unknown. However, symptoms of compromised
ecosystem function that may be attributable to chronic exposure to bioavailable nitrogen are widespread;
many forests routinely leach nitrogen to surface and groundwater and nitrate-N concentration in estuaries
perturbs aquatic food-webs and affects fisheries. These observations, among others, support the hypoth-
esis that ecosystem function can be (and has been) deleteriously impacted by chronic exposure to low
doses of bioavailable nitrogen. To investigate this, in 1998 we initiated an integrated multi-disciplinary
study investigating the effects of chronic exposure of ecosystems to low doses of bioavailable nitrogen.
We investigated several aspects of ecosystem response to chronic exposure to low doses of bioavailable
nitrogen on sixteen 40x40-m study plots in south-central Oklahoma in conjunction with complementary
short-term field and laboratory studies. Plots were manipulated in a factorial arrangement such that 4 plots
each received fertilizer only (16.3 kg N ha~1 yr1), herbivory manipulation only (fenced), a combination of
fertilizer and herbivory manipulation, or were left as controls. Herbivory population was manipulated by
constructing a 2-m tall chain link fence of 2.5-cm wire mesh. In this nitrogen-limited system, the ability
of the soil system to adapt to new nitrogen inputs was compromised after 1 year of exposure. Concen-
trations of nitrate-N in the soil peaked at 1169% more than expected and averaged 254% greater than
expected. Plant growth was affected by nitrogen application, wherein biomass increased on fertilized
plots and diversity was related to distribution of Festuca arundinacea. Microbial activity was naturally
limited in this system by carbon availability, but this tendency was exacerbated by additional inputs of
nitrogen: further, microbial population response was not qualitatively different in soils that received small
nitrogen additions vs. soils that received larger nitrogen additions. The presence of large numbers of
small mammals coincided with high concentrations of soil nitrate-N. We estimate that herbivores may
be able to re-circulate up to 67% of the bioavailable nitrogen deposited back into the plant and microbial
pathways, thereby producing a self reinforcing positive feedback loop leading to ever greater concentra-
tions of soil nitrate-N. This could lead to increased nitrate-N leaching to surface and ground water. The
ability of detritus pathways to process nitrogen inputs was compromised after 6 months and this tendency
was increased when macroinvertebrate communities were restricted. These experiments demonstrate
that even the relatively small amounts of bioavailable nitrogen that are deposited in precipitation have
the capacity to change multiple aspects of ecosystem nitrogen retention, sequestration, and processing.
The changes observed are always deleterious in that they lead to greater concentrations of nitrate-N
and thereby make more available for leaching to surface and groundwater. As outputs of nitrogen to the
atmosphere can reasonably be expected to increase in the foreseeable decades, it is prudent to identify
and develop management options now to both restore ecosystems that are already compromised and
to buffer effects to ecosystems that are at risk from new nitrogen inputs.
Keywords:
Atmospheric Deposition, Bioavailable Nitrogen, Nitrate, Ecosystem Response,
Ecosystem Management, Trophic Interactions
IV
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Contents
Foreword iii
Abstract iv
Keywords iv
Contents v
Acknowledgments x
Introduction 1
Soil Nitrogen Chemistry 2
Bioavailable Nitrogen 2
Microbiology; Patterns and Constraints 2
Nitrogen Leaching 3
Plant Communities 3
Primary Consumers - Herbivory 3
Litter Decomposition - Detritivory 4
Study Site 5
Methods 7
Soil Nitrogen Chemistry 7
Bioavailable Nitrogen 7
Microbiology; Patterns and Constraints 7
Microbial Response to Applied N 7
Denitrification Assays 7
Gross N Transformation Rates and Soil Moisture 8
Nitrogen Leaching 8
Plant Communities 8
Primary Consumers - Herbivory 8
Population Ecology 8
Physiology 9
Litter Decomposition - Detritivory 9
Results 11
Soil Nitrogen Chemistry 11
Bioavailable Nitrogen 11
Microbiology; Patterns and Constraints 12
Microbial Response to Added N 12
Denitrification Assays 12
Gross N Transformation and Mineral N 13
Moisture Effects 14
Nitrogen Leaching 15
Plant Communities 16
Primary Consumers - Herbivory 20
Population Ecology 20
Physiology 20
Litter Decomposition - Detritivory 22
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Discussion 25
Soil Nitrogen Chemistry 25
Bioavailable Nitrogen 25
Microbiology; Patterns and Constraints 25
Microbial Response to Added N 25
Denitrification Assays 25
Nitrogen Leaching 26
Plant Community 26
Primary Consumers - Herbivory 27
Population Ecology 27
Physiology 27
Litter Decomposition - Detritivory 27
Implications 29
References 31
VI
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Figures
Figure 1. These data, adapted from Wedin and Tiiman (1996), indicate that most of the
ecological response associated with nitrogen exposure occurs in the first
100 kg N ha~1yr1 of application 2
Figure 2. Percent extractable soil nitrate-N content relative to control treatment 11
Figure 3. Both fertilization (+Fertilized) and increased small mammal density
(+Fenced) increased soil nitrate-N 12
Figure 4. Soil nitrate chemistry changed quantitatively with nitrogen dose; while the
amount of nitrate-N measured in the soil increased with dose, the increase
remained proportional to the input 12
Figure 5. Denitrification in soils provided with ad libitum amounts of nitrogen was
greatest in soils collected from control plots 13
Figure 6. Denitrification in soils provided with ad libitum amounts of carbon was
greatest in soils collected from fertilized plots, especially during wet
spring periods 13
Figure 7. Gross N mineralization, nitrification, ammonium consumption, and nitrate
consumption rates in old-field soil measured under field conditions 14
Figure 8. Gross mineralization and NH/ consumption rates (A) and gross nitrification
and NO3 consumption rates (B) from old-field soils under three different water
potentials measured under laboratory conditions 15
Figure 9. Plant biomass production in the control and fertilized treatments 16
Figure 10. Site scores for the first two axes of the RDA following in an Oklahoma old-field
from 1999 (•) to 2001 (O) 17
Figure 11. Correlation between Festuca canopy cover and functional group canopy cover
on an Oklahoma old field from 1999 to 2001 18
Figure 12. Plot of the correlation between the change in Festuca canopy cover and the
change in warm-season native canopy cover from 1999 to 2001
(P = 0.0410, r = -0.56) 19
Figure 13. Plot of species richness as a function of litter mass 19
Figure 14. Estimates and standard errors (±1 SE) of minimum number known alive (MNA) for
Sigmodon hispidus across a landscape manipulated with fertilization and enclosure
fencing at the Center for Subsurface and Ecological Assessment Research,
Pontotoc County, Oklahoma, 1999-2000 20
VII
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Figure 15. Estimates and standard errors (±1 SE) of minimum number known alive (MNKA) for
Reithrodontomys montanus across a landscape manipulated with fertilization and
enclosure fencing at the Center for Subsurface and Ecological Assessment
Research, Pontotoc County, Oklahoma, 1999-2000 21
Figure 16. Estimates of minimum number known alive (MNKA) for Reithrodontomys fulvescens
across a landscape manipulated with fertilization and enclosure fencing at the Center
for Subsurface and Ecological Assessment Research, Pontotoc County, Oklahoma,
1999-2000 21
Figure 17. In litter which was placed in plots that did receive fertilizer amendments, litter
nitrogen declined relatively slowly overtime 23
Figure 18. In litter that allowed access by all decomposers and detritivores, litter nitrogen
declined consistently over time 23
VIII
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Tables
Table 1. Total Nitrogen Inputs (kg N ha1 yr1) for Water Year 2002-2003 16
Table 2. Nitrogen Leaching below 90 cm for Water Year 2002-2003 (kg N ha1 yr1)3 16
Table 3. Mean (x) and Standard Error (+SE) of Species Richness and Festuca Canopy Cover
in an Oklahoma Old-Field (n=16) 17
Table 4. Range of Estimates of Potential Sigmodon hispidus Urinary and Fecal Nitrogen
Deposition under Experimental Conditions Observed in this Experiment and under
Conditions Reported Elsewhere in the Literature 22
IX
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Acknowledgments
The U.S. Environmental Protection Agency through its Office of Research and Development funded
the research described here both through in-house efforts and through Interagency Agreements with
the Biological Resources Division of the U.S. Geological Survey and as administered by the Oklahoma
Cooperative Fish and Wildlife Research Unit (Oklahoma State University, Oklahoma Department of
Wildlife Conservation, U.S. Geological Survey, and Wildlife Management Institute cooperating) and the
U.S. Department of Energy as administered by Oak Ridge Institute for Science and Education (ORISE).
ManTech Environmental conducted some of the chemical analyses under contract with the U.S. EPA.
It has not been subjected to Agency review and therefore does not necessarily reflect the views of the
Agency, and no official endorsement should be inferred. Particular thanks goto D.M. Leslie, Jr. (Oklahoma
Cooperative Fish and Wildlife Research Unit, USGS), J. Williams, S. Schmelling, and C. Hall (National
Risk Management Research Laboratory, U.S. EPA) for facilitating this work. We thank M. Bahm, J.
Bahm, C. Bilder, T. Bodine, K. Brazil, A. Burrow, M. Day, C. Deerenberg, B. Faulkner, S. Fuhlendorf,
J.Gray, M. Hamilton, P Hanson, D. Hutchings, J. Jones, S. Kovash, M. Leslie, L. Levesque, J. Lindsey,
R. Lochmiller, S. Lyon, K. McBee, L. Mulkey, H. Murray, L. Newby, D. Niyogi, R Nunn, T. Oppelt, M.
Palmer, M. Payton, C. Peterson, D. Pope, H. Purvis, A. Roper, G. Sampson, J. Sheffield, E. Webb, and
J. Wilson for their assistance, advice, and counsel.
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Introduction
Throughout the eastern United States, from the Front Range of the Rocky Mountains to the Atlantic Ocean, bioavailable
nitrogen has been falling in the rain since the industrial revolution (i.e., Smil, 1990; Vitousek et ai, 1997). In a trend
that is expected to continue, these additions have been increasing (Brimblecombe and Stedman, 1982; Galloway et ai,
1994; U.S. EPA, 1995; Vitousek et ai, 1997). Because nitrogen is frequently a limiting nutrient for plants and animals,
increased quantities of nitrogen in ecosystems alter competitive relationships among terrestrial and aquatic organisms.
Nitrogen, particularly nitrate-N, easily moves from terrestrial ecosystems into surface and groundwaters, including lakes,
streams, rivers, and estuaries (i.e., Baker, 1992; Kahl et ai, 1993; Peterjohn et ai, 1996). As nitrogen concentrates in
surface and groundwater sinks, increasingly frequent observations of undesirable effects associated with eutrophication,
algae blooms, hypoxia, and toxicity are observed (Kelly et ai, 1990; Likens, 1992; Glibert and Terlizzi, 1999). Today,
acid rain phenomena in North America are largely associated with excess nitrogen (Aber et ai, 1989; Gilliam et ai,
1996). Wedin and Tilman (1996) suggested that increasing amounts of nitrogen in the environment may be associated
with global warming and climate change (See also Vitousek et ai, 1997; Shaver et ai, 2000).
Excess nitrogen is not tightly retained by ecosystems, but is highly mobile (i.e., Vitousek et ai, 1997). It occurs in
ecosystems under a variety of guises (i.e., nitrogen species; N03, NH4, NO2, DON, TN, etc.), each of which varies in
mobility, potential for use by organisms, and expression in site biogeochemistry. Therefore, concern about nitrogen
management in ecosystems is focused not only on the amount of nitrogen present, but also its transport and cycling.
While the effects of large doses of nitrogen are well documented, it is only recently that attention has been focused
toward risks associated with chronic low-level exposure to nitrogen, such as that accompanying atmospheric deposition
(Likens, 1992; Jorgensen et ai, 2002; Jorgensen et ai, 2003). In order to weigh risks and assess management op-
tions, it is important that a thorough understanding of the interactions and transport of nitrogen in terrestrial and aquatic
ecosystems and the atmosphere be developed.
The nitrogen cycle is well studied. While many of the cycle's components are important to consider for nitrogen man-
agement, there are relatively few that interact closely with atmospherically deposited nitrogen. In this study, we were
most interested in those components that are directly affected by nitrogen deposition and their response. Wedin and
Tilman (1996) published data demonstrating that exposure to excess bioavailable nitrogen degrades ecosystems in a
number of notable ways; 1) retained nitrogen decreases with increasing exposure, 2) biodiversity declines, and 3) plant
C/N ratio declines. Of perhaps greater importance is the observation that most of the ecological response measured
by Wedin and Tillman (1996) occurs in the first 100 kg N ha1 yr1 of deposition. Elsewhere, others and ourselves have
termed this the "ecologically significant dose" (Figure 1) (Jorgensen et ai, 2003).
Whereas atmospheric deposition in the eastern United States is measured at approximately 10-30 kg N ha1 yr1 (National
Atmospheric Deposition Program (NADP-3)/National Trends Network, 2002), it is reasonable to believe ecosystems
throughout this region are affected by chronic exposure to excess bioavailable nitrogen. Such effects may be related
in part to Perakis and Hedin's (2002) observations concerning relative availability of organic vs. inorganic nitrogen
in South America. Ecosystems process deposition in a few ways. Initially, direct deposition to terrestrial systems is
processed differently than direct deposition to aquatic systems (note: the research described in this paper does not
consider deposition to aquatic systems). Deposition to terrestrial systems may either be processed by biota (i.e., plants
and/or microbes), or it may escape to aquatic systems through runoff, percolation, or it may in-part volatilize. Improved
nitrogen management will occur where the probability of interaction with biota is high. The probability of interacting
with biota is not constant or fixed, because biota may change in response to many environmental conditions. Some of
these conditions (i.e., temperature and precipitation) are essentially outside of the scope of management intervention.
However, many of the conditions are susceptible to management once their response is better understood.
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LJJ
250
200 -i
150
100
50
0 <
Dose of Ecological Significance
50
100
150
200
250
300
350
Nitrogen addition (kg N ha"1 yr"1)
Figure 1. These data, adapted from Wedin and Tilman (1996), indicate that most of the ecological response associated
with nitrogen exposure occurs in the first 100 kg N ha~' yr1 of application. We term this level of exposure the
"Dose of Ecological Significance."
Soil Nitrogen Chemistry
Bioavailable Nitrogen
Ultimately, it is excess bioavailable nitrogen in soil (i.e., bioavailable nitrogen present in amounts greater than can be
used by plants and microbes) that leaches to waterbodies and causes the undesirable effects previously identified (Mahli
and Nyborg, 1986; Luo et al., 2000). Fortunately, terrestrial ecosystems are nitrogen limited; thus they usually use a
greater proportion of the bioavailable nitrogen they receive. When terrestrial ecosystems are no longer nitrogen limited,
they leach nitrogen to aquatic systems. As terrestrial ecosystems lose their inherent tendency toward nitrogen limitation,
they will begin to express elevated concentrations of bioavailable nitrogen in the soil; nitrogen that is immediately avail-
able to waterbodies via runoff or leaching (i.e., Kahl et al., 1993; Wedin and Tilman, 1996; Peterjohn et al., 1996).
Microbiology; Patterns and Constraints
Soil microbial communities can remove nitrogen from terrestrial ecosystems through denitrification, thereby reducing
the potential for contamination (i.e., Mahli and Nyborg, 1986; Luo et al., 2000). Denitrification is the biogeochemical
process in which bacteria use nitrate instead of oxygen to produce energy during which nitrate is reduced to gaseous
nitrogen (NO, N2O, or N2 depending on oxygen concentration). Denitrification and the related process of nitrification
strongly affect soil N chemistry (Hutchinson and Davidson, 1993; Whitehead, 1995). During nitrification, nitrite (NO2)
is formed through bacteria mediated chemical decomposition of intermediates between ammonium (NH4+) and nitrate
(NO3-). Nitrite is an unstable product and may easily convert to N20 under anaerobic conditions (Wrage et al., 2001).
NO,
NO,
NO
N2O
Complete denitrification is regarded as beneficial because of its potential to reduce NO3 concentrations in soil. How-
ever, partial denitrification products (N2O and NO) have undesirable environmental effects and are potentially harmful
greenhouse gases (Bouwman, 1990; Duxbury and Mosier, 1993).
Denitrification is mainly enabled by bacteria (i.e., Denitrifiers) that are generally heterotrophic; they rely on organic com-
pounds as electron donors. Denitrification in soils has been shown to be limited by nitrogen availability (Groffman et al.,
1993; Jordan et al., 1998) and carbon availability (i.e., Schnabel et al., 1996; Ashby et al., 1998; Frank and Groffman,
1998). However, information on how these limiting factors fit into a larger ecological context is sparse, incomplete, and
complex (Silva et al., 2005[b]). Soil denitrification has been shown to be a spatially and temporally variable phenomenon
(e.g., Luoetal., 2000; Frank and Groffman, 1998; Jordan ef al., 1998). Further, denitrification can also be constrained by
oxygen supply, temperature, soil moisture, soil pore status, soil depth, and pH. Due to the complexity of denitrification,
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its importance to agriculture, greenhouse gas emissions, and a widespread desire to better understand soil N chemistry,
investigators have studied it throughout the world (i.e., Stanford et al., 1975; Westerman and Tucker, 1978; Firestone,
1982; Knowles, 1982; Starr and Gillham, 1993; Weier et al., 1993a). These studies, and others, have demonstrated
the high degree of apparent variability for soil nitrogen transformation processes. Such variability seriously complicates
efforts to develop applied products and management recommendations. Better understanding and characterization of
the extent and variability of soil N chemistry will aid development of N management strategies in the future.
Nitrogen Leaching
Nitrogen leaching, in the context of this research, is the movement of inorganic nitrogen (i.e., nitrate-N) from upper soil
horizons where processing by either plant or microbial biota occurs into lower soil horizons (essentially below plants' root
zones; about 50 cm). Inorganic nitrogen that successfully passes through this region will eventually reach groundwater
and be released to surface waters (i.e., Baker, 1992; Kahl etal., 1993; Peterjohn et al., 1996) where, as already noted,
undesirable (and noticeable) consequences occur.
Plant Communities
Old-fields have been used extensively as model ecosystems for the investigation of ecosystem N dynamics and ef-
fects (Christensen and MacAller, 1985; Kalisz, 1986; Pastor et al., 1987; Robertson et al., 1988; Dormaar et al., 1990;
Gross et al., 1995). Soil properties, including nitrogen (Wedin and Tilman, 1990; Knops and Tilman, 2000), soil organic
matter (Zedler and Zedler, 1969), phosphorus, potassium, calcium, magnesium, and pH (Kalisz, 1986) change during
succession. For example, nitrogen is usually the most limiting nutrient to plant growth during the first 40 to 60 years of
old-field succession (Gleeson and Tilman, 1990), and through time, resource limitation can shift from nitrogen to light
availability (Tilman, 1988). It is thought that early-successional species are best adapted to low nitrogen availability,
whereas late-successional species are associated with more elevated nitrogen levels (Tilman, 1987).
As an ecological effect, nitrogen enrichment seldom occurs in isolation. More frequently, multiple effects occur together.
In landscapes where past use has been intensive, as in the eastern United States, current ecological effects are often
influenced by the past uses of the land. Reduction or elimination of grazing is frequently used as an early restoration
intervention technique, especially in riparian areas. Such landscapes frequently have a history of intentional and/or
incidental release and proliferation of non-native species.
In temperate mosaic grasslands, removal of heavy grazing can result in an increase of late-successional warm-season
grasses (Freeman, 1998; Engleefa/., 2000), litter accumulation, and a decrease in abundance of non-dominant grasses
and forbs (Weaver, 1968; Knapp ef al., 1998). Accumulation of litter in the absence of grazing can result in decreasing
species richness (Collins 1987; Carson and Peterson 1990; Foster and Gross 1998) which Wedin and Tilman (1996)
associated with reduced nitrogen use efficiency. Moreover, introduction or presence of non-native species may alter
secondary succession in grasslands following cessation of grazing (Tremmel and Peterson, 1983; Fike and Niering,
1999), further effecting soil N processing.
Tall fescue (Festuca arundinacea) is an invasive perennial grass native to Eurasia (Gibson and Newman 2001). Wide-
spread use of tall fescue for forage, turf, and soil conservation purposes began in the 1940s, and fescue gained status
as a commonly planted species in the eastern United States (Ball et al., 1993; Hoveland, 1993). Although fescue is
considered by pastoralists to be a valuable forage species in planted pastures, native ecosystems lacking disturbance
may be at risk of fescue invasion resulting in fescue becoming a transformer species (Richardson et al., 2000).
Studies of tall fescue have focused on species richness and plant-soil interactions of endophyte-infected fescue viz
endophyte free fescue (Clay and Holah, 1999; Matthews and Clay, 2001). Our study, on the other hand, provided the
opportunity to investigate the overall effect of relatively small amounts of fescue in an old-field ecosystem: an old-
field ecosystem released from cattle grazing and experimentally exposed to low-levels of bioavailable nitrogen on an
on-going basis.
Primary Consumers - Herbivory
Nutrient enrichment, particularly nitrogen input, affects trophic interactions and nutrient cycling. In irrigated shortgrass
prairie, nitrogen supplementation converted experimental plots to "islands of tallgrass" in the shortgrass landscape (Grant
ef al., 1977). Enrichment with nitrogen-rich sludge or fertilizer treatments in old-field communities caused decreases
in population density, recruitment, and survival of meadow voles (Microtus pennsylvanicus; Hall et al., 1991). Primary
consumers can have significant top-down effects on the cycling of energy and nutrients through ecosystems, especially
grasslands (Gessaman and MacMahon, 1984; McNaughton, 1985). These effects range from physical damage and
thrash, deposition of feces and urine to change nutrient status for the plant community composition, seed dispersal
and soil impacts (Gessaman and MacMahon, 1984; Heske et al., 1994; Silva et al., 2005a). In this way, modifications
in herbivore assemblages may interact with plant communities to amplify the effects of bioavailable nitrogen (Vitousek,
1994).
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Litter Decomposition - Detritivory
Plant litter decomposition is a key process in C and N cycles in terrestrial ecosystems. Litter decomposition is influ-
enced by various factors including litter quality, detritivore diversity, and nutrient availability (Swift et al., 1979; Aber and
Melillo, 1991; Schlesinger, 1997). For example, inorganic N inputs may alter microbial decomposition by changing N
availability and eliminating microbial nutrient limitation or, conversely, inhibiting decomposition (Carreiro et al., 2000).
Detritivore and decomposer diversity can influence decomposition rates (Mikola and Setala, 1998; Van der Heijden et
al., 1998; Naeem et al., 2000; Hobbie and Vitousek, 2000) where the presence of diverse macroinvertebrate functional
groups allows more efficient litter processing. Quantifying the influence of factors controlling decomposition is critical
to understanding litter dynamics and developing better models of nitrogen flux in ecosystems.
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Study Site
This study was conducted in southeastern Oklahoma at the Center for Subsurface and Ecological Assessment Research
(CSEAR), operated by U.S. EPA, Robert S. Kerr Environmental Research Center, Ada, Oklahoma, USA. CSEAR is
located in an area of interspersed old-field and oak-forest patches characteristic of the Cross Timbers ecotone, histori-
cally, a mosaic of mixed grasslands and oak-dominated forest between the southern Great Plains and eastern deciduous
forests of Texas, Oklahoma, and Kansas, USA (Hoagland ef a/., 1999). Cultivation at the CSEAR was abandoned ca.
1950, and cattle grazing was halted in 1998. The soil has been classified as clay loam (Vertic Argiustolls, USDA).
Within a contiguous old-field, sixteen 40 x 40-m plots were established to investigate ecosystem interactions associ-
ated with additions of low-level nitrogen and manipulations of herbivore populations. Plots were separated by creating
a 5-m mowed pathway. N availability and mammals were manipulated in plots in a randomized factorial experimental
design such that 4 plots each received fertilizer only, mammal manipulation only, a combination of fertilizer and mam-
mal manipulation, or neither (control). Plots were fertilized with granular 34% ammonium nitrate at an annual rate of
16.3 kg N ha1 yr1 beginning in February 1999 and every 3 months thereafter for 5 years. Mammal populations were
manipulated by a 2-m tall chain-link fence of 2.5-cm mesh that effectively excluded intermediate to large sized mammals
such as white-tailed deer (Odocoileus virginiana), armadillos (Dasypus novemcinctus), rabbits (Sylvilagus floridana),
coyote (Canis latrans), gray fox (Urocyon cinereoargentious), and striped skunk (Mephitus mephtus) while supporting
greater abundance of cotton rats (Sigmodon hispidus) and mice (i.e. Reithrodontomys montanus) inside fenced plots.
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Methods
Soil Nitrogen Chemistry
Bioavailable Nitrogen
Soil extractable nitrogen was measured in the 16 experimental plots three times per quarter for 5 years. Two soil samples
(separated by 10-m) were used from each plot. Samples were processed within 24 hours of collection. Samples were
hand picked of gross contaminants (i.e., plant material, worms, stones) and homogenized prior to analysis. Two sub-
samples from each homogenized sample were extracted with 2-M KCI using a soil to extractant ratio of 1:2. Extraction
procedures included shaking soil slurry for 1 hour, centrifuging at 1800 rpm for 10 minutes at 15°C (Scharf and Alley,
1988) and filtering using pretreated (with 2-M KCI and deionized water) Whatman 42 filter papers (Sparrow and Masiak,
1987). Extractions were completed within two days of sampling. Filtrates were analyzed for mineral-N using LACHAT
QuikChem flow injection analyzer (FtA).
In order to better understand the processes that account for variability in soil nitrogen and litter decomposition (i.e.,
concentrated excretion spot, fine-scale nitrogen transformation processes, local environmental conditions such as soil
moisture and porosity), we complemented studies on our primary old-field plots with related short-term field and labora-
tory studies using the same soils that were taken close to (but outside of) the already described 0.16-ha experimental
plots. Details of the experimental techniques for these experiments are delineated below.
Microbiology; Patterns and Constraints
Microbial Response to Added N .- Topsoil (0-10 cm) was collected from an adjacent old-field and homogenized by air-
drying and passed through a 4-mm sieve and mixing. Microcosms were prepared in plastic cups with each microcosm
containing approximately 290 grams of homogenized soil. Five treatments including a 1) control with no N applied,
2) 100 kg N ha1, 3) 200 kg N ha1, 4) 500 kg N ha1 and 5) 1000 kg N ha1 were established in randomized design with
3 replicates. Nitrogen was applied as 34% ammonium nitrate. Soil water potential in the microcosms was maintained
every alternate day by adding water to bring the microcosms to field capacity. All microcosms were sampled on the
following intervals; day 0 (prior to N application), 14, 28, 64, and 130. Soil moisture in the samples was determined gravi-
metrically by drying soils to a constant weight at 104-105°C. A 1:2 (soil to reverse osmosis water) soil extract was prepared
for pH measurement, followed by analysis by EPA method 150.1. Soil mineral nitrogen was measured using Lachat FIA
as described under bioavailable nitrogen section.
Denitrification /Assays.- Soil samples were taken from field plots set up for experimental nitrogen and herbivore ma-
nipulations (previously described). Denitrification of these samples was examined during February 21 and 23, June 28,
and July 19 and 24 of 2000. On dates when these plots were sampled, three soil samples were taken from each plot
using segments of PVC pipe (5.5 cm diameter x 7 cm depth). One of the intact soil cores was used for assessment of
denitrification rates by placing the pipe with soil into a 1-L jar, and using the acetylene block method (i.e., Groffman et
a/., 1999). The other two soil samples from each plot were composited, and subsamples were used to measure CO2
production, soil moisture, soil NH4+, and NO3 contents.
The main focus of our study was to examine the impact of N on denitrification; however, alteration of nitrogen may
change plant composition as well as soil carbon status. Thus, in our study we also included carbon treatment to study
its impact in comparison to nitrogen. To measure denitrification potential, both carbon (3-mg C g1 soil) and nitrogen
(0.3-mg N g1 soil) were added. Denitrification was measured by placing soils (10-g dw) in serum vials using the acety-
lene block method. Similarly, carbon dioxide production was also measured by placing soils (10-g dw) in serum vials
(160 ml) and monitoring the accumulation of CO2 in the headspace with a TCD (thermal conductivity detector) gas
chromatograph. Gravimetric soil moisture was determined by drying soils to a constant weight at 100°C. Soil ammo-
nium-N, and nitrate-N were determined using a Lachat FIA. Additionally, net nitrification and net nitrogen mineralization
were determined by incubating composited soils (20-g dw) in water-tight (130 mL) containers at room temperature and
measuring the change in soil nitrate-N and total inorganic-N, respectively.
-------
Gross N Transformation Rates and Soil Water.- Intact soil cores (4-cm diameter x 10-cm depth) collected from 4 ran-
domly selected locations from an adjacent old-field were used to measure gross N mineralization and nitrification rates.
Intact soil cores were collected on five occasions (between May and July 2002) at 2, 4, 6, 8, and 12 weeks. 15N dilution
technique (Di etal., 2000) was used to measure gross mineralization and nitrification rates. Silva et a/., (2005b) pro-
vided additional detailed descriptions of these methods. Since our N application rate was low (16.3 kg ha~1 yr1), there
is a very high probability to collect intact cores (4 cm diameter) with greater variability in fertilized plots; therefore, we
did not sample from fertilized plots for this short-term experiment.
Additionally, the effect of various soil moisture contents on N transformation process was examined using sieved old-
field soil (with no-fertilizer) placed in lysimeters and compacted to achieve original field bulk density (1.4 g cm'3). Water
potentials of 0 kPa, -0.5 kPa, and -1.0 kPa were maintained using a vacuum gauge and a vacuum pump, with water
being applied to laboratory soils in concert with precipitation events. Specific water potentials were applied to the soil
through a ceramic cup, which was attached to the bottom of each lysimeter. Lysimeters were incubated in the labora-
tory at a constant temperature of 25°C, and soil collected from lysimeters was analyzed at 2, 4, 6, 8, and 12 weeks to
determine the gross mineralization and nitrification rates at different water levels as they occur under field conditions.
Nitrogen Leaching
Nitrate-N, ammonium-N, and total Kjeldahl N (TKN) concentrations were measured in soil water and rain water. Soil
water was collected in ceramic cup tension lysimeters installed at approximately 90 cm below the soil surface in each
of the plots. Lysimeters were evacuated to 50 centibars and monitored every two weeks to check for water and to
re-apply vacuum. Three precipitation collectors consisting of an open funnel at 2 m height connected to a collection
vessel were placed across the site. Water, when present, was collected from the lysimeters and the precipitation col-
lectors without filtering. Precipitation samples were occasionally fouled by bird droppings and were discarded. All
samples were stored in a freezer to await analysis using a Lachat FIA. Dissolved organic N (DON) was calculated as
TKN minus ammonium-N, because the total Kjeldhal digestion method (H2S04 and HgO mixture) excluded reduction
of nitrate compounds to ammonium. Total N is the sum of Nitrate-N, ammonium-N, and DON.
Nitrogen fluxes were calculated by multiplying nitrogen concentrations by water flux volume. For rainfall inputs the
water-flux volume measured by the on-site meteorological station rain gauge was used to calculate the nitrogen flux.
Soil water flux volume was calculated using a simple model that used daily soil matric potential and precipitation data
to estimate water volume. On days where the soil matric potential was greater than -33 kPa (field capacity), the previ-
ous day's precipitation was assumed to be available for leaching. The total leachable water was summed for each
two-week period prior to lysimeter sampling. This model is not precise, but provides a good estimate of the volume of
water available for leaching and is not biased toward any treatment over another. It might slightly over-estimate fluxes
during heavy rain events because runoff is not subtracted from rainfall inputs.
Plant Communities
We calculated average quadrat (0.1-m2) canopy cover for each species by plot from 1999 through 2003 and clipped
enclosed vegetation for biomass measurement in the same quadrat. We performed redundancy analysis (RDA) (ter Braak
& Smilauer 1998) on the species data with Festuca as the explanatory variable to investigate community composition.
RDA is an ordination technique used when there is a linear relationship between two variables. We also examined rela-
tionships between Festuca canopy cover and functional group canopy cover using Pearson's correlation coefficients (r).
Plant canopy by species was determined using quadrat sampling in May 1999, August 1999, May 2000, and August 2000.
Plant species were identified using twenty five 0.1-m2 quadrats per plot (Stohlgren ef a/. 1998; Jorgensen and Tunnel!
2001). Tunnell (2002) and Tunnell etal. (2004) provide additional detailed descriptions of these methods.
Primary Consumers - Herbivory
Population Ecology
We sampled small mammals with Sherman live traps (7.6 x 8.9 x 22.9 cm) for 3 consecutive days at 3-5 week intervals
from July 1999 to December 2000. Each plot consisted of 25 traps systematically spaced at 7-m intervals. We released
captured animals immediately after marking with a unique identifier. Identification was done without considering animal's
sex. We used minimum number known alive (MNKA; Krebs 1966) as an index to abundance for each plot at each
sampling period. We made statistical comparisons of MNKA between treatment plots using a 2-way analysis of variance
with repeated measures (PROC MIXED, SAS 1990). We fitted a multiple variance model and used the Kenward-Roger
approximation to calculate effective degrees of freedom (PROC MIXED, SAS 1990; Kenward and Roger 1997) and
used least-squared means separation tests for all significant main effects.
-------
Physiology
To determine nitrogen dynamics and requirements at various life phases, animals underwent a series of feeding trials
under laboratory conditions. A captive research colony was formed using wild-caught individuals trapped at various sites
in Oklahoma. Not all animals used for feeding trials were wild-caught; some experimental animals were the captive-
born progeny of the wild individuals. All free-ranging animals were captured using Sherman live traps (7.6 x 8.9 x 22.9
cm), following standards established by the Animal Care and Use Committee of the American Society of Mammalogists
(1998). Research subjects were housed at the Laboratory Animal Resources facility at Oklahoma State University for the
duration of the pretrial and experimental periods. Animals were housed individually (except during the breeding phase
of the reproduction trial) and kept at 20-25°C under a 12L12D cycle for the duration of the study. We operated under
Animal Care and Use Protocol 723, Oklahoma State University. Mice were housed in 28- x 18- x 13-cm wire-topped
plastic cages with corn-cob bedding, and cotton rats were housed in similar cages that were 48 x 25 x 20 cm.
Litter Decomposition - Detritivory
Live and dead grasses and forbs were collected for litter on 18 December 1998 and 18 February 1999 from immediately
outside of our experimental plots. Cut litter was turned by hand to form a homogenous mix. A known quantity of dried
litter weighing (mean + 1 SE) 9.568 g + 0.106 was placed inside a bag made from a 0.5 x 0.5 m piece of nylon mesh
secured at the top with a locking plastic tie. Coarse (6.35 mm) and fine (0.33 mm) meshes were used to make litter bags.
Fine mesh was intended to exclude macro-detritivores such as earthworms (Annelida), soil mites (Arachnida), insects
(Insecta), pill bugs (Isopoda), and snails (Gastropoda), whereas coarse mesh was intended to allow access to the litter
by all micro and macro-detritivores. A total of 480 litter bags was constructed, half of fine mesh and half of coarse.
-------
Results
Soil Nitrogen Chemistry
Bioavailable Nitrogen
Extractable nitrate-N concentrations never averaged more than 1 mg L1 during the growing season and were always
less than 2 mg L1 on control plots during the dormant season. The introduction of an additional 16.3 kg N ha~1 yr1
resulted in little or no excess N at Fall 99 (Figure 2). However, during the first dormant season, nitrate-N concentra-
tion on fertilized plots was 525% greater than expected (i.e., compared to control plots), and concentrations remained
elevated on average throughout the experiment (Figure 2).
Soil Nitrate: 3 Month Average Deviation
•o
£
u
ex
X
UJ
0
Q.
Year and Season
•+Fert/+Fen +Fert
+Fen
Figure 2. Percent extractable soil nitrate-N content relative to control treatment. The change from Fall 1999 to Winter
2000 and continuing throughout the experiment is notable because it shows that the ability of soil to adapt
to new nutrient inputs is compromised by prior inputs. Further, the magnitude of variation was relatively
small during the first year, but increased thereafter, as did the overall variability.
While most of the elevated soil nitrate-N response is attributable to fertilization, a moderate level of soil nitrate-N was
measured in the fenced only treatment. Further, the amount of excess nitrate-N measured on fertilized only and fenced
only treatments closely corresponds to the excess nitrate-N measured on fertilizer and fence combined treatment.
Therefore, these effects appear to be additive and occur independently. It is noteworthy to add that N measurements
on the combined treatment plots during spring exceeded the sum of inputs measured on fenced and fertilized plots
alone (Figure 3). This is likely attributable to N cycling from other seasonal sources (e.g., insects) and carryover of N
from the preceding winter.
11
-------
Excess Soil Nitrate
Summer Fall Winter
Season / MEAN for Entire Experiment
MEAN
I Expected • + Fenced • + Fertilized • Combined
Figure 3. Both fertilization (+Fertilized) and increased small mammal density (+Fenced) increased soil nitrate-N. The
cumulative effects of these appear to be additive (Combined) during most of the year.
In summary, soil nitrate-N levels averaged over 250% higher on fertilized plots viz control plots, and excess nitrate-N was
particularly low during fall. The concentrations attributable to fertilization were sufficient to cause observable leaching
through the root zone (see Nitrogen Leaching results) while those attributable to increased density of small mammals
alone were not. Even though the overall average results from small mammals alone were relatively low, individual ex-
cretion patches, latrines, or other local hot-spots can considerably deviate from this trend due to high nitrogen loading
as seen in our simulated N application experiment (Figure 4).
Changes in Soil Nitrate Concentration
12 28 64
Days of Exposure
130
Figure 4. Soil nitrate chemistry changed quantitatively with nitrogen dose; while the amount of nitrate-N measured in
the soil increased with dose, the increase remained proportional to the input. This suggests that microbial
responses to nitrogen additions may be relatively fixed.
Microbiology; Patterns and Constraints
Microbial Response to Added N.- Qualitatively, the biogeochemistry of soils treated with 100, 200, 500, and 1000 kg ha~1
of nitrogen responded similarly (Figure 4). Response followed load in the expected fashion, with progressively greater
nitrate-N concentrations being present in those microcosms that received greater nitrogen loads. Quantitatively, while
the biogeochemistry of the treatments responded differently, the difference was in the expected direction (i.e., greater
concentrations observed in those microcosms that received higher loading of nitrogen). However, it is important to
note that by the end of the experiment, the observed quantitative difference among the treatments was converging, a
point that will be discussed later.
Denitrification Assays.- Even though stimulation of denitrification by nitrogen tended to occur after major precipitation
events, denitrification was more often stimulated by carbon than by nitrogen (135 out of 161 composited soil samples).
As soils dried and warmed with the onset of summer drought, denitrification became carbon-limited. Denitrification was
most stimulated by nitrogen after a rain event in April 2000. At this time, denitrification was also highly stimulated by
12
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carbon. As soils dried and warmed in May and June 2000, denitrification became more exclusively limited by carbon.
By July, denitrification activity diminished altogether. Rain events of September and October 2000 renewed denitrifica-
tion activity, but unlike denitrification after the April rain events, this activity was stimulated only by carbon and not by
nitrogen. Anaerobic conditions enhanced denitrification in carbon-amended soils compared to aerobic treatments, but
they were not sufficient to stimulate denitrification without added carbon.
Response of denitrification to laboratory additions of nitrogen and carbon was altered by fertilization (Figures 5 and 6).
Carbon stimulated denitrification more than nitrogen in fertilized soils, whereas nitrogen stimulated denitrification more
in control soils.
o>
I
c
o
'•5
re
u
c
0)
Q
100
Denitrification in Nitrate-rich Soil
Early Late
Cool Season Growing Time
End
Figure 5. Denitrification in soils provided with ad libitum amounts of nitrogen was greatest in soils collected from control
plots. This means that these soils'ability to respond to new nitrogen inputs (ad libitum nitrogen addition) was
reduced by prior exposure to low-level fertilization.
Denitrification in Carbon-rich Soil
0>
4-1
re
Oi
c
O
"re
u
c
0)
Q
1000
800
600
400
200
0
Early Late
Cool Season Growing Time
End
Figure 6. Denitrification in soils provided with ad libitum amounts of carbon was greatest in soils collected from fertil-
ized plots, especially during wet spring periods. This reveals that these soils' natural tendency for carbon
limitation was even further enhanced when exposed to chronic low-level fertilization.
Gross N Transformation and Mineral N.- Gross N mineralization and NH4+ consumption rates changed slightly between
sampling times when compared to nitrification and NO3 consumption rates. Additionally, both consumption rates (NH4+
and NO3) were greater than gross mineralization and nitrification rates (Figure 7); however, differences were fairly
constant over the experimental period. Similarly, both ammonium-N (1.5 -2.0 mg kg-1) and nitrate-N (0.8 -1.0 mg kg-1)
concentrations did not change over the late-spring to mid-summer period, suggesting that internal N transformation
processes in this old-field are fairly consistent under control conditions, and as one process shifts even slightly, oth-
ers change to counterbalance the system. However, as bioavailable N increases with N additions, the balance of the
internal N transformation may alter (i.e., lower N consumption) due to limitations of other nutrients (particularly carbon)
leaving excess N in the system.
13
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en
.2 15-
co
v
'ro
c.
o
'S.
e
o
CO
c
(D
CO
O
25 n
2.0 -
1.0 -
0.5 -
00 -
-0.5
10-May 24-May 6-Jun. 20-Jun.
Time
18-Jul.
—•— Mineralization
—.— Ammonium Consumption
—f— Nitrification
—^*— Nitrate consumption
Figure 7. Gross N mineralization, nitrification, ammonium consumption, and nitrate consumption rates in old-field soil
measured under field conditions. Bars indicate the ± standard error of mean.
Moisture Effects.- Soil N transformation processes were also examined through variation of response in association
with water potentials. Average gross NH4+ consumption rates were greater than gross nitrogen mineralization rates
for a given water potential (Figure 8A), but the differences were constant as observed in intact soil cores under field
conditions. Nitrogen transformation rates increased as soil water potential decreased. Greater mineralization and NH4+
consumption rates from lower water potentials could be attributable to greater O2 availability as more soil pores dry with
decreasing water potentials. Such differences were not observed in nitrification and NO3 consumption rates; however,
the magnitude of effect varied during the course of the experiment. This could not be only a result of soil water poten-
tials, but also the effect of soil incubation processes (Figure 8). Our results also reveal that N mineralization is a robust
process when compared with nitrification.
14
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en
S
TO
C
g
'S.
E
I
z
o
01
N
-------
Table 1. Total Nitrogen Inputs (kg N ha'1 yr1) for Water Year 2002-2003
Nitrate-N
Ammonium-N
DON0
Total N
Control3
3.7
3.6
3.8
11.0
Fertilized Plots'3
11.8
11.7
3.8
27.3
Current year's rainfall = 73.1cm (100 cm annual average)
alnputs from rainfall
"Fertilizer inputs (16.3 kg N ha'1 yr1 as NH4NO3) plus rainfall inputs
c Dissolved Organic N from total Kjeldahl N minus ammonium N.
Table 2. Nitrogen Leaching Below 90 cm for Water Year 2002-2003 (kg N ha"1 yr1)
Treatment
Fert x Fence
Fert Only
Fence Only
Control
NO3
Leaching Flux
8.2
12.8
0.5
0.1
NH4+
Leaching Flux
0.1
0.3
0.1
0.2
DON"
Leaching Flux
0.7
2.3
2.0
1.0
Total N
Leaching Flux
9.0
15.4
2.6
1.3
Total water leached: 38.7 cm
a Calculated from nitrogen concentrations in solutions collected bi-weekly from ceramic cup tension
lysimeters multiplied by leachable water (precipitation onto soil at or above field capacity) over the
same time period.
"Dissolved Organic N from Total Kjeldahl N minus ammonium
Plant Communities
Although no differences in total canopy cover were observed between fenced and control treatments, we observed a
slightly positive nitrogen effect on total biomass in later years. Further, even after 5 years of grazing release, biomass
continued to increase including in the control plots (Figure 9).
Total Plant Biomass
O)
1998
1999
2000
2001
2002
2003
Year
—• - Control
1 + Fertilizer
Figure 9. Plant biomass production in the control and fertilized treatments. There was no measurable difference
between control and fenced plots over the experimental period.
16
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Festuca cover increased nearly five-fold on average across plots from 1999 to 2001 (Table 3), which is reflected along
RDA Axis 1. Further, vegetation dynamics in this old-field were explained by increasing abundance of Festuca. The
change in species composition within plots is visually represented in the ordination diagram in which plots with the
greatest amount of Festuca (>10% canopy cover) are located in the right half of the diagram (Figure 10).
Table 3. Mean (x) and Standard Error (±SE) of Species Richness and Festuca Canopy Cover in an Oklahoma Old-
Field (n=16). There was No Treatment Effect on Species Richness
Species Richness
Festuca Canopy Cover (%)
Year
x
SE
X
SE
1999
2000
2001
44
37
41
±2
±1
±2
2.3
4.9
10.9
±0.6
±1.0
±2.0
(N
co
Plots with >10% Festuca canopy cover
RDA Axis 1
Figure 10. Site scores for the first two axes of the RDA following in an Oklahoma old-field from 1999(*)to2001 (O). RDA
axis 1 is represented by increasing Festuca canopy cover and has an eigenvalue of 0.135 (P=0.005).
Festuca was correlated with the two dominant functional groups, warm-season native grasses and non-legume forbs
(Figure 10). Festuca canopy cover was correlated negatively with warm-season native grass cover in 2001. Warm-
season native grass cover was not correlated with Festuca cover in the first two growing seasons when Festuca cover
was less (Figure 11). However, plots with the greatest warm-season native grass canopy cover had the least amount
of Festuca cover in 2001, and plots with low warm-season native grass canopy cover had the greatest amount of Fes-
tuca cover (Figure 11).
17
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45
9T 40
fc 35
6 30
& 25
O 20
15
10
5
0
O
10
15
20
25
fe
0
O
a
c
O
O
0
PH
1
0)
J
O
Z
45 -,
40 -
35 -
30 -
25 -
20 -
15 -
10 -
5 -
0 -
^ ^
"
•»• • " "*" ^
• ^ \^^^ ' 3 ° " "
^^£~ " <=>* -r
ft ° ° ^ o ^°
°=
0 5 10 15 20 25
Figure 11. Correlation between Festuca canopy cover and functional group canopy cover on an Oklahoma old field from
1999 to 2001. a. Correlation between Festuca canopy cover and warm-season native (C4N) grass canopy
cover in 1999 (P = 0,1305), 2000 (P = 0.7644), and 2001 (P = 0.0020, r = -0.25). b. Correlation between
Festuca canopy cover and non-legume forb canopy cover in 1999 (P = 0.0335, r = 0.45), 2000 (P = 0.1176),
and 2001 (P = 0.4915). • = 1999, o = 2000, and T = 2001.
The relationship appears to be causal in that, on the average, warm-season native grasses decreased most on plots
where Festuca increased (Figure 12).
Species richness (change in canopy cover) was negatively related to Festuca canopy cover (Figure 12), but species
richness was not correlated to litter mass (Figure 13).
18
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15
O
u
O
G
03
U
U
G
10 -
5 -
0 -
-5 -
J3
U
-10
5 10 15 20
Change in Festuca Canopy Cover (%)
25
Figure 12. Plot of the correlation between the change in Festuca canopy cover and the change in warm-season native
canopy cover from 1999 to 2001 (P = 0.0410, r - -0.56).
OH
C/3
70
60
50
40
30
20 J
100
200 300
Litter Mass
400
500
600
c
o
5
M
U
n.
GO
70
60
50
"3 40
30
20
10
15
20
25
Festuca Canopy Cover (%)
Figure 13. Plot of species richness as a function of litter mass. Regression models were not significant in 1999
(P = 0.3616), 2000 (P = 0.5560), and 2001 (P = 0.6913). • = 1999, O = 2000, and V = 2007. b. Plot of
species richness as a function of Festuca cover. Regression model in 1999 was not significant (P=0.1069).
Regression model in 2000 was significant (P = 0.0003), with Y = 42.3- 1.1X where Y = species richness
and X = Festuca canopy cover; R2 = 0.62. Regression model in 2001 was significant (P = 0.0039), with
Y = 46.6- 0.5535X where Y = species richness and X = Festuca canopy cover; R2 = 0.46. • = 1999,
o o = 2000, and V—V = 2001.
19
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Primary Consumers - Herbivory
Population Ecology
Between July 1999 and December 2000, we recorded 7,955 small-mammal captures in 20 sampling periods (i.e.,
24,000 potential trap nights). Cotton rats (Sigmodon hispidus) accounted for 5,468 (68.8%) captures; two species of
harvest mice accounted for 1,971 captures (Reithrodontomys montanus N=1229 [15.5%]; Reithrodontomys fulvescens
N=742 [9.3%]).
We observed 2-way interactions for abundance of cotton rats between the fenced treatment and time (F15170 = 1.91,
P = 0.024) and between fertilizer and fence treatments (F, 243 = 10.87, P = 0.003). Abundance of cotton rats tended to
be higher on fertilizer-fenced plots (x = 18.4, SE = 0.8, P < 0.001) compared with other treatments (control: x = 9.8,
SE = 1.0; fenced only: x = 11.2, SE - 0.8; fertilized only: x = 9.7, SE = 1.0; Figure 14). Abundance of R. montanus
tended to be higher on fertilized-only treatment, but lowest on fertilizer-fenced treatment (3-way interaction: nitrogen x
fence x time, F15171 = 2.22, P = 0.007; Figure 15). We observed no distinct patterns in relation to the treatment plots
for abundance estimates off?, fulvescens (Figure 16).
_o
D.
8 2 22 20 18 21 11 9 6 5 2 5 22 29 28 20 10 8 29 3
JLAASONDJ FMAMMJNJLASOOD
99 99 99 99 99 99 99 00 00 00 00 00 00 00 00 00 00 00 00 00
Sampling date
• Control —o— Fence --•-- Fertilizer—a- Fertilizer/Fence
Figure 14. Estimates and standard errors (±1 SE) of minimum number known alive (MNKA) for Sigmodon hispidus
across a landscape manipulated with fertilization and enclosure fencing at the Center for Subsurface and
Ecological Assessment Research, Pontotoc County, Oklahoma, 1999-2000.
Physiology
Estimates of mean daily urinary and fecal nitrogen deposition for an average individual Sigmodon hispidus were of
100.9 mg day1 (assuming an average weight of 150 g). There are many factors to consider when extrapolating this
number to calculate an estimated deposition for each plot. For this presentation, a range of potential outcomes are given
that assume constant Sigmodon hispidus density for an entire year. Thus, the intent is to estimate a range of outcomes
including the most extreme case while providing evidence that may more closely reflect actual conditions.
20
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18
16
14
~ 12
_o
9: 10
2
0
8 2 22 20 18 21 11 9 6 5 2 5 22 29 28 20 10 8 29 3
JLAASONDJ FMAMMJNJLASOOD
99 99 99 99 99 99 99 00 00 00 00 00 00 00 00 00 00 00 00 00
Sampling date
-•- Control —c^-Fence - •-- Fertilizer —a- Fertilizer/Fence
Figure 15. Estimates and standard errors (± 1 SE) of minimum number known alive (MNKA) for Reithrodontomys mon-
tanus across a landscape manipulated with fertilization and enclosure fencing at the Center for Subsurface
and Ecological Assessment Research, Pontotoc County, Oklahoma, 1999-2000.
o
Q_
Z
^.
7.0
6.0
5.0
4.0
3.0
2.0
1.0
0.0
n-T
8
JL
99
2 22 20 18 21 11 9 6 5 2 5 22 29 28 20 10 8 29 3
AASONDJ FMAMMJNJLASOOD
99 99 99 99 99 99 00 00 00 00 00 00 00 00 00 00 00 00 00
Sampling date
• Control —o—Fence • Fertilizer—n—Fertilizer/Fence
Figure 16. Estimates of minimum number known alive (MNKA) for Reithrodontomys fulvescens across a landscape
manipulated with fertilization and enclosure fencing at the Center for Subsurface and Ecological Assess-
ment Research, Pontotoc County, Oklahoma, 1999-2000. Standard errors were not included because no
significant differences (P > 0.05) were detected between treatment combinations.
21
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In our study, we observed a maximum abundance of Sigmodon hispidus of 200 ha1 during October of 2000 (minimum
abundance was 0 ha'1) (note: abundance relative to particular experimental treatments has already been given). This
means that Sigmodon hispidus may have been depositing up to 32.6 kg ha'1 yr1 of nitrogen back to the plots as urine
and feces (Table 4).
Table 4. Range of Estimates of Potential Sigmodon hispidus Urinary and Fecal Nitrogen Deposition under Experimental
Conditions Observed in this Experiment and under Conditions Reported Elsewhere in the Literature
Condition/Observation
Observations from this experiment
Maximum abundance
+Nitrogen +Fence
+Nitrogen - Fence
-Nitrogen +Fence
-Nitrogen -Fence
Nitrogen Mean
-Nitrogen Mean
Observations from literature
Jorgensen et al. 1994
Langley and Shure 1988
Schetter et al. 1998
Doorian and Slade 1995
Schetter et al. 1998
Stafford and Stout 1983
Doorian and Slade 1995
Fleharty et al 1972
Cameron 1977
Abundance
(#/ha)
200.0
115.0
60.6
70.0
61.3
87.7
65.6
244.0
119.0
111.0
100.0
90.0
46.9
39.5
20.6
14.0
Urine and Fecal Deposition
(kg ha"1 yr1)
32.6
18.8
9.9
11.4
10.0
14.3
10.7
39.8
19.4
18.1
16.3
14.7
7.7
6.4
3.4
2.3
Litter Decomposition - Detritivory
Statistical analysis reveals that litter decomposition was unaffected by either fertilization or exclusion of intermediate
and large mammals. All litter showed an increase in nitrogen concentration over time; however, nitrogen in litter placed
in fertilized plots displayed a reduced rate of nitrogen loss. Further, it appears that the litter would actually again be
accumulating nitrogen relative to its initial concentration after little more than a year of decomposition (Figure 17).
Change of nitrogen loss in litter is not only a function of fertilization. There are several other factors that can influence
litter decomposition and subsequently nitrogen loss. Detritivore exclusion altered litter nitrogen dynamics. After an initial
loss of nitrogen in both control and exclusion litter, nitrogen began to rapidly accumulate in exclusion litter. After little
more than 6 months of decomposition, nitrogen actually began to accumulate in exclusion litter (Figure 18).
22
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Effect of Low Level Fertilization
5? « "
(0
O IE
_j 13
C
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Discussion
These results demonstrate that even the relatively small amounts of bioavailable nitrogen that are deposited in precipita-
tion have the capacity to change multiple aspects of ecosystem function. These studies provide insight into the results
of Wedin and Tilman (1996), where most ecosystem effects attributable to exposure to excess bioavailable nitrogen
occurred in the first 100 kg ha~1 yr1 of deposition. In this study, changes among the responses of different trophic levels
to low-level nitrogen inputs were always in a direction that favored low nitrogen turnover and/or increased nitrogen flux,
both conditions that could contribute to Wedin and Tilman's (1996) observations. These characteristics directly lead to
increased levels of exposure to nitrogen and to other trophic levels that will in turn further reduce the ability to process
new nitrogen inputs. The changes observed are always potentially deleterious in that they lead to greater concentrations
of nitrate-N in soil and thereby make more nitrogen available for leaching to surface and groundwater.
Soil Nitrogen Chemistry
Bioavailable Nitrogen
In this experiment, soil nitrogen chemistry is the main focus of all of the trophic interactions that were investigated.
There needed to have been a change in soil nitrogen chemistry if excess nitrate-N were to be available for leaching.
While we expected this change to occur, we did not have a clear idea of how long it would take. In fact, it took only a
single growing season for changes in the soil's ability to adapt to nitrogen inputs to be evident.
Ultimately, it is excess bioavailable nitrogen in soil that leaches to water-bodies and causes the undesirable effects pre-
viously identified (Mahli and Nyborg, 1986; Luo et al., 2000). This study demonstrates that even regions of the eastern
United States that receive seemingly modest or small amounts of atmospheric nitrogen deposition may have undergone
a long-term change to their ability to process further nitrogen inputs. These data are consistent with Perakis and Hedin's
(2002) hypothesis that currently observed ecosystem biogeochemistry in much of the Northern Hemisphere may be the
product of an historical alteration to biogeochemical cycles that has not yet been identified or understood.
Our data suggest that the inherent capacity of terrestrial ecosystems to process bioavailable nitrogen is alterable by
low-level exposures; however, the long-term effects of such exposure remain uncertain.
Microbiology; Patterns and Constraints
Microbial Response to Added N.- We found no indication of a qualitative difference in soil biogeochemical response,
expressed as concentration of nitrogenous compounds, among control soils or those dosed with an equivalent of 100,
200, 500, and 1000 kg N ha'1 applied as 34% ammonium nitrate (Figure 4). Based upon the findings of Wedin and Til-
man (1996), this was somewhat surprising. An important difference between our study and that of Wedin and Tilman
(1996) was that our lowest load was an equivalent of 100 kg ha~1. For Wedin and Tilman (1996), ecosystem structure
and function were already seriously affected at this load. So, it could be that soil biogeochemical response differs
qualitatively with load at low doses (i.e., <100 kg ha~1), but not at higher doses. While our data are silent on this ques-
tion, the qualitative response we observed for our experimental doses leaves little room for such a response to occur.
Therefore, we expect that most of the exponential decay in ecosystem structure and function observed by Wedin and
Tilman (1996) is ultimately attributable to plant community dynamics.
Denitrification Assays.- Our results contrast with studies which found denitrification to be N limited (Luo et al., 2000;
Jordan et al., 1998; Groffman et al., 1993), but are similar to other studies which found denitrification to be limited by
carbon availability (Frank and Groffman, 1998; Luo et al., 1998; Schnabel et al., 1996). However, after rain events,
such as in April 2000, denitrification was sometimes also stimulated by nitrogen, indicating that carbon and nitrogen
can transiently co-limit denitrification depending on soil moisture conditions. Similarly, Ashby et al. (1998) found that
nitrogen was more important to denitrification when soils were wet and carbon was available.
Our data suggest that the ability of microbial consortia to conduct denitrification will be limited and that it will be affected
by prior exposure to nitrogen. The ability of microbial consortia to respond to new nitrogen inputs may already be ad-
versely affected throughout many areas of the eastern and central United States. We suggest that integrated ecological
25
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studies, including the consideration of the effects of plants, animals, and upland sites may be essential for constructing
predictive models of denitrification throughout a potentially wide area of the eastern and central United States.
Denitrification is less efficient than aerobic utilization of the same carbon compounds, so denitrification rates are highest
where soils are saturated and become anoxic. We measured denitrification in upland soils that are generally considered
aerobic, perhaps occurring in the anoxic microsites of these soils. Across the central United States, upland soils cover a
larger surface area than saturated soils. Therefore, although there is concentrated denitrifying activity in riparian zones,
hyporheic zones, and wetlands, denitrification in upland soils may be just as important regionally, especially considering
their large spatial extent (Ashby et al., 1998).
Nitrogen Leaching
Inorganic N deposition across the United States ranges between near 0 kg N ha'1 yr1 in west coast sites to 10 kg N ha~1 yr1
in sites in New England. Inorganic N deposition measured at our site was higher (7.3 kg N ha~1 yr1) than the 4.4 kg N ha1 yr1
ten-year average (1993-2002) of three National Atmospheric Deposition Program (NADP) sites in the region (sites OK17,
AR27, and TX56). This could be due to undetected other depositions (i.e. bird-related contamination), differences in
sampling protocol, or inter-annual variation. The NADP sites do not measure DON, and the results of this study dem-
onstrate that DON is also an important source of N inputs, making up about one third of the total at our site.
In unfertilized plots, only about 10% of rainfall N inputs were detected in leaching water; however, nitrate-N leaching
at our site increased as a result of ammonium nitrate additions. About two thirds of the annual addition of fertilizer
nitrogen was detected in leachable water as nitrate-N. Some of this nitrogen could have come from previous years' N
additions, because we did not monitor nitrate leaching in previous years. The relatively low rainfall during this water
year also could have contributed to higher nitrate-N leaching in fertilized plots in two ways: 1) Lower plant growth could
have lowered the demand for added N and 2) Drying of the vertic clays at the site may have increased macropore flow
by-pass because of the extensive cracking that developed.
Ammonium-N concentrations and flux were very low indicating either strong potential for nitrification or for strong retention
of ammonium-N in the cation exchange sites in the soil. Interestingly, DON appeared unaffected by fertilizer addition
indicating that inorganic fertilizer does not lead in a direct way to the formation of leachable DON.
Plant Community
Other than increased biomass, we did not observe changes to the plant community that could be explained by additional
exposure to the small amounts of bioavailable nitrogen we applied during the first two years. Additionally, following
release from grazing, changes to the plant community on this study site were complex, interacting with presence and
proliferation of a non-native species, tall fescue. However, changes to the plant community structure remain a virtual
certainty with higher N doses (Wedin and Tilman, 1996; Silva et al., 2005a).
Festuca increased and altered community composition and short-term succession without decreasing species richness
(Table 3), but this was intermittent and did not persist throughout the term of this study. Others have observed that when
Festuca is the major vegetation component, Festuca influences vegetation dynamics by decreasing species richness,
(Clay and Holah, 1999) and decreasing litter accumulation (Wieder et al., 1983).
Our study was conducted on a serai old-field with small amounts of Festuca, whereas other studies have examined
vegetation dynamics of Festuca monocultures or simple mixtures with the primary focus on endophyte-infected and
endophyte-free Festuca (Wieder era/., 1983; Clay and Holah, 1999; Matthews and Clay, 2001). Festuca is an invasive
and competitive species that overrides vegetation dynamics in monocultures and simple mixtures, so the possibility this
plant might dominate in old-fields remains a legitimate concern for ecosystem management and restoration.
In the absence of an invasive transformer, vegetation change within plots should represent chronosequences of species
composition from early- to mid- and late-successional species (Collins and Adams, 1983; Engle et al., 2000) consistent
with previous observations of similar old-fields following cessation of chronic intense grazing (Engle et al., 2000). However,
we observed a distinct separation among plots representing a difference in increasing abundance of Festuca canopy
cover (Figure 12). In the three years following cessation of heavy grazing, changes in species composition of plots were
expressed with increasing abundance of Festuca rather than increases in late-successional species. That is, increas-
ing Festuca canopy cover altered succession dynamics. Increases in warm-season grasses were expected to occupy
the major successional pathway in this grassland (Collins and Adams, 1983; Engle et al., 2000), but on plots in which
Festuca cover exceeded 10%, Festuca became the driver of succession in place of warm-season native grasses.
These relationships and potential interactions with nitrogen biogeochemistry warrant further investigation for 2 reasons;
1) the time and site specific limitations of our study are insufficient to definitively rule out relationships among nitrogen
availability and Festuca, 2) because of both the widespread nature of atmospheric nitrogen deposition and Festuca
distribution, ecosystem effects would be widespread.
26
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Primary Consumers - Herbivory
Population Ecology
Cotton rat density tended to be highest on fertilized-fenced treatment for a considerable time compared to the other
treatments. Population density of cotton rats in fertilized only, fenced only treatments was similar to control treatment
(Figure 14), but density fluctuated among treatments with significantly greater densities during winter from fertilized-
fenced treatments. Thus, we did not observe a fertilizer or fence effect on population densities, but rather a fertilizer-fence
interaction. The combination of increased above-ground live mass and protection from predation on fertilized-fenced
plots likely accounted for differences in population density of cotton rats among the treatment plots.
Predation has regulatory effects on population densities of small mammals (Desy and Batzli, 1989, Tait and Krebs,
1983). Enclosures that control access by predators have been used to address hypotheses and effects of predation
on population characteristics of small mammals (Schnell, 1968; Vaughan and Keith, 1981; Desy and Batzli, 1989). En-
closures at CSEAR were designed primarily to control herbivory by medium and large herbivores with less emphasis
on controlling predation. If our observations are in-part attributable to interactions between predation risk and nitrogen
availability, then the potential implications for nitrogen management of small or isolated landscapes where herbivore
population dynamics are perturbed become quite complicated and worthy of considerable future research.
Physiology
Deposition of urine and feces by Sigmodon hispidus may be a significant source of nitrogen for plants and microbes.
Further, when it is considered that there are many herbivorous consumer organisms present, the estimates associated
with a single species (albeit a potentially dominant and abundant one) definitely reflect under-estimates of the extreme
case. However, the fact that additional leaching was not observed on fenced-only treatments, coupled with the obser-
vation that soil nitrate-N (although elevated) was lower on fenced plots compared to fertilized plots, indicates that this
source of input itself is insufficient to drive ecosystem changes during early succession in old-fields.
In this experiment, we are able to present extreme case conditions associated with Sigmodon hispidus (Table 4) with
considerable certainty. Further, data collection and/or modeling may help us to better estimate actual field depositions
of nitrate and exposures to plants and animals associated with herbivory.
Litter Decomposition - Detritivory
This study suggests limits to which ecosystems can process N through the decomposer pathway. Specifically, both
the concentration and total amount of N remaining in litter increased after exogenous N addition, indicating that the
ecosystem had quickly reached a limit to its N processing capacity (Figure 17). Furthermore, when macro-detritivores
were excluded from litter, total N in the litter increased, eventually becoming an apparent sink for the N inputs, imply-
ing that a diverse detritivore community is better able to process N even under increased N loads. These observations
have 2 potential consequences for ecosystems. First, when N deposition (simulated here by fertilization) increases, the
resulting excess N contained in litter may eventually be available for transport through leaching and runoff (i.e., Kahl
et a/., 1993; Peterjohn et al., 1996) potentially adversely impacting ecosystems through eutrophication and acidification
(Kelly et al., 1990; Likens, 1992; Carpenter et al., 1998; Gilbert and Terlizzi, 1999). Reducing N inputs to the decom-
poser pathway points logically to source control of fertilizers and atmospheric N, which contribute significantly to the
global input of synthesized N currently produced at rates exceeding natural terrestrial N fixation (Galloway and Cowl-
ing, 2002). Second, the influence of macro-detritivores on N flux observed here has important implications for systems
where detritivore communities may be perturbed through habitat alteration or chemical use. For example, the ecological
modification that characterizes urbanized or agricultural systems is likely to adversely affect detritivore communities
and their function (Blair, 1999; Paoletti and Hassall, 1999). Direct exposure to chemicals and/or chemical drift may also
impact macro-invertebrates (Hershey ef a/., 1998) potentially reducing detritivore diversity and causing indirect effects
on litter dynamics and nutrient flux.
Terrestrial ecosystems with impaired ability to process N due to impacted detritivore communities may represent non-point
sources of N for ground and surface water. Resource management to reduce risks to ecosystem function (U.S. EPA,
1995) by improving the ability of ecosystems to retain, sequester, and process N may need to address detritivores and
their influence on litter biogeochemistry and nutrient flux (Seastedt, 1984; Beare et al., 1992; Griffiths, 1994), especially
if global inputs of reactive N increase at current rates (Brimblecombe and Stedman, 1982; Smil, 1990; Vitousek, 1994;
Vitousek et al., 1997, Galloway et al., 2002).
N concentration in litter increased throughout the experiment, but final concentrations did not differ between coarse and
fine mesh, a result consistent with studies on Eucalyptus litter (Reddy and Venkataiah, 1989); however, total N loss was
significantly different between the two mesh types (Figure 18). This paradox may be explained by the conservation of
mass in fine mesh bags relative to coarse. Little, if any, additional litter mass was lost in fine mesh bags after the eighth
month of exposure, whereas mass was continually lost from coarse mesh bags throughout the experiment; therefore,
27
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mesh effect on total N was a function of reduced mass loss from fine meshed bags rather than from an effect on N
concentration. Our results support the suggestion that detritivore diversity and species' characteristics (e.g. size, feeding
preference) are important factors in dictating decomposition processes (Beare et a/., 1992; Mikola and Setala, 1998).
The significance of macro-detritivore effects on N cycling may be a function of their large size relative to microbes and
fungi and thus, their ability to translocate large quantities of mass.
28
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Implications
This study provides evidence for several important phenomena:
1) Soil nitrogen chemistry was altered after one growing season of exposure to low-level nitrogen deposi-
tion.
2) The ability of the soil ecosystem to respond to nitrogen inputs can be compromised by previous expo-
sures.
3) Microbial denitrification is primarily carbon-limited during most of the year.
4) The ability of microbes to respond to nitrogen inputs may be largely fixed.
5) Soils that have had prior exposure to low level nitrogen inputs are more severely carbon-limited than soils
that have not been so exposed.
6) High abundance of herbivores, in combination with their deposition of urine and feces, can mimic the effect
of fertilization and increase exposure of the soil ecosystem to bioavailable nitrogen.
7) Herbivorous consumer communities can likely produce more than inconsequential amounts of bioavailable
nitrogen.
8) The nitrogen use efficiency of decomposer and detritivore pathways is deleteriously altered by exposure
to excess bioavailable nitrogen.
9) The nitrogen use efficiency of decomposer and detritivore pathways is deleteriously altered by perturbations
to those pathways.
10) Changes to ecosystems that are chronically exposed to low doses of bioavailable nitrogen are frequently
in a direction that tends to increase the flux of bioavailable nitrogen through the system, thereby increasing
the risk of deleterious ecosystem responses.
Ecosystems may be at risk from doses of atmospherically deposited nitrate that is generally considered to be low or
modest. As deposition of bioavailable nitrogen from the atmosphere can reasonably be expected to increase in the
foreseeable decades, it is prudent to identify and develop management options now to both restore ecosystems that
are already compromised and to buffer effects to ecosystems that are at risk from new nitrogen inputs.
We suggest that integrated ecological studies, including the consideration of the effects of plants, animals, and upland
sites, may be essential for constructing predictive models of watershed nitrogen risk and management throughout a
potentially wide area of the eastern and central United States.
29
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References
Aber, J.D., K.J. Nadelhoffer, R Steudler, and J.M. Melillo. 1989. Nitrogen saturation in northern forest ecosystems.
Bioscience 39:378-386.
Aber, J.D. and J.M. Melillo. 1991. Terrestrial ecosystems. Saunders College Publishing, Orlando, Florida, USA.
Animal Care and Use Committee. 1998. Guidelines for the capture, handling, and care of mammals as approved by
the American Society of Mammalogists. Journal of Mammalogy 79:1416-1431.
Ashby, J.A., W.B. Bowden, and RS. Murdoch. 1998. Controls on denitrification in riparian soils in headwater catch-
ments of a hardwood forest in the Catskill Mountains. Soil Biology and Biochemistry 30:853-864.
Baker, L.A. 1992. Introduction to nonpoint source pollution in the United States and prospects for wetland use. Eco-
logical Engineering 1:1-26.
Ball, D.M., J.F. Pedersen, and G.D. Lacefield. 1993. The tall-fescue endophyte. American Scientist 81: 370-379.
Beare, M.H., R.W. Parmelee, RF. Hendrix, W. Cheng, D.C. Coleman, and D.A. Crossley. 1992. Microbial and faunal
interactions and effects on litter nitrogen and decomposition in agroecosystems. Ecological Monographs 62:569-
591.
Blair, R.B. 1999. Birds and butterflies along an urban gradient: surrogate taxa for assessing biodiversity? Ecological
Applications 9:164-170.
Bouwman, A.F., 1990. Exchange of greenhouse gases between terrestrial ecosystems and the atmosphere. Pages
61-127 In Bouwman, A.F. (Ed.), Soils and Greenhouse Effects. John Wiley and Sons Ltd, Chichester.
Brimblecombe, P. and D.H. Stedman. 1982. Historical evidence for a dramatic increase in the nitrate component of
acid rain. Nature 298:460-462.
Cameron, G. N. 1977. Experimental species removal: demographic responses by Sigmodon hispidus and Reithro-
dontomys fulvescens. Journal of Mammalogy 58:488-506.
Carpenter, S., N.F. Caraco, D.L. Correll, R.W. Howarth, A.N. Sharpley, and V.H. Smith. 1998. Nonpoint pollution of
surface waters with phosphorous and nitrogen. Issues in Ecology, No. 3. Ecological Society of America, Washing-
ton, D.C., USA.
Carreiro, M.M., R.L. Sinsabaugh, D.A. Repert, and D.F. Parkhurst. 2000. Microbial enzyme shifts explain litter decay
responses to simulated nitrogen deposition. Ecology 81:2359-2365.
Carson, W.R and C.J. Peterson. 1990. The role of litter in a old-field community: impact of litter quantity in different
seasons on plant species richness and abundance. Oecologia 85: 8-13.
Christensen, N.L. and T. MacAller. 1985. Soil mineral nitrogen transformations during succession in the piedmont of
North Carolina. Soil Biology and Biochemistry 17: 675-681.
Clay, K. and J. Holah. 1999. Fungal endophyte symbiosis and plant diversity in successional fields. Science 285:1742-
1744.
Collins, S.L 1987. Interaction of disturbances in tallgrass prairie: a field experiment. Ecology 68: 1243-1250.
Collins, S.L. and D.E. Adams. 1983. Succession in grasslands: thirty-two years of change in a central Oklahoma
tallgrass prairie. Vegetation 51:181-190.
Desy, E. A., and G.O. Batzli. 1989. Effects of food availability and predation on prairie vole demography: a field ex-
periment. Ecology 70:411-421.
Doonan, T. J., and N. A. Slade. 1995. Effects of supplemental food on population dynamics of cotton rats, Sigmodon
hispidus. Ecology 76:814-826.
Dormaar, J.F, S. Smoliak, and W.D. Williams. 1990. Soil chemical properties during succession from abandoned crop-
land to native range. Journal of Range Management 43:260-265.
Duxbury, J.M. and A.R. Mosier. 1993. Status and issues concerning agricultural emissions of green house gases.
Pages 229-258 In Kaiser, H.M., Drennen, T.E. (Eds.), Agricultural Dimensions of Global Climate Change. St. Luice
Press, Forida, USA.
Engle, D.M., M.W. Palmer, J.S. Crockett, R.L. Mitchell, and R. Stevens. 2000. Influence of late season fire on early
successional vegetation of an Oklahoma prairie. Journal of Vegetation Science 1:135-144.
Fike, J. and W.A. Niering. 1999. Four decades of old field vegetation development and the role of Celastrus orbiculatus
in the northeastern United States. Journal of Vegetation Science 10: 483-492.
31
-------
Firestone, M.K., 1982. Biological denitrification. Pages 289-326 In Stevenson, F.J. (Ed.), Nitrogen in Agricultural Soils.
American Society of Agronomy. Madison, Wl, USA.
Fleharty, E. D., J. R. Choate, and M. A. Mares. 1972. Fluctuations in population density of the hispid cotton rat: factors
influencing a "crash." Bulletin of the Southern California Academy of Sciences 71:132-138.
Foster, B.L and K.L. Gross. 1998. Species richness in a successional grassland: effects of nitrogen enrichment and
plant litter. Ecology 79: 2593-2602.
Frank, D.A. and P.M. Groffman. 1998. Denitrification in a semi-arid grazing ecosystem Oecologia. 117:564-569.
Freeman, C.C. 1998. The flora of Konza Prairie: a historical review and contemporary patterns. Pages 69-80 In A.K.
Knapp, J.M. Briggs, D.C. Hartnett, and S.L. Collins. (Eds.) Grassland dynamics: long-term ecological research in
tallgrass prairie. Oxford University Press, Oxford, United Kingdom.
Galloway, J.N. and E.B. Cowling. 2002. Reactive nitrogen and the world: 200 years of change. Ambio 31:64-71.
Galloway, J.N., H. Levy II, and RS. Kasibhatla. 1994. Year 2020: consequences of population growth and development
on deposition of oxidized nitrogen. Ambio 23:120-123.
Gessaman, J.A., and J.A. MacMahon. 1984. Mammals in ecosystems: their effects on the composition and production
of vegetation. Acta Zoologica Fennica 172:11 -18.
Gibson, D.J. and J.A. Newman. 2001. Festuca arundinacea Schreber (F, elatior L. spp. arundinacea (Schreber) Hackel.
Journal of Ecology 89: 304-324.
Gilliarn, F.S., M.B. Adams, and B.M. Yurish. 1996. Ecosystem nutrient responses to chronic nitrogen inputs at Fernow
Experimental Forest, West Virginia. Canadian Journal of Forest Research 26:196-205.
Gleeson, S.K. and D. Tilman. 1990. Allocation and the transient dynamics of succession on poor soils. Ecology
71:1144-1155.
Glibert, RM. and D.E. Terlizzi. 1999. Cooccurrence of elevated urea levels and dinoflagellate blooms in temperate
estuarine aquaculture ponds. Applied and Environmental Microbiology 65:5594-5596.
Grant, W.E., N.R. French, and D.M. Swift. 1977. Response of a small mammal community to water and nitrogen treat-
ments in a shortgrass prairie ecosystem. Journal of Mammalogy 58:637-652.
Griffiths, B.S. 1994. Microbial-feedmg nematodes and protozoa in soil: Their effects on microbial activity and nitrogen
mineralization in decomposition hotspots and the rhizosphere. Plant and Soil 164:25-33.
Groffman, RM., C.W. Rice, and J.M. Tiedje. 1993. Denitrification in a tallgrass prairie landscape. Ecology 74:855-
862.
Groffman, RM., E.A. Holland, D.D. Myrold, G.R Robertson, and X. Zou. 1999. In Denitrification, in Standard Soil
Methods for Long Term Ecological Research, G.R Robertson, D.C. Coleman, C.S. Bledsoe, and P. Sollins (Eds),
Oxford University Press: New York, NY, USA.
Gross, K.L., K.S. Pregitzer, and A.J. Burton. 1995. Spatial variation in nitrogen availability in three successional plant
communities. Journal of Ecology 83:357-367.
Hall, AT., RE. Woods, and G.W. Barrett. 1991. Population dynamics of the meadow vole (Microtus pennsylvanicus)
in nutrient-enriched old-field communities. Journal of Mammalogy 72:332-342.
Hershey, A.E., A.R. Lima, G.J. Niemi, and R.. Regal. 1998. Effect of Bacillus thuringiensis israelensis (Bti) and me-
thoprene on nontarget macroinvertebrates in Minnesota wetlands. Ecological Applications 8:41-60.
Heske, E.J., J.H. Brown, and S. Mistry. 1994. Long-term experimental study of a Chihuahuan Desert rodent com-
munity: 13 years of competition. Ecology 75:438-445.
Hoagland, B.W., I. Butler, F.L. Johnson, and S. Glenn. 1999. The Cross Timbers. In The Savanna, Barren and Rock
Outcrop Communities of North America, R.C. Anderson, J. Fralish, and J. Baskin (Eds), Cambridge University
Press.
Hobble, S.E. and RM. Vitousek. 2000. Nutrient limitation of decomposition in Hawaiian forests. Ecology 81:1867-
1877.
Hoveland, C.S. 1993. Importance and economic significance of the Acremonium endophytes to performance of animals
and grass plant. Agriculture, Ecosystems, and Environment 44: 3-12.
Hutchinson, G.L., E.A. Davidson, 1993. Process for Production and Consumption of Gaseous Nitrogen Oxide in Soil.
Pages 79-83 In L.A. Harpet (Ed.), Agricultural Ecosystem Effects on Trace Gases and Global Climate Change.
ASA Special Publication 55ASA, CSSA and SSSA, Madison, Wisconsin, USA.
Jordan, T.E., D.E. Weller, and D.L. Correll. 1998. Denitrification in surface soils of a riparian forest: effects of water,
nitrate and sucrose additions. Soil Biology and Biochemistry 30:833-843.
32
-------
Jorgensen, E.E. 2002. Opportunities in nitrogen management research; improving applications for proven technologies
and identifying new tools for managing nitrogen flux and input in ecosystems. Advances in Restoration Science;
Number 2. U.S. EPA. EPA/600/S-02/003.
Jorgensen, E.E. and S.J. Tunnell. 2001. The effectiveness of quadrats for measuring vascular plant diversity. Texas
Journal of Science 53:365-368.
Jorgensen, E.E., S. Demarais, and W.R. Whitworth. 1994. The effect of box trap design on rodent captures. South-
western Naturalist. 39:291-294.
Jorgensen, E.E., RM. Mayer, A.E. West, M.E. Gonsoulin, S.J. Tunnell, J.E. Clark, J.L. Parsons, D.M. Engle, and E.G.
Hellgren. 2002. "Early indicators of nitrate stress; effects to ecosystems of chronic exposure to low doses of
bioavailable nitrogen." In Proceedings of the National TMDL Science and Policy Conference, Water Environment
Federation, Phoenix, AZ.
Jorgensen, E.E., S. Holub, R.G. Silva, P.M. Mayer, A.E. West, M.E. Gonsoulin, S.J. Tunnell, J.E. Clark, J.L. Parsons,
D.M. Engle, and E.G. Hellgren. 2003. "Sources and estimated load of bioavailable nitrogen attributable to chronic
nitrogen exposure and changed ecosystem structure and function." In Proceedings of the First Interagency Confer-
ence on Research in the Watersheds, Benson, AZ.
Kahl, J.S., S.A. Norton, I.J. Fernandez, K.J. Nadelhoffer, C.T. Driscoll, and J.D. Aber. 1993. Experimental inducement
of nitrogen saturation at the watershed scale. Environmental Science and Technology 27:565-568.
Kalisz, PJ. 1986. Soil properties of steep Appalachian old fields. Ecology 67:1011-1023.
Kelly, C.A., J.W.M. Rudd, and D.W. Schindler. 1990. Acidification by nitric acid -future considerations. Water, Air, and
Soil Pollution 50:49-61.
Kenward, M.G., and J.H. Roger. 1997. Small sample inference for fixed effects from restricted maximum likelihood.
Biometrics 53:983-998.
Knapp, A.K., J.M. Briggs, J.M. Blair, and C.L Turner. 1998. Patterns and controls of aboveground net primary produc-
tion in tallgrass prairie. Pages 193-221 In A.K. Knapp, J.M. Briggs, D.C. Hartnett, and S.L. Collins, (eds.) Grassland
dynamics: long-term ecological research in tallgrass prairie. Oxford University Press, Oxford, United Kingdom.
Knops, J.M.H. and D. Tilman. 2000. Dynamics of soil nitrogen and carbon accumulation for 61 years after agricultural
abandonment. Ecology 81:88-98.
Knowles, R. 1982. Denitrification. Microbiological Review 46:43-70.
Krebs, C.J. 1966. Demographic changes in fluctuating populations of Microtus califomicus. Ecological Monographs
36:239-273.
Langley, A. K., Jr., and D. J. Shure. 1988. The impact of climatic extremes on cotton rat (Sigmodon hispidus) popula-
tions. American Midland Naturalist 120:136-143.
Likens, G.E. 1992. The ecosystem approach: its use and abuse. Excellence in Ecology 3. Ecology Institute, Olden-
dorf/Luhe, Germany. 166pp.
Luo, J., R.W. Tillman, R.E. White, and PR. Ball. 1998. Variation in denitrification activity with soil depth under pasture.
Soil Biology and Biochemistry 30:897-903.
Luo, J., R.W. Tillman, and PR. Ball. 2000. Nitrogen loss through denitrification in a soil under pasture in New Zealand.
Soil Biology and Biochemistry 32:497-509.
Mahli, S.S., and M. Nyborg. 1986. Increase in mineral N in soils during winter and loss of mineral N during early spring
in north-centra! Alberta. Canadian Journal of Soil Science 66:397-409.
Matthews, J.W. and K. Clay. 2001. Influence of fungal endophyte infection on plant-soil feedback and community
interactions. Ecology 82:500-509.
McNaughton, S.J. 1985. Ecology of a grazing ecosystem: the Serengeti. Ecological Monographs 55:259-294.
Mikola, J, and H. Setala. 1998. Relating species diversity to ecosystem functioning: mechanistic backgrounds and
experimental approach with a decomposer food web. Oikos 83:180-194.
Naeem, S., D.R. Hahn, and G. Shuurman. 2000. Producer-decomposer co-dependency influences biodiversity effects.
Nature 403:762-764.
National Atmospheric Deposition Program (NADP-3)/National Trends Network. 2002. NADP Program Office, Illinois
State Water Survey, 2204 Griffith Drive, Champaign, IL 61820.
Paoletti, M. G. and M. Hassall. 1999. Woodlice (Isopoda: Oniscidea): their potential for assessing sustainability and
use as bioindicators. Agriculture, Ecosystems, and Environment 74:157-165.
33
-------
Pastor, J., M.A. Stillwell, and D. Tilman. 1987. Nitrogen mineralization and nitrification in four Minnesota old fields.
Oecologia 71:481-485.
Perakis, S.S. and L.O. Hedin. 2002. Nitrogen loss for unpolluted South American forests mainly via dissolved organic
compounds. Nature:416-419.
Peterjohn, W.T., M.B. Adams, and F.S. Gilliam. 1996. Symptoms of nitrogen saturation in two central Appalachian
hardwood forest ecosystems. Biogeochemistry 35:507-522.
Reddy, M.V. and B. Venkataiah. 1989. Influence of microarthropod abundance and climatic factors on weight loss and
mineral nutrient contents of Eucalyptus leaf litter during decomposition. Biology and Fertility of Soils 8:319-324.
Richardson, D.M., P Pysek, M. Rejmanek, M.G. Barbour, F.D. Panetta, and C.J. West. 2000. Naturalization and inva-
sion of alien plants: concepts and definitions. Diversity and Distributions 6:93-107.
Robertson, G.P, M.A. Huston, F.C. Evans, and J.M. Tiedje. 1988. Spatial variability in a successional plant community:
patterns of nitrogen availability. Ecology 69:1517-1524.
SAS Institute Inc. 1990. SAS/STAT User's Guide. Release 6. SAS Institute Inc., Gary, NC.
Scharf, P, and M.M. Alley. 1988. Centrifugation: a solution to the problems posed by ammonium and nitrate contamina-
tion of filters in soil extract analysis. Soil Science Society of America Journal 52:1508-1510.
Schetter, T.A., R.L. Lochmiller, D.M. Leslie, Jr., D.M. Engle, and M.E. Payton. 1998. Examination of the nitrogen limitation
hypothesis in non-cyclic populations of cotton rats (Sigmodon hispidus). Journal of Animal Ecology 67:705-721.
Schlesinger, W.H. 1997. Biogeochemistry: An analysis of global change. Academic Press. San Diego, CA.
Schnabel, R.R., L.F. Cornish, W.L. Stout, and J.A. Shaffer. 1996. Denitrification in a grassed and a wooded, valley
and ridge, riparian ecotone. Journal of Environmental Quality. 25:1230-1235.
Schnell, J.H. 1968. The limiting effects of natural predation on experimental cotton rat populations. Journal of Wildlife
Management 32:698-711.
Seastedt, T.R. 1984. The role of arthropods in decomposition and mineralization processes. Annual Review of Ento-
mology 29:25-46.
Shaver, G.R., J. Canadell, F.S. Chapin III, J. Gurevitch, J. Harte, G. Henry, P Ineson, S. Jonasson, J. Melillo, L. Pitelka,
and L. Rustad. 2000. Global warming and terrestrial ecosystems: a conceptual framework for analysis. Biosci-
ence 50:871-882.
Silva, R.G., K.C. Cameron, H.J. Di, and E. E. Jorgensen. 2005(a). A lysimeter study to investigate the effect of dairy
effluent and urea on cattle urine nitrogen losses, plant uptake and soil retention. Water, Air, and Soil Pollution (In
Press).
Silva, R.G., E.E. Jorgensen, S.M. Holub, and M.E. Gonsoulin. 2005(b). Relationship between culturable soil microbial
populations and gross nitrogen transformation processes in a clay loam soil across ecosystems. Nutrient Cycling
in Agroecosystems (In Press).
Smil, V. 1990. Nitrogen and phosphorus. In The Earth as Transformed by Human Action; Global and Regional Changes
in the Bioshpere over the Past 300 Years, B.L. Turner II, W.C. Clark, R.W. Kates, J.F. Richards, J.T. Mathews, and
W.B. Meyer (Eds). Cambridge University Press, United Kingdom. 713pp.
Sparrow, S.D., and D.T. Musiak. 1987. Errors in analysis for ammonium and nitrate caused by contamination from filter
papers. Soil Science Society of America Journal 51:107-110.
Stafford, S.R., and I.J. Stout. 1983. Dispersal of the cotton rat, Sigmodon hispidus. Journal of Mammalogy 64:210-
217.
Stanford, G., R.A. Vander Pol, and S. Dzienia. 1975. Denitrification rates in relation to total and extractable soil carbon.
Soil Science Society of America Proceedings 39:284-289.
Starr, R.C. and R.W. Gillham. 1993. Denitrification and organic carbon availability in two aquifers. Ground Water
31:934-947.
Stohlgren, T.J., K.A. Bull, and Y. Otsuki. 1998. Comparison of rangeland vegetation sampling techniques in the Central
Grasslands. Journal of Range Management 51:164-172.
Swift, M.J., O.W. Heal, and J.M. Anderson. 1979. Decomposition in terrestrial ecosystems. Studies in Ecology; Volume
5, University of California Press, Berkeley and Los Angeles. 372 pp.
Tait, M.J., and C.J. Krebs. 1983. Predation, cover, and food manipulations during a spring decline oiMicrotus townsen-
dii. The Journal of Animal Ecology 52:837-848.
ter Braak, C.J.F. and R Smilauer. 1998. CANOCO Reference Manual and User's Guide to Canoco for Windows: Soft-
ware for Canonical Community Ordination (Version 4). Microcomputer Power, Ithaca, NY.
34
-------
Tilman, D. 1987. Secondary succession and the pattern of plant dominance along experimental nitrogen gradients.
Ecological Monographs 57:189-214.
Tilman, D. 1988. Plant Strategies and the Dynamics and Structure of Plant Communities. Princeton University Press,
Princeton, NJ.
Tremmel, D.C. and K.M. Peterson. 1983. Competitive subordination of a Piedmont old field successional dominant by
an introduced species. American Journal of Botany 70:1125-1132.
Tunnell, S.J. 2002. Vegetation dynamics following cessation of severe disturbance in an old field. Ph.D. dissertation.
Oklahoma State University, Stillwater, OK.
Tunnell, S.J., D.E. Engle, and E.E. Jorgensen. 2004. Old field successional dynamics following cessation of chronic
disturbance. Journal of Vegetation Science 15:431-436 15:431-436.
USDA. 1973. Soil survey of Pontotoc County, Oklahoma. USDA Soil Conservation Service.
U.S. EPA. 1995. An SAB report: ecosystem management; imperative for a dynamic world. EPA-SAB-EPEC-95-003.
Van der Heijden, M.G.A., J.N. Klironomos, M. Ursic, R Moutoglis, R. Streitwolf-Engel, T. Boiler, A. Wiemken, and R.
Sanders. 1998. Mycorrhizal fungal diversity determines plant diversity, ecosystem variability and productivity.
Nature 396:69-72.
Vaughan, M.R. and L.B. Keith. 1981. Demographic response of experimental snowshoe hare populations to overwinter
food shortage. Journal of Wildlife Management 45:354-380.
Vitousek, PM. 1994. Beyond global warming: ecology and global change. Ecology 75:1861-1876.
Vitousek, P.M., J. Aber, R.W. Howarth, G.E. Likens, PA. Matson, D.W. Schindler, W.H. Schlesinger, and G.D. Tilman.
1997. Human alteration of the global nitrogen cycle: causes and consequences. Issues in Ecology 1:1-16.
Weaver, J.E. 1968. North American Prairie. University of Nebraska Press, Lincoln, NE.
Wedin, D.A. and D. Tilman. 1990. Species effects on nitrogen cycling: a test with perennial grasses. Oecologia
84:433-441.
Wedin, D.A. and D. Tilman. 1996. Influence of nitrogen loading and species composition on the carbon balance of
grasslands. Science 274:1720-1723.
Weier, K.L., J.W. Doran, J.F. Power, and D.T. Walters. 1993. Denitrification and dinitrogen/nitrous oxide ratio as affected
by soil water, available carbon and nitrate. Soil Science Society of America Journal 57:66-72.
Westerman, R.L. and T.C. Tucker. 1978. Factors affecting denitrification in a Sonoran Desert soil. Soil Science Society
of America Journal 42:596-599.
Whitehead, D.C. 1995. Grassland Nitrogen. CAB International, Wellingford, United Kingdom, pp. 405.
Wieder, R.K., J.E. Carrel, J.K. Rapp, and C.L. Kucera. 1983. Decomposition of tall fescue (Festuca elatior var. arun-
dinacea) and cellulose litter on surface mines and a tallgrass prairie in central Missouri, USA. Journal of Applied
Ecology 20:303-321.
Wrage, N., G.L. Velhof, M.L. van Beusichem, and 0. Oenema. 2001. Role of nitrifier denitrification in the production
of nitrous oxide. Soil Biology and Biochemistry 33: 1723-1732.
Zedler, J.B. and PH. Zedler. 1969. Association of species and their relationship to microtopography within old fields.
Ecology 50:432-442.
35
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