Download Nitrogen enrichment and plant communities

Ann. N.Y. Acad. Sci. ISSN 0077-8923
ANNALS OF THE NEW YORK ACADEMY OF SCIENCES
Issue: The Year in Ecology and Conservation Biology
Nitrogen enrichment and plant communities
Elsa E. Cleland1 and W. Stanley Harpole2
1
Ecology, Behavior and Evolution Section, University of California San Diego, La Jolla California, USA. 2 Ecology, Evolution and
Organismal Biology, Iowa State University, Ames, Iowa, USA
Address for correspondence: Elsa E. Cleland, 9500 Gilman Dr. #0116, University of California San Diego, La Jolla, CA 92093,
USA. [email protected]
Anthropogenic nitrogen (N) enrichment of many ecosystems throughout the globe has important ramifications
for plant communities. Observational and experimental studies frequently find species richness declines with N
enrichment, in concert with increasing primary production. Nitrogen enrichment also reorders species composition,
including species turnover through gains and losses of species, changes in dominance and rarity, and shifts in
the relative abundance of particular functional groups. Nitrogen has traditionally been considered the primary
limiting nutrient for plant growth in terrestrial ecosystems, but recent synthetic work suggests that colimitation by
phosphorus (P), water, and other resources is widespread, consistent with theoretical predictions. At the same time,
disproportionate increases in ecosystem N input are expected to exacerbate limitation by P and other resources.
Similarly, synthetic research has pointed out the important role of consumers and pathogens in determining plant
community structure, especially with respect to shifting resource availability. We argue here that environmental and
biotic contexts, including limitation by multiple resources, herbivores and pathogens, play important roles in our
understanding of plant community responses to N enrichment.
Keywords: nitrogen; phosphorus; community; diversity; richness; composition; multiple resource limitation;
stoichiometry
Introduction
Nitrogen (N) is an essential element for all forms
of life, as a building block of amino acids and ultimately proteins. Although all life is bathed in N2
gas, only bacteria capable of N fixation—its conversion to available inorganic forms—can access this
vast pool of N, resulting in fundamental N limitation of plants and their consumers in both aquatic
and terrestrial ecosystems.1 In the early 1900s, the
development of the Haber–Bosch process enabled
industrial production of ammonia (NH3 ) from N2
gas, without which the global increase in crop yields
and associated rise in human population growth
would not have been possible.2 However, most (ca.
86%) of the reactive N created for food production is lost to the environment through a number of
pathways, and does not result in human consumption.3 In addition to fertilizer production, reactive
N is emitted by fossil fuel combustion. By the 1990s,
anthropogenic N fixation exceeded the total amount
of biological N fixation on land or in the ocean.4–6
Nitrogen enrichment of natural ecosystems as a result of human activities has had a variety of negative
consequences including reductions in species richness and community structure, which will be the
focus of this review.
The topic of plant community responses to nutrient enrichment has been explored by excellent
earlier reviews by DiTommaso and Aarssen7 and
Bobbink et al.8 . This review builds on these earlier efforts by synthesizing recent literature, comparing how species richness, evenness, and functional
composition of terrestrial plant communities vary
along natural gradients versus with anthropogenic
N enrichment. In particular, this review seeks to
(1) understand the mechanisms that underlie plant
community responses to N enrichment, (2) identify aspects of environmental context that modify
these responses, and (3) incorporate the concept
doi: 10.1111/j.1749-6632.2010.05458.x
46
c 2010 New York Academy of Sciences.
Ann. N.Y. Acad. Sci. 1195 (2010) 46–61 !
Cleland & Harpole
of multiple-resource limitation that may constrain
community and ecosystem responses to N enrichment.
Nutrients and plant community structure
in natural systems
Nutrient inputs in natural systems
In preindustrial times, inputs of N in terrestrial
ecosystems arose predominantly through bacterial
N fixation, estimated to be between 100 teragrams
(Tg)/year2 and 195 Tg/year,9 with an additional estimated 5 Tg N/year fixed by lightning.10 “Natural”
fires (as opposed to human-caused fires) result in
additional volatilization and deposition of N, a minimum of 3.2 Tg N/year is likely to be deposited as
a result of “natural” fires (this number only considers fire at high latitudes).5 As N is recycled through
detritus by decomposition, N is also volatilized to
the atmosphere in the form ammonia (NH3 ), and
related forms (NOy and NHx ). This natural rate
of N deposition (as opposed to anthropogenic N
deposition, which will be discussed later) has been
estimated at 6.6 Tg/year in the form of NOy and
10.8 Tg/year in the form of NHx .5 Although the
inorganic forms NH4 + and NO3 − are the predominant sources of plant N, a number of plants utilize
amino acids and other organic forms of N, especially
in highly N-limited systems (reviewed by Chapin11
and Schimel and Bennett12 ). Species also appear
to differ in their preference for different forms of
N and species abundances correlate to the relative
availability of the form of N they prefer.13
There is great variation among ecosystems in the
rate of biological N fixation, the major source of
natural inputs of N. The highest rates of N fixation take place in tropical ecosystems, resulting in
relatively N-rich systems14 ; as a result, other micronutrients frequently limit production in tropical systems. Legumes that have the highest rates of
N fixation are frequently found to be phosphorus
(P) limited, highlighting an important link between
N and P cycling in terrestrial ecosystems. This observation has led to the speculation that P may be
the ultimate nutrient limiting production in many
ecosystems via its control over N fixation rates,15
even though short-term experiments often find that
growth is stimulated when N is added.
In contrast to N, P inputs to natural systems are
largely physical and chemical. Root exudates and
Nitrogen enrichment and plant communities
mycorrhizae can increase the rate at which P is made
available to plants, but this biological influence is
small compared to primary importance of physical weathering. Weathering of parent material is the
primary source of P for terrestrial ecosystems, estimated globally at 3 megagrams,16 with local variation from 0.05 to 5 kg P/ha year.17 Additional P
is deposited in the form of dust, ash, or pollen but
global estimates of these Aeolian inputs are not yet
available; estimates from funnel traps range from
0.07 to 1.7 kg P/ha year.17
Ecosystem development
The relative importance of N and other limiting
factors changes with the development of ecosystems
over geological time. Soils in the early states of primary succession often contain abundant mineral
nutrients with the exception of inorganic N; deposition and biological N fixation and the accumulation of organic matter increases total N over time
during both primary and secondary succession.18,19
Over geological time periods, accumulation of N is
in contrast to net losses of P and other nutrients
due to weathering20 . On Hawaiian islands with volcanic soils that are millions of years old the soils are
P-limited, whereas soils on recent lava flows that are
just a few hundred years old tend to be N limited,
and soils of intermediate ages are often N&P colimited.21 Viewing successional gradients merely as
transitioning from N to P limitation over time may
be overly simplistic. Colimitation can occur across a
gradient of soil age and also can occur on very young
soils.22 Despite the general patterns of strong N limitation on early successional soils, in addition to P,
Mg, and other micronutrients may also be limiting.
At the other end of the soil development gradient,
on geologically old soils, weathering leads to a loss
of other base cations in addition P.20
Stoichimetry competition, and R∗ Theory
The stoichiometry of biological organisms dictates
that multiple elements are required in particular ratios simultaneously for optimal growth. Because of
physiological and life history differences, species will
necessarily differ in their stoichiometry.23 Organism
physiology can strongly drive organism stoichiometry: vertebrate endoskeletons result in relatively
low organismal N:P compared to arthropods, with
their protein-based chitinous exoskeletons. Species
with greater growth rates have higher P demand
c 2010 New York Academy of Sciences.
Ann. N.Y. Acad. Sci. 1195 (2010) 46–61 !
47
Nitrogen enrichment and plant communities
Cleland & Harpole
because of greater allocation to P-rich ribosomal
nucleic acids.24 In chronically nutrient limited environments, evolution should lead to adaptations that
either increase the access of an organism to limiting nutrients (e.g., N fixation, phosphatase production) or decrease the demand for particular limiting nutrients: marine phytoplankton recently have
been found that use nonphosphorus membrane
lipids in P-limiting environments.25 Thus, because
different species represent alternative evolutionary
“solutions” to dealing with multiple and complex
constraints, they differ in both their requirements
for and ability to compete for different limiting
nutrients.
Because species have evolved different adaptations to multiple limiting factors, including nutrients, species show trade-offs, which can allow many
species to coexist when they compete for multiple
limiting nutrients.26 Nutrient availability in ecosystems is a balance between input and loss rates as
well as the ability of plants to grow and deplete
nutrients. The ability of plants to deplete nutrients
leads to plant competition. The converse prediction
is that increasing the supply of a limiting nutrient
eventually makes it nonlimiting and thus no longer
competed for: fertilization will cause competition
to shift to a different limiting factor. Because species
have trade-offs for different nutrients, changes in
nutrient ratios are predicted to lead to changes in
species composition and diversity.26
Increasing inputs of N or other nutrients will
eventually lead to imbalanced nutrient supply and
should result in strong limitation by fewer resources
for which fewer species should be the best competitors (Fig. 1).27 Thus, fertilization of ecosystems can
reduce the number or change the identity of the
factors that were originally limiting and to which
the species in that ecosystem were evolutionarily
adapted, resulting in loss of diversity.27 For example,
colimitation by multiple below ground resources
such as N and P may allow multiple species to coexist
if they trade-off the ability to acquire and use these
different resources; fertilization with N should ultimately make N nonlimiting and cause greater limitation by P (fewer limiting resources) or increased
limitation by light (increased importance of a different limiting factor) (Fig. 1).28 Fertilization may
also homogenize environments and decreased heterogeneity should lead to loss of diversity through
diminished opportunities for spatial coexistence,26
48
Figure 1. If species A–D trade off the ability to compete
for nitrogen (N) and phosphorus (P), they can coexist
pairwise at resource supply points in the gray triangular
regions. The right-angled lines indicate resource combinations where each species has zero net growth. For
example, A is the best competitor for N because it can
maintain growth at the lowest levels of N. If the blue
circular region represents preindustrial levels of N and
P supply, N deposition (red arrow) might move the set
of supply points to the red circular region. Under the
new conditions of decreased N limitation, only species
D would persist because it is the best competitor for
limiting P.
although experimental support for this mechanism
has been mixed.29
Multiple resource limitation
Multiple resource limitation is predicted from several lines of theory. Economic and optimal foraging theory predicts that the fitness of an organism
should be maximized when it is simultaneously limited by all limiting factors.30 Nonoptimal allocation
that results in unbalanced or excessive uptake of one
nutrient relative to other nutrients represents a fitness cost because excess uptake must be stored or
excreted; fitness should be increased by adjusting
allocation efforts to allow uptake of all essential nutrients in optimally balanced ratios.30 However, the
degree to which multiple limitation is observed will
depend on the extent to which plants are able to
deplete soil nutrient pools,31 as well as the degree
to which uptake of multiple nutrients covaries.32
For example, allocation to roots allows a plant to
increase access to multiple below ground nutrients
c 2010 New York Academy of Sciences.
Ann. N.Y. Acad. Sci. 1195 (2010) 46–61 !
Cleland & Harpole
simultaneously (positive covariance), whereas allocation to symbiotic mycorrhizae may maximize P
uptake at the expense of other micronutrients (negative covariance).
Plants can also exert direct control over nutrient availability to minimize limitation by any single nutrient, thereby increasing the likelihood that
multiple nutrient limitation will be observed. For
instance, plants have evolved various adaptations to
acquiring P that include variation in root architecture, phosphatase and organic acid root exudates,
and mycorrhizal symbiosis.33 In addition to biological N fixation through symbiosis, plants can also
exert feedback control over N cycling although effects of litter chemistry on decomposition and soil
microorganism communities34,35 (but see Ref. 36).
This concept of multiple resource limitation is at
odds with traditional views of resource limitation
stemming from von Liebig’s Law of the Minimum
that there is a single resource that is the most limiting
to plant growth at any given time; but this idea was
developed for single species in an agricultural setting, rather than naturally diverse communities.37
Whether or not individual plants are equally limited by multiple nutrients, communities of multiple plant species should necessarily be colimited by
multiple nutrients based on the mechanisms already
discussed.37 Thus, to understand the impacts of N
on individual plants and on communities, it is necessary to view changing N supply in the context of
multiple resource limitation.
Species composition along natural gradients
in nutrient availability
Species’ competitive trade-offs for different nutrients, as well as trade-offs for growth, reproduction, and defense are often correlated with patterns of species abundances along N gradients, and
they form the basis for mechanistic theoretical solutions to species coexistence questions.38,39 Plant
functional traits reflect strategies for resource capture that vary both within and among species, and
can indicate important trade-offs that contribute
to species coexistence. High trait dispersion (more
evenly distributed trait values than expected by
chance) can be indicative of interspecific competition, resulting in diverse strategies for resource capture in a community.40 Supporting this idea, trait
dispersion has been shown to increase with increasing soil fertility.40 N-fixing legumes and tree species
Nitrogen enrichment and plant communities
(e.g., Alder spp.) are often found on early successional soils or on soils following a fire that are relatively rich in P and micronutrients and with high
light availability but relatively depleted in N. Highnutrient soils favor good light competitors with high
allocation to aboveground leaves and stems. Plants
with various adaptations to acquiring P (e.g., root
architecture, phosphatase root exudation) can be
found on low P soils.
Changes in nutrient supply should affect species
differently. Plants that have evolved in low-nutrient
environments may respond less to increases in nutrient availability than would plants adapted to highnutrient conditions.41 For instance, species adapted
to low N conditions show a broad spectrum of traits
corresponding to either N fixing or N conserving
strategies including low N tissue, high root to shoot
allocation, and correspondingly low relative growth
rates.42 Species adapted to low N conditions also
tend to have low extractable soil nitrate when grown
in monocultures low extractable soil nitrate; this
metric is indicative of the level to which a plant can
deplete limiting soil inorganic N, corresponds to a
plant’s competitive ability for nitrate (i.e., its R∗ ).26
Low R∗ for nitrate (better competitor for limiting
N) correlates with other traits like high C:N, high
root:shoot.42 Plants that are better N competitors
(lower R∗ ) are also predictably more abundant in a
wide variety of contexts: increasing with time along
successional gradients (potential trade-off with colonization), decreasing with increasing N fertilization (N becoming less limiting), and are consistently dominant in low N old fields at multiple sites,
and in community assembly experiments on low
N soil.43
Plant community responses to
N enrichment
Anthropogenic N inputs
Nitrogen (N) deposition into terrestrial ecosystems
has more than doubled due to human activities,
principally fossil-fuel combustion, deforestation,
and agricultural intensification.2,5,44,45 These activities result in different forms of N entering the environment. For instance fossil fuel combustion results
primarily in deposition of oxidized forms (NOy ). In
contrast, agricultural activities result in deposition
of reduced forms (NHx ). This can result in regional
differences in the form of deposition. For instance,
c 2010 New York Academy of Sciences.
Ann. N.Y. Acad. Sci. 1195 (2010) 46–61 !
49
Nitrogen enrichment and plant communities
Cleland & Harpole
Holland et al.46 synthesized wet-deposition fluxes
(those occurring in precipitation) across the United
States and found peak NH4 + deposition over the
Midwest. In contrast, NO3 − deposition peak over
the Northeastern states. In many areas, the combination of N and other atmospheric pollutants has
led to acidification of soils and freshwater, causing
a suite of environmental changes including longterm nutrient losses and base cation leaching,47–50
sometimes referred to as the nitrogen cascade.51
Evidence from observational studies
Atmospheric N deposition has also been associated with observed declines in plant species richness
(i.e., the number of species found in a given area),
especially in Europe.8,52 There is variation among
ecosystems in diversity decline; particularly susceptible communities are those on poorly buffered
soils8 and wetlands (reviewed by Morris53 and
Bedford et al.54 ). Although sensitive plant species
and lichens have declined in areas with high deposition in the Western United States,55 comprehensive observational data on plant species richness
responses to N deposition in North America are
lacking. Even less attention has been focused on
tropical regions. A recent analysis of global chemistry transport models reported that in 34 of the
world biodiversity hotspots the 1990 level of N deposition was 50% greater than the global terrestrial
average, with increases predicted for all but one location by the year 2050.56 This calls attention to
potential negative impacts of N deposition in some
of the most diverse habitats.
Evidence from experimental studies
Most of our understanding of how plant communities respond to nutrient enrichment comes from experiments. Nitrogen fertilization experiments have
shown that N addition almost always increases primary production, but reduces species richness in
terrestrial systems, and there have been a number of excellent reviews (e.g., Ref. 7) and metaanalyses,57–59 which have documented this consistent pattern. There is great variation in the degree
of productivity increases and richness declines, suggesting that environmental and biotic context play
an important role in modifying community responses to enrichment (see section “The importance
of environmental and biotic context”).
50
Beyond richness responses, community structure
has also been shown to respond to N enrichment.
For instance, Collins et al.60 showed that N fertilization caused consistent changes in dominance in
a variety of North American herbaceous plant communities, frequently resulting in increased dominance by one or a few species. In a recent analysis of 274 N fertilization experiments, Hillebrand
et al.59 found that N enrichment tends to reduce
evenness, although positive effects on evenness have
also been reported.61 This analysis59 found that inherent ecosystem productivity predicted the degree
of evenness change, with highly productive ecosystems experiencing the greatest decline in evenness
(and hence increase in dominance) with nutrient
enrichment.
These observed changes in richness and evenness
beg the questions: Which species are most likely to
be lost, and which species may actually increase in
abundance as communities are reordered? Several
tentative generalizations can be made. For instance,
rare species are more likely to be lost with N enrichment than abundant species, as evidenced both in
an observational synthesis along gradients of N enrichment54 and in meta-analyses of N fertilization
experiments.58,62
There are often clear responses at the functional
group level. Xia and Wan62 found that herbaceous
species had significantly greater biomass responses
to N fertilization than woody species, in a global
meta-analysis including 456 terrestrial plant species.
For herbaceous species, Suding et al.58 found that
perennial, N fixing, and native species were more
likely to be lost under experimental fertilization. In a
review comparing the performance of co-occurring
native and invasive species, Daehler63 found that invaders frequently had a performance advantage over
native species under conditions of nutrient enrichment. This finding has been particularly robust for
exotic annual grasses, which increase in abundance
with N addition and can drive diversity declines of
forbs and N fixers.64–66
It is important to note that the spatial and temporal scales of experiments bias our understanding
of plant community responses to N enrichment. Although experiments can control for many of the
kinds of co-occurring changes that complicate observational studies of N deposition, most of the results of such experiments are based on small plots
(often 1 m2 ). If individuals increase in size with N
c 2010 New York Academy of Sciences.
Ann. N.Y. Acad. Sci. 1195 (2010) 46–61 !
Cleland & Harpole
enrichment, then per m2 there will be fewer individuals and decreased numbers of species because sampling area is fixed; with N enrichment the researcher
is sampling a smaller effective area.67 If sampling
area scaled to individual size, then the null expectation would be that increasing individual size would
not be related to numbers of species if the local area
was sufficiently larger than the sample area.
Chalcraft et al.68 examined diversity responses to
N addition at multiple spatial scales, using many of
the same herbaceous plant communities analyzed
by Suding et al.58 . Although the response of alpha
diversity (plot scale) nearly always declined with
N addition (+5% to −61%), beta diversity (species
turnover among plots) responses were variable. This
variation was associated with productivity; beta diversity increased in response to N enrichment at
low-productivity sites (up to +22%), but decreased
at high-productivity sites (up to −18% reduction).
This effect on turnover in species identity among
plots tended to ameliorate fertilization induced reductions in gamma diversity (i.e., species pools in
control and fertilized plots integrated across whole
experiments). This study points out the importance
of understanding spatial heterogeneity and turnover
among plant communities when scaling up from
plot level experiments to regional level predictions
of anthropogenic impacts.
Temporal trends are also important to consider.
Most experimental fertilization studies are short in
duration and add nutrients in excess of demand,
and at levels much higher than those experienced by
natural systems due to atmospheric N deposition.
One long-term gradient study added 1–10 g/m2 N
to native prairie grassland in Minnesota (equivalent to 10–100 kg/ha) and found significant declines
in species richness even at the lowest level of addition,69 although it took the longest time to see a
significant reduction in richness at the lowest level
of addition. This study is important because it shows
that species are responding to cumulative N loads,
that detrimental effects can result from the accrual
of even small amounts of N deposition, and also
suggests that short-term experiments with high N
addition rates are reasonable proxies for chronic,
low-level N deposition.
The longest running ecological experiment, the
Park Grass experiment at Rothamsted, Harpenden,
UK, has received annual application of combinations of NH4 + , NO3 − , P, Na, and Mg for over
Nitrogen enrichment and plant communities
150 years. Fertilization has lead to a persistent loss
of species and changes in composition.70–72 The
most dramatic effect of nutrient addition on species
number (a decline in species number from greater
than 40 to three species per plot) was largely attributable to decreased soil pH from 7.3 to 3.6 due
to the acidifying effect of chronic addition of ammonium sulfate.71 This strong effect of soil acidification prompted the establishment of additional
liming subplots to mitigate the impact on pH: liming
the ammonium addition subplots effectively added
about 15 species.71 It should be noted that communities at Park Grass are strongly colimited by multiple nutrients including N, and the greatest loss
of species occurred with the addition of multiple
nutrients.27,73 In contrast to species richness, the
genetic diversity of Anthoxanthum odoratum populations increased in plots with greater numbers
of added resources.73 This result is consistent with
the hypothesis that the addition of greater numbers
of limiting resources should decrease niche dimension, and thus species diversity27 ; as a consequence
of species loss, niche breadth of the surviving species
should increase and be reflected in the potential for
greater genetic diversity of the surviving species.73
Plant community responses to multiple
altered nutrients
A recent meta-analysis by Elser et al.74 shows that
N and P strongly colimit production across terrestrial, marine, and freshwater systems, challenging
a long-standing paradigm that ecosystem production can be viewed as primarily limited by the single
nutrient in lowest supply—typically thought to be
N in most terrestrial systems. Addition of a nutrient makes it not limiting and thus takes away
a niche axis along which some species were specialized or for which they traded off the ability
to compete along some other niche axis.27 Addition of greater numbers of limiting resources
is then, essentially, a “niche-destruction” experiment, which follows from Hutchinson’s75 prediction that greater numbers of niche axes, or niche
dimension, should allow greater numbers of coexisting species. Harpole and Tilman27 tested this
prediction experimentally and found that addition of increasing numbers of resources (e.g.,
combinations of N, P, K, or water), and therefore decreased niche dimension, led to a joint
c 2010 New York Academy of Sciences.
Ann. N.Y. Acad. Sci. 1195 (2010) 46–61 !
51
Nitrogen enrichment and plant communities
Cleland & Harpole
increase in grassland community biomass and decrease in species diversity. The relationship between
species richness and productivity can differ depending on the degree of N, P, and/or K limitation.76
Resource ratio theory26 makes a similar prediction: greatest species diversity should occur
when multiple resources are supplied at optimal or
balanced ratios; extreme or unbalanced resource ratios should correspond to reduced diversity.77,78 Orthogonal to resource ratio gradients is the total supply of resources: greater total supply, if at balanced
ratios, might increase diversity by allowing greater
numbers of individuals to persist, thus reducing
likelihood of extinction for small populations.79
Thus, increased N deposition will change both ratios
(balance) and amount of nutrients (total supply);
along with these two components of fertilization.
Cardinale et al. also include the causal positive effect of diversity on productivity and provide a conceptual multivariate, resource-based hypothesis for
the productivity–diversity relationship.79 In summary, resource imbalance decreases both diversity
and productivity (resource ratio theory); increasing
total resource supply at balanced ratios tends to promote diversity and increase productivity (Species
Energy theory); changes in resources that decrease
diversity should additionally reduce productivity
(biodiversity and ecosystem function theory); thus
resource ratios, resource availability and diversity
jointly and mechanistically can explain the often
observed hump-shaped productivity–diversity correlation patterns.79
Changing inputs, changing limitations
Despite the impressive impact human activities have
had on the global N cycle, human activities have had
an even larger impact on the global P cycle, increasing P flux by 400%.16 High levels of P are applied to
agricultural lands but more strongly impact aquatic
systems than natural terrestrial systems because
P-loss pathways are directed toward waterways and
the P cycle does not have a gaseous component like
N that can allow widespread atmospheric deposition. Thus, alterations to the N cycle, rather than
the P-cycle, have had a disproportionate effect on
natural terrestrial systems because of atmospheric
N deposition.
In addition to high rates of N inputs to terrestrial
systems that should progressively decrease N limitation and increase limitation by other nutrients over
52
time, the evidence that most systems appear to be
colimited by N, P, and perhaps other nutrients, suggests that N limitation should be diminishing and
there should be increasing limitation by other nutrients. Ratios of resource inputs are changing over
time. Ratios of N to other nutrients (e.g., Na, K, Mg,
S) in wet deposition samples from the U.S. National
Atmospheric Deposition Program appear to have
been increasing in recent years in the continental
United States (Fig. 2). Increasing levels of N relative to other potentially limiting nutrients suggest
that terrestrial systems may experience increasing
limitation by nutrients other than N over time. Although N emissions in parts of the United States
and Europe have plateaued or decreased, N emissions are increasing in others areas of the world,
in particular in countries with growing industry
such China.80 Although average total inorganic N
wet deposition may not be changing significantly
in the United States, the relative amounts of different forms of N do show stronger trends: NHy
deposition is increasing relative to NOx wet deposition (Fig. 2). This trend has implications for
altered species composition because studies have
shown species’ preferences for different N forms
that correlate with their abundance.13 Given that
N deposition is increasing in some areas, and that
other areas have already experienced about a sixfold increase in N deposition from preindustrial
levels,81 we need to explore the importance of and
potential limitation of other nutrients in natural
ecosystems.
The importance of environmental
and biotic context
Although species richness generally declines with
anthropogenic N enrichment, there is considerable variation in this response across ecosystems
(section “Plant community responses to N enrichment”). In addition to stoichiometric requirements
that could constrain species and hence community
responses to N enrichment (section “Plant community responses to multiple altered nutrients”), there
has been a growing recognition that environmental
context plays an important role in predicting variation in community responses to N enrichment.
For instance, species often respond to N addition in
predictable ways: for species that occurred across
multiple N fertilization experiments, 10 of
c 2010 New York Academy of Sciences.
Ann. N.Y. Acad. Sci. 1195 (2010) 46–61 !
Cleland & Harpole
Nitrogen enrichment and plant communities
Figure 2. Nutrient ratio wet deposition trends from annual means of 290 continental USA National Atmospheric
Deposition Program monitoring sites (excluding Alaska). Polynomial regressions suggest that historically increasing
deposition rates of inorganic N relative to other nutrients are decelerating, and may reflect recent decreases in
inorganic N deposition in some regions of the United States.123 Although recent rates of inorganic N (NH4 plus NO3 )
deposition, on average, show no significant increase, the ratios of NO3 to NH4 have declined, and ratios of N to other
nutrients have generally increased.
20 consistently increased or decreased, but abiotic
factors were also important.82 Clark et al.83 used a
structural equation model to ascertain the factors
that best explained the variation in species response
(0–65%) in 23 fertilization experiments. Greater
species loss occurred in communities with lower
cation exchange capacity, colder regional temperatures, and stronger production responses to fertilization. Similarly, Stevens et al. found that high
mean annual precipitation and high soil pH moderated the negative impact of N deposition on species
richness across a gradient of N deposition in Great
Britain.52 In this section, we summarize the importance of environmental context (in particular water
and light availability) and biotic interactions (herbivores, pathogens, soil microbes) that can mod-
ify community responses to N enrichment, and can
also potentially reveal the mechanisms that underlie
these changes (Fig. 3).
Water
Community responses to N enrichment may depend on precipitation, both because of the fundamental constraint of water limitation in many
ecosystems84 and because water availability influences the diffusion of soil N to plant roots and N
mineralization. Frequently, N addition can lead to
decreased soil moisture,65,85 either because of increased water demand through increasing biomass,
or through increased stomatal conductance as a consequence of increased leaf N and photosynthetic
rates.86 Furthermore, N deposition rates tend to
c 2010 New York Academy of Sciences.
Ann. N.Y. Acad. Sci. 1195 (2010) 46–61 !
53
Nitrogen enrichment and plant communities
Cleland & Harpole
Figure 3. Photographs of experimental plots under (A) control conditions or (B) N addition at a rate of 7 g N/m2 year
in the form of calcium nitrate, in central coastal California (Cleland and Suding, unpublished). Notice that N addition
increased production and dominance of exotic annual grasses and reduced diversity of native forbs. A summary of
hypothesized mechanisms leading to declines in species richness and evenness with N enrichment follows.
be correlated with rainfall levels,87,88 as do N loss
rates through leaching and runoff.89 Using a global
data set, Xia and Wan62 found that the percent
increase in production with fertilization increased
linearly with mean annual precipitation. In contrast, in a synthesis of fertilization experiments in
arid or semiarid ecosystems, the absolute response
of production to N enrichment increased with increasing annual precipitation, but the relative response did not change.90 The conclusion from both
of these meta-analyses was that N and water colimited production across the range of ecosystems
included.
54
Factorial N and water addition experiments have
found different effects of their interaction on species
richness. Factorial experiments are illustrative for
identifying the mechanisms that might drive responses at multiple levels for organization. In an annual grassland adding N and water both increased
productivity, but only N addition lowered species
richness.91 In this case timing was important: N addition increased biomass and decreased light levels
earlier in the growing season, resulting in a decline
in short-statured species. Harpole et al.85 found a
slightly different response to factorial additions of N
and water in California annual grassland, where N
c 2010 New York Academy of Sciences.
Ann. N.Y. Acad. Sci. 1195 (2010) 46–61 !
Cleland & Harpole
and water only increased productivity when added
simultaneously, but N addition caused a significant
decline in species richness both alone and in concert
with water addition, suggesting it was not the increase in biomass that drove the richness response.
Nitrogen enrichment and plant communities
Light
Light limitation associated with increased aboveground production is one of the primary mechanisms hypothesized to explain diversity loss in
response to N enrichment.38,92 In addition, high
production can result in high litter loads, limiting
light availability for seedlings.93 An elegant experiment by Hautier et al.28 strongly suggested that competition for light was the driver of richness decline
following nutrient enrichment. In their experimental grassland communities, the decline in species
richness following fertilization (N, P, and micronutrients) was ameliorated by lights hung inside the
plant canopy, which eliminated light limitation in
the understory.
Other recent work, however, suggests that light
limitation may not be responsible for richness declines (e.g., Ref. 94). Lamb61 manipulated soil N,
water, and light availability in a rough fescue prairie
in Canada, and analyzed their interacting effects on
species richness and evenness using structural equation models. This showed that the negative effect
of N enrichment on richness was mediated largely
through increased litter accumulation but not via
decreased light availability. Instead other effects of
increased litter (via herbivore or pathogen loads, see
later sections) may have been responsible.
that the combination of herbivory and fertilization
had the strongest positive effects on richness with
low-productivity responses to fertilization.
Herbivore performance and population sizes generally show positive responses to anthropogenic N
deposition, a survey of the literature found that plant
N concentration generally increased and carbonbased defense compounds decreased in response
to deposition.96 Nitrogen content of plant tissues
is one of the best predictors of herbivore performance,97 and it has been suggested that the
greater frequency of bark beetle outbreaks may
be related to increasing N deposition rates.96 This
recalls the theoretical concept of the “paradox of
enrichment,” whereby increasing resource availability for plants may induce a population increase in
their consumers, ultimately driving the consumed
species near extinction.98 Hence, although herbivory rates may mediate species richness declines in
response to fertilization in small-scale experiments,
anthropogenic deposition could have negative consequences for diversity via increased populations of
herbivores.
The degree of stoichiometric mismatch between
plants and herbivores is also likely to mediate the
effect of herbivory on plant growth and community composition.99 As an example of stoichiometric requirements changing with trophic level, livestock production in California rangelands can be
limited by forage protein. One management strategy has been to fertilize pastures with P and S, thus
particularly increasing the production of N-fixing
legumes.100
Herbivores
Using a meta-analysis of herbivore exclosure and
fertilization experiments, Gruner et al.95 showed
that herbivory greatly limits the apparent productivity increase with fertilization in both terrestrial
and aquatic ecosystems. In a complementary metaanalysis using many of the same studies, Hillebrand
et al.59 showed that species richness and evenness responded in complex ways to manipulation of both
nutrients and herbivores. Herbivores generally increased richness and evenness, both alone and in
combination with fertilization. The researchers hypothesized that herbivory on dominant species may
have contributed both to increased evenness and
removal of biomass, hence preventing light limitation. This hypothesis was supported by the fact
Pathogens
Pathogens can dramatically reduce plant productivity,101 and can alter plant community composition
and diversity.102 In general, N enrichment increases
plant susceptibility to pathogenic fungi,103–105 although declines in susceptibility have also been reported (e.g., Ref. 106) and the effect likely varies
among pathogen types.107 For example, in Sweden
the parasitic fungus Valdensia heterodoxa played a
key role in vegetation responses to nutrient enrichment,103 whereby the shrub Vaccinium myrtillus had increased fungal infection with N enrichment, resulting in increased leaf shedding, increased
light availability in the understory, and ultimately
increased abundance of the grass Deschampsia
flexulosa.108
c 2010 New York Academy of Sciences.
Ann. N.Y. Acad. Sci. 1195 (2010) 46–61 !
55
Nitrogen enrichment and plant communities
Cleland & Harpole
Pathogens may also play a role in altered plant–
soil feedbacks with N enrichment.109 For instance,
N addition tends to alter microbial communities
by increasing the relative abundance of fungi over
bacteria,110 although dominant plant composition
had a greater effect on overall microbial biomass and
composition in this study. Mycorrhizal composition
also changes in response to N addition,111,112 although it is not known if the facilitative/pathogenic
balance in this essential plant–microbe interaction
is altered.
Synthesis and conclusions
Synthesis and future research needs
Recent empirical and conceptual advances in our
understanding of resource limitation and the impacts of changing nutrient inputs to ecosystems have
contributed to a better understanding of the ecology of terrestrial ecosystems, but they also highlight
several areas that are not well understood. N and
P colimitation appears to be common; we know
relatively little about the importance of other nutrients in natural systems, in particular whether they
are interactively colimiting with N and P. Greater
efforts are being made to monitor nutrient deposition around the world. Such data will provide important regional and continental scale estimates of
deposition trends. Smaller scale data are also needed
because regional scale extrapolations can greatly
underestimate local deposition rates and it is the
local scale variation that is important for understanding plant community response to atmospheric
deposition.
Because ecological systems are enormously complex, many studies focus on single factors, yet the
pattern we see is that interactions are not only common but often expected theoretically: interactions
between plant species, multiple resources, trophic
levels, climate change factors, etc. Our conceptual
ability to interpret higher order interactions is limited, as is our data to test them.
We study species that for the most part evolved in
a preindustrial world and all areas of the world are
impacted by some aspect of global change. Species
traits reflect adaptations to environmental conditions that either no longer exist or are changing
rapidly. There is likely to be a growing mismatch
between evolved species traits and novel resource
and limiting factor environments. We need to bet-
56
ter synthesize evolution and ecology to understand
the implications for biodiversity and conservation
of a world and its species, which may all be changing
at different rates.
Many of our recent advances in understanding
have come from meta-analysis across many studies that are similar. A greater challenge will be to
synthesize across different kinds of studies (e.g., observational and experimental) and identify gaps in
our understanding that can only be filled by imagining new types of studies that are different than our
conventional approaches.
Management implications
In Europe, there have been intense research efforts
to set and refine “critical loads” for N deposition,
defined as “a quantitative estimate of an exposure
to one or more pollutants below which significant
harmful effects on specified sensitive elements of
the environment do not occur according to present
knowledge.”113 This research was driven by demand
for estimates of critical loads by international European air pollution treaties; critical loads have been
an established component of European pollution
policy for over 20 years and have been successful
in reversing ecosystem acidification trends. Lacking
an analogous legislative demand, critical loads have
not been developed for the United States, although
there are federal mandates to protect public lands
from the effects of N deposition (reviewed by Porter
et al.114 ). The recent formation of a Critical Loads
sub-committee within the National Atmospheric
Deposition Program within the Environmental Protection Agency suggests that there is growing interest
in this concept in the United States.115 . A variety of
approaches have been proposed for the definition of
critical loads in the United States, including empirical approaches based on long-term ecological monitoring,116 identifying sensitive indicator species,117
“ecological hindcasting,”81 and simulation models
of N saturation and subsequent nitrate leaching and
acidification.47,118
It is unclear how the rate of community and
ecosystem recovery will proceed once N inputs are
reduced below critical loads,119 and whether systems will return to their original state or experience hysteresis, remaining in an altered state.120,121
Careful experiments are required to identify threshold responses of sensitive communities and ecosystems to N deposition,69,122 both in terms of initial
c 2010 New York Academy of Sciences.
Ann. N.Y. Acad. Sci. 1195 (2010) 46–61 !
Cleland & Harpole
responses to enrichment, and levels below which
recovery from enrichment is possible.121
Conclusion
Nitrogen is an important limiting nutrient across
ecosystems, but consistent and globally general patterns of colimitation highlight the importance of
other resources such as phosphorus availability and
precipitation inputs. We know far less about the importance of other limiting nutrients in structuring
terrestrial plant communities, in particular when
there are important connections between nutrient
cycles, such as between N and P via N fixation.
Super-abundant N (and lower pH) should lead to
new scarcities and potential for release of other elements in soil that could increase toxicity; species
best adapted to those novel conditions are likely to
be different than those found in current communities adapted to low N environments. Numerous
experimental and observational studies suggest that
as N enrichment proceeds, communities become increasingly dominated by a fewer species, particularly
when productivity is increased. Although competition for other resources such as water or light is
frequently invoked as the mechanism to explain diversity declines in concert with increasing productivity, we are beginning to understand the importance of other biotic interactions such as herbivory
or disease that could mediate these responses. As
policymakers and land managers seek to limit N
deposition through development of critical loads
or other legislative mechanisms, ecologists have
the opportunity to inform the outcome by communicating their knowledge about the importance
of N in structuring natural plant communities, as
well as the importance of environmental and biotic
context that mediates community responses to N
deposition.
Acknowledgments
The authors are grateful to an anonymous reviewer
for detailed and constructive comments on an earlier draft of this review. The authors are also grateful
to Steven Porder and Jim Dalling for constructive
conversations that greatly improved particular sections of this review.
Conflicts of interest
The authors declare no conflicts of interest.
Nitrogen enrichment and plant communities
References
1. Vitousek, P.M. & R.W. Howarth. 1991. Nitrogen limitation on land and in the sea – how can it occur?
Biogeochemistry 13: 87–115.
2. Smil, V. The Earth as Transformed by Human Action.
1990. B.L. Turner et al., Eds.: 423–436. Cambridge University Press. Cambridge.
3. Galloway, J. & E. Cowling. 2002. Reactive nitrogen and
the world: 200 years of change. Ambio 31: 64–71.
4. Gruber, N. & J. N. Galloway. 2008. An Earth-system
perspective of the global nitrogen cycle. Nature 45: 293–
296.
5. Galloway, J.N. et al. 2004. Nitrogen cycles: past, present,
and future. Biogeochemistry 70: 153–226.
6. Vitousek, P. et al. 1997. Human alteration of the global
nitrogen cycle: sources and consequences. Ecol. Appl. 7:
737–750.
7. DiTommaso, A. & L. W. Aarssen. 1989. Resource manipulations in natural vegetations: a review. Vegetation
84: 9–29.
8. Bobbink, R., M. Hornung & J. G. M. Roelofs. 1998. The
effects of air-borne nitrogen pollutants on species diversity in natural and semi-natural European vegetation.
J. Ecol. 86: 717–738.
9. Cleveland, C.C. et al. 1999. Global patterns of terrestrial
biological nitrogen (N-2) fixation in natural ecosystems. Global Biogeochem. Cycles 13: 623–645.
10. Lelieveld, J. & F. J. Dentener. 2000. What controls tropospheric ozone? J. Geophys. Res.-Atmos. 105(D3): 3531–
3551.
11. Chapin, F.S. 1980. The mineral-nutrition of wild plants.
Annu. Rev. Ecol. Syst. 11: 233–260.
12. Schimel, J. P. & J. Bennett. 2004. Nitrogen mineralization: challenges of a changing paradigm. Ecology 85:
591–602.
13. McKane, R.B. et al. 2002. Resource-based niches provide a basis for plant species diversity and dominance
in arctic tundra. Nature 415: 68–71.
14. Matson, P.A., W.H. McDowell, A.R. Townsend & P.M.
Vitousek. 1999. The globalization of N deposition:
ecosystem consequences in tropical environments. Biogeochemistry 46: 67–83.
15. Vitousek, P.M., S. Porder, B.Z. Houlton & O.A.
Chadwick. 2010. Terrestrial phosphorus limitation:
mechanisms, implications, and nitrogen-phosphorus
interactions. Ecol. Appl. 20: 5–15.
16. Falkowski, P. et al. 2000. The global carbon cycle: a test
of our knowledge of Earth as a system. Science 290:
291–294.
c 2010 New York Academy of Sciences.
Ann. N.Y. Acad. Sci. 1195 (2010) 46–61 !
57
Nitrogen enrichment and plant communities
Cleland & Harpole
17. Newman, E.I. 1995. Phosphorus inputs to terrestrial
ecosystems. J. Ecol. 83: 713–726.
18. Chapin F.S. III, P.A. Matson & H.A. Mooney. 2002. Principles of Terrestrial Ecosystem Ecology. Springer. New
York.
19. Knops, J.M.H. & D. Tilman. 2000. Dynamics of soil
nitrogen and carbon accumulation for 61 years after
agricultural abandonment. Ecology (Washington, DC)
81: 88–98.
20. Walker, T.W. & J.K. Syers. 1976. The fate of phosphorus
during pedogenesis. Geoderma 15: 1–19.
21. Vitousek, P.M. & H. Farrington. 1997. Nutrient limitation and soil development: experimental test of a
biogeochemical theory. Biogeochemistry 37: 63–75.
22. Raich, J.W., A.E. Russell, T.E. Crews et al. 1996. Both
nitrogen and phosphorus limit plant production on
young Hawaiian lava flows. Biogeochemistry 32: 1–14.
23. Sterner, R.W. & J.J. Elser. 2002. Ecological Stoichiometry:
The Biology of Elements from Molecules to the Biosphere.
Princeton University Press. Princeton.
24. Elser, J.J. et al. 2003. Growth rate-stoichiometry couplings in diverse biota. Ecol. Lett. 6: 936–943.
25. Van Mooy, B.A. et al. 2009. Phytoplankton in the ocean
use non-phosphorus lipids in response to phosphorus
scarcity. Nature 458: 69–72.
26. Tilman, D. 1982. Resource Competition and Community
Structure. Princeton University Press. Princeton.
27. Harpole, W.S. & D. Tilman. 2007. Grassland species loss
due to reduced niche dimension. Nature 446: 791–793.
28. Hautier, Y., P.A. Niklaus & A. Hector. 2009. Competition for light causes plant biodiversity loss after eutrophication. Science 324: 636–638.
29. Reynolds, H.L., G.G. Mittelbach, T.L. Darcy-Hall et al.
2007. No effect of varying soil resource heterogeneity
on plant species richness in a low fertility grassland.
J. Ecol. 95: 723–733.
30. Bloom, A.J., F.S. Chapin & H.A. Mooney. 1985. Resource limitation in plants: an economic analogy. Annu.
Rev. Ecol. Syst. 16: 363–392.
31. Rastetter, E.B. & G.R. Shaver. 1992. A model of
multiple-element limitation for acclimating vegetation.
Ecology 73: 1157–1174.
32. Gleeson, S.K. & D. Tilman. 1992. Plant allocation and
the multiple limitation hypothesis. Am. Nat. 139: 1322–
1343.
33. Vance, C.P., C. Uhde-Stone & D.L. Allan. 2003. Phosphorus acquisition and use: critical adaptations by
plants for securing a nonrenewable resource. New Phytol. 157: 423–447.
34. Chapman, S.K., J.A. Langley, S.C. Hart & G.W. Koch.
58
35.
36.
37.
38.
39.
40.
41.
42.
43.
44.
45.
46.
47.
48.
49.
2006. Plants actively control nitrogen cycling: uncorking the microbial bottleneck. New Phytol. 169: 27–
34.
Wedin, D.A. & D. Tilman. 1990. Species effect on nitrogen cycle: a test with perennial grasses. Oecologia 84:
433–441.
Knops, J.M.H., K.L. Bradley & D.A. Wedin. 2002. Mechanisms of plant species impacts on ecosystem nitrogen
cycling. Ecol. Lett. 5: 454–466.
Danger, M., T. Daufresne, F. Lucas et al. 2008. Does
Liebig’s law of the minimum scale up from species to
communities? Oikos 117: 1741–1751.
Wedin, D. & D. Tilman. 1993. Competition among
grasses along a nitrogen gradient: initial conditions and
mechanisms of competition. Ecol. Monogr. 63: 199–229.
Tilman, D. 1987. The importance of mechanisms of
interspecific competition. Am. Nat. 129: 769–774.
Weiher, E., G.D.P. Clarke & P.A. Keddy. 1998. Community assembly rules, morphological dispersion, and the
coexistence of plant species. Oikos 81: 309–322.
Chapin F.C. III, P.M. Vitousek & K. V. Cleve. 1986. The
nature of nutrient limitation in plant communities. Am.
Nat. 127: 48–58.
Craine, J.M. et al. 2002. Functional traits, productivity
and effects on nitrogen cycling of 33 grassland species.
Funct. Ecol. 16: 563–574.
Harpole, W.S. & D. Tilman. 2006. Non-neutral patterns
of species abundance in grassland communities. Ecol.
Lett. 9: 15–23.
Howarth, R.W., E.W. Boyer, W.J. Pabich & J.N.
Galloway. 2002. Nitrogen use in the United States from
1961–2000 and potential future trends. Ambio 31: 88–
96.
Vitousek, P.M. et al. 1997. Human alteration of the
global nitrogen cycle: sources and consequences. Ecol.
Appl. 7: 737–750.
Holland, E.A., B.H. Braswell, J. Sulzman & J.F.
Lamarque. 2005. Nitrogen deposition onto the United
States and western Europe: synthesis of observations
and models. Ecol. Appl. 15: 38–57.
Schulze, E.D., W. d. Vries & M. Hauhs. 1989. Critical loads for nitrogen deposition on forest ecosystems.
Water, Air and Soil Pollution 48: 451–456.
Aber, J. et al. 1998. Nitrogen saturation in temperate
forest ecosystems – hypotheses revisited. Bioscience 48:
921–934.
Roem, W.J., H. Klees & F. Berendse. 2002. Effects of
nutrient addition and acidification on plant species diversity and seed germination in heathland. J. Appl. Ecol.
39: 937–948.
c 2010 New York Academy of Sciences.
Ann. N.Y. Acad. Sci. 1195 (2010) 46–61 !
Cleland & Harpole
50. Matson, P.A., K.A. Lohse & S.J. Hall. 2002. The globalization of nitrogen deposition: consequences for terrestrial ecosystems. Ambio 31: 113–119.
51. Galloway, J.N. et al. 2003. The nitrogen cascade. Bioscience 53: 341–356.
52. Stevens, C.J., N.B. Dise, J.O. Mountford & D.J.
Gowing. 2004. Impact of nitrogen deposition on the
species richness of grasslands. Science 303: 1876–1879.
53. Morris, J.T. 1991. Effects of nitrogen loading on wetland
ecosystems with particular reference to atmospheric deposition. Annu. Rev. Ecol. Syst. 22: 257–279.
54. Bedford, B.L., M.R. Walbridge & A. Aldous. 1999. Patterns in nutrient availability and plant diversity of temperate North American wetlands. Ecology 80: 2151–
2169.
55. Fenn, M.E. et al. 2003. Ecological effects of nitrogen
deposition in the western United States. Bioscience 53:
404–420.
56. Phoenix, G.K. et al. 2006. Atmospheric nitrogen deposition in world biodiversity hotspots: the need for
a greater global perspective in assessing N deposition
impacts. Global Change Biol. 12: 470–476.
57. Gough, L., C. Osenberg, K. Gross & S. Collins. 2000.
Fertilization effects on species density and primary productivity in herbaceous plant communities. Oikos 89:
428–439.
58. Suding, K.N. et al. 2005. Functional- and abundancebased mechansisms explain diversity loss due to
N fertilization. Proc. Natl. Acad. Sci. 102: 4387–
4392.
59. Hillebrand, H. et al. 2007. Consumer versus resource
control of producer diversity depends on ecosystem
type and producer community structure. Proc. Natl.
Acad. Sci. 104: 10904–10909.
60. Collins, S.L. et al. 2008. Rank clocks and community
dynamics. Ecology 89: 3534–3541.
61. Lamb, E.G. 2008. Direct and indirect control of grassland community structure by litter, resources, and
biomass. Ecology 89: 216–225.
62. Xia, J. & S. Wan. 2008. Global response patterns of terrestrial plant species to nitrogen addition. New Phytol.
179: 428–439.
63. Daehler, C.C. 2003. Performance comparisons of cooccurring native and alien invasive plants: implications
for conservation and restoration. Annu. Rev. Ecol. Evol.
Syst. 34: 183–211.
64. Stevens, C.J., N.B. Dise, D.J.G. Gowing & J.O.
Mountford. 2006. Loss of forb diversity in relation to
nitrogen deposition in the UK: regional trends and potential controls. Global Change Biol. 12: 1823–1833.
Nitrogen enrichment and plant communities
65. Zavaleta, E. et al. 2003. Grassland responses to three
years of elevated temperature, CO2, precipitation, and
N deposition. Ecol. Monogr. 73: 585–604.
66. Huenneke, L.F., S.P. Hamburg, R. Koide, et al. 1990. Effects of soil resources on plant invasion and community
structure in Californian serpentine grassland. Ecology
71: 478–491.
67. Oksanen, J. 1996. Is the humped relationship between
species richness and biomass an artefact due to plot
size? J. Ecol. 84: 293–295.
68. Chalcraft, D.R. et al. 2008. Scale-dependent responses
of plant biodiversity to nitrogen enrichment. Ecology
89: 2165–2171.
69. Clark, C. M. & D. Tilman. 2008. Loss of plant species
after chronic low-level nitrogen deposition to prairie
grasslands. Nature 451: 712–715.
70. Brenchley, W. E. & K. Warington. 1958. The park grass
plots at rothamsted 1856–1949. The Park Grass plots at
Rothamsted 1856–1949.
71. Crawley, M.J. et al. 2005. Determinants of species richness in the Park Grass Experiment. Am. Nat. 165: 179–
192.
72. Tilman, D. et al. 1994. Long-Term Experiments in Agricultural and Ecological Sciences R. A. Leigh & A. E.
Johnston, Eds.: CAB International. Wallingford.
73. Silvertown, J., P.M. Biss & J. Freeland. 2009. Community genetics: resource addition has opposing effects on
genetic and species diversity in a 150-year experiment.
Ecol. Lett. 12: 165–170.
74. Elser, J.J. et al. 2007. Global analysis of nitrogen and
phosphorus limitation of primary producers in freshwater, marine, and terrestrial ecosystems. Ecol. Lett. 10:
1135–1142.
75. Hutchinson, G.E. 1957. Concluding remarks. Cold
spring harbor symposium. Quant. Biol. 22: 415–427.
76. Olde Venterink, H., M.J. Wassen, A.W.M. Verkroost
& P.C. de Ruiter. 2003. Species richness-productivity
patterns differ between N-, P-, and K-limited wetlands.
Ecology 84: 2191–2199.
77. Braakhekke, W.G. & D.A.P. Hooftman. 1999. The resource balance hypothesis of plant species diversity in
grassland. J. Veg. Sci. 10: 187–200.
78. Herbert, D.A., E.B. Rastetter, L. Gough & G.R. Shaver.
2004. Species diversity across nutrient gradients: an
analysis of resource competition in model ecosystems.
Ecosystems 7: 296–310.
79. Cardinale, B.J., H. Hillebrand, W.S. Harpole et al. 2009.
Separating the influence of resource ‘availability’ from
resource ‘imbalance’ on productivity–diversity relationships. Ecol. Lett. 12: 475–487.
c 2010 New York Academy of Sciences.
Ann. N.Y. Acad. Sci. 1195 (2010) 46–61 !
59
Nitrogen enrichment and plant communities
Cleland & Harpole
80. Richter, A., J.P. Burrows, H. Nüss et al. 2005. Increase
in tropospheric nitrogen dioxide over China observed
from space. Nature 437: 129–132.
81. Baron, J.S. 2006. Hindcasting nitrogen deposition to
determine an ecological critical load. Ecol. Appl. 16:
433–439.
82. Pennings, S.C. et al. 2005. Do individual plant species
show predictable responses to nitrogen addition across
multiple experiments? Oikos 110: 547–555.
83. Clark, C.M. et al. 2007. Environmental and plant community determinants of species loss following nitrogen
enrichment. Ecol. Lett. 10: 596–607.
84. Churkina, G. & S.W. Running. 1998. Contrasting climatic controls on the sstimated productivity of global
terrestrial biomes. Ecosystems 1: 206–215.
85. Harpole, W.S., D.L. Potts & K.N. Suding. 2007. Ecosystem responses to water and nitrogen amendment in a
California grassland. Global Change Biol. 13: 1–8.
86. Field, C. & H.A. Mooney. 1986. On the Economy of
Plant form and Function. T. Givnish 25–55. Cambridge
University Press. Cambridge.
87. Fenn, M.E. et al. 2003. Nitrogen emissions, deposition,
and monitoring in the western United States. Bioscience
53: 391–403.
88. Báez, S., J. Fargione, D.I. Moore et al. 2006. Atmospheric nitrogen deposition in the northern Chihuahuan desert: temporal trends and potential consequences. J. Arid Environ. 68: 640–651.
89. Mitchell, M.J. et al. 1996. Climatic control of nitrate
loss from forested watersheds in the northeast United
States. Environ. Sci. Technol. 30: 2609–2612.
90. Hooper, D.U. & L. Johnson. 1999. Nitrogen limitation
in dryland ecoystems: responses to geographical and
temporal variation in precipitation. Biogeochemistry 46:
247–293.
91. Goldberg, D.E. & T. Miller. 1990. Effects of different
resource additions on species diversity in an annual
plant community. Ecology 71: 213–225.
92. Tilman, D. 1988. Plant Strategies and the Dynamics and
Structure of Plant Communities, Vol. 26. Princeton University Press. Princeton.
93. Foster, B.L. & K.L. Gross. 1998. Species richness in a
successional grassland: effects of nitrogen enrichment
and plant litter. Ecology 79: 2593–2602.
94. Rajaniemi, T.K. 2002. Why does fertilization reduce
plant species diversity? Testing three competition-based
hypotheses. J. Ecol. 90: 316–324.
95. Gruner, D.S. et al. 2008. A cross-system synthesis of
consumer and nutrient resource control on producer
biomass. Ecol. Lett. 11: 740–755.
60
96. Throop, H.L. & M.T. Lerdau. 2004. Effects of nitrogen
deposition on insect herbivory: implications for community and ecosystem processes. Ecosystems 7: 109–
133.
97. Mattson, W.J. 1980. Herbivory in relation to plant nitrogen content. Annu. Rev. Ecol. Syst. 11: 119–161.
98. Rosenzweig, M.L. 1971. Paradox of enrichment: destabilization of exploitation ecosystems in ecological time.
Science 171: 385–387.
99. Hillebrand, H. et al. 2009. Herbivore metabolism and
stoichiometry each constrain herbivory at different organizational scales across ecosystems. Ecol. Lett. 12:
516–527.
100. Jones, M.B. et al. 1990. Effects of phosphorus and
sulfur fertilization on subclover-grass pasture production as measured by lamb gain. J. Prod. Agric. 3: 534–
539.
101. Mitchell, C.E. 2003. Trophic control of grassland production and biomass by pathogens. Ecol. Lett. 6: 147–
155.
102. Gilbert, G.S. 2002. Evolutionary ecology of plant diseases in natural ecosystems. Annu. Rev. Phytopathol. 40:
13–43.
103. Strengbom, J., A. Nordin, T. Nasholm & L. Ericson.
2002. Parasitic fungus mediates change in nitrogenexposed boreal forest vegetation. J. Ecol. 90: 61–67.
104. Nordin, A., T. Näsholm & L. Ericson. 1998. Effects of
simulated N deposition on understorey vegetation of a
boreal coniferous forest. Funct. Ecol. 12: 691–699.
105. Paul, N.D. 1990. Pests, Pathogens and Plant Communities. J.J. Burdon & S.R. Leather, Eds.: 81–96. Blackwell.
Oxford.
106. Strengbom, J. & P.B. Reich. 2006. Elevated [CO2] and
increased N supply reduce leaf disease and related photosynthetic impacts on Solidago rigida. Oecologia 149:
519–525.
107. Hoffland, E., M.J. Jeger & M.L. van Beusichem. 2000.
Effect of nitrogen supply rate on disease resistance in
tomato depends on the pathogen. Plant Soil 210: 239–
247.
108. Strengbom, J., T. Näsholm & L. Ericson. 2004. Light,
not nitrogen, limits growth of the grass Deschampsia
flexuosa in boreal forests. Can. J. Bot. 82: 430–435.
109. Manning, P., S.A. Morrison, M. Bonkowski & R.D.
Bardgett. 2008. Nitrogen enrichment modifies plant
community structure via changes to plant–soil feedback. Oecologia 157: 661–673.
110. Bardgett, R.D. et al. 1999. Plant species and nitrogen
effects on soil biological properties of temperate upland
grasslands. Funct. Ecol. 13: 650–660.
c 2010 New York Academy of Sciences.
Ann. N.Y. Acad. Sci. 1195 (2010) 46–61 !
Cleland & Harpole
111. Treseder, K.K. 2004. A meta-analysis of mycorrhizal responses to nitrogen, phosphorus, and atmospheric CO2 in field studies. New Phytol. 164: 347–
355.
112. Johnson, N.C., D.L. Rowland, L. Corkidi et al. 2003.
Nitrogen enrichment alters mycorrhizal allocation at
five mesic to semiarid grasslands. Ecology 84: 1895–
1908.
113. Nilsson, J. & P. Grennfelt. 1988. Critical loads for sulfur
and nitrogen. Report for the Nordic Council of Ministers,
Copenhagen, Denmark.
114. Porter, E., T. Blett, D.U. Potter & C. Huber. 2005. Protecting resources on federal lands: implications of critical loads for atmospheric deposition of nitrogen and
sulfur. Bioscience 55: 603–612.
115. Burns, D.A., T. Blett, R. Haeuber & L.H. Pardo. 2008.
Critical loads as a policy tool for protecting ecosystems
from the effects of air pollutants. Front. Ecol. Env. 6:
156–159.
116. Williams, M.W. & K.A. Tonnessen. 2000. Critical loads
for inorganic nitrogen deposition in the Colorado Front
Range, USA. Ecol. Appl. 10: 1648–1665.
117. Bowman, W.D., J.L. Gartner, K. Holland & M.
Wiedermann. 2006. Nitrogen critical loads for alpine
Nitrogen enrichment and plant communities
118.
119.
120.
121.
122.
123.
c 2010 New York Academy of Sciences.
Ann. N.Y. Acad. Sci. 1195 (2010) 46–61 !
vegetation and terrestrial ecosystem response—are we
there yet? Ecol. Appl. 16: 1183–1193.
Driscoll, C.T. et al. 2003. Nitrogen pollution in the
northeastern United States: sources, effects, and management options. Bioscience 53: 357–374.
Nordin, A., J. Strengbom, J. Witzell et al. 2005. Nitrogen deposition and the biodiversity of boreal forests:
implications for the nitrogen critical load. Ambio 34:
20–24.
Folke, C. et al. 2004. Regime shifts, resilience, and biodiversity in ecosystem management. Annu. Rev. Ecol.,
Evol. Syst. 35: 557–581.
Clark, C.M., S.E. Hobbie, R. Venterea & D. Tilman.
2009. Long-lasting effects on nitrogen cycling 12
years after treatments cease despite minimal longterm nitrogen retention. Global Change Biol. 15: 1755–
1766.
Groffman, P. et al. 2006. Ecological thresholds: the key
to successful environmental management or an important concept with no practical application? Ecosystems
9: 1–13.
Monteith, D.T. et al. 2007. Dissolved organic carbon
trends resulting from changes in atmospheric deposition chemistry. Nature 450: 537–540.
61