Ecology, 85(5), 2004, pp. 1179–1192
q 2004 by the Ecological Society of America
CARBON SEQUESTRATION IN ECOSYSTEMS:
THE ROLE OF STOICHIOMETRY
DAG O. HESSEN,1,6 GÖRAN I. ÅGREN,2 THOMAS R. ANDERSON,3 JAMES J. ELSER,4
AND
PETER C.
DE
RUITER5
1University of Oslo, Department of Biology, P.O. Box 1027 Blindern, 0316 Oslo, Norway
Department of Ecology and Environmental Research, Swedish University of Agricultural Sciences, Box 7072,
SE-750 07 Uppsala, Sweden
3Southampton Oceanography Centre, Waterfront Campus, European Way, Southampton SO14 3ZH, UK
4School of Life Sciences, Arizona State University, Tempe, Arizona 85287 USA
5Department of Environmental Studies, Copernicus Research Institute for Sustainable Development and Innovation,
Utrecht University, P.O. Box 80115, 3508 TC Netherlands
2
Key words: carbon; C sequestration; ecosystems; C-use efficiency; grazing; nitrogen; phosphorus; stoichiometry, role in carbon sequestration.
INTRODUCTION
From the classical works of Lindeman (1942) onwards, the concept of ‘‘energy flow’’ in food webs has
been and remains a cornerstone in ecological theory in
which the height of the trophic pyramid, and thus community structure of the food web, is limited by the
efficiency of energy transfer between trophic levels
(Hutchinson 1959, Hairston et al. 1960). The extent to
which trophic structure controls trophic efficiency or
vice versa is, however, not trivial. Hairston and Hairston (1993) argued that trophic structure should be the
main determinant of energy transfer and trophic efficiency. They also pointed to the fact that aquatic food
chains have high transfer efficiency of energy and matter compared with terrestrial food chains. A more detailed account of this has been provided by Cebrian
(1999), who illustrated clearly how variation in trophic
efficiency across different ecosystems was linked with
autotroph growth rate and nutrient quotas, with a strong
gradient from forest ecosystems at one extreme to
planktonic ecosystems at the other. The high transfer
Manuscript received 26 April 2002; revised 4 June 2003; accepted 21 July 2003; final version received 10 September 2003.
Corresponding Editor: R. M. Nisbet. For reprints of this Special
Feature, see footnote 1, p. 1177.
6 E-mail: dag.hessen@bio.uio.no
efficiency of C in planktonic systems may be attributed
to both high cell quotas of N and P relative to C (highquality food for reasons given below) and to decreased
importance of low-quality structural matter like lignins
and cellulose that are poorly assimilated. These factors
point to fundamental differences between ecosystems,
not only with regard to the transfer and sequestration
of carbon, but also with regard to community composition and ecosystem function in more general terms.
One key determinant of these ecosystem processes is
the relative abundance of key elements, i.e., ‘‘ecological stoichiometry’’ (Sterner and Elser 2002).
While energy is most appropriately given in terms
of joules, it is more convenient to use carbon (C) as a
basic currency for the overall flow of both energy and
matter. First of all, C is the major element in living
materials, contributing roughly 50% of dry mass. Second, C is easily measured simultaneously with other
key elements in organic tissue. Third, the C cycle has
per se become a theme of major interest in a ‘‘globalchange’’ context, and to a large extent the global effects
are the sum of uptake and sequestration of C in aquatic
and terrestrial ecosystems. Ecological stoichiometry is
a tool for analyzing how the balance of the multiple
elements required by organisms affects production, nutrient cycling, and food-web dynamics (Sterner and El-
1179
Special Feature
Abstract. The fate of carbon (C) in organisms, food webs, and ecosystems is to a major
extent regulated by mass-balance principles and the availability of other key nutrient elements. In relative terms, nutrient limitation implies excess C, yet the fate of this C may
be quite different in autotrophs and heterotrophs. For autotrophs nutrient limitation means
less fixation of inorganic C or excretion of organic C, while for heterotrophs nutrient
limitation means that more of ingested C will ‘‘go to waste’’ in the form of egestion or
respiration. There is in general a mismatch between autotrophs and decomposers that have
flexible but generally high C:element ratios, and consumers that have lower C:element
ratios and tighter stoichiometric regulation. Thus, C-use efficiency in food webs may be
governed by the element ratios in autotroph biomass and tend to increase when C:element
ratios in food approach those of consumers. This tendency has a strong bearing on the
sequestration of C in ecosystems, since more C will be diverted to detritus entering soils
or sediments when C-use efficiency is low due to stoichiometric imbalance. There will be
a strong evolutionary pressure to utilize such excess C for structural and metabolic purposes.
This article explores how these basic principles may regulate C sequestration on different
scales in aquatic and terrestrial ecosystems.
Special Feature
1180
DAG O. HESSEN ET AL.
Ecology, Vol. 85, No. 5
FIG. 1. Conceptual diagram illustrating the potential fate of assimilated C in a homeostatic organism. The threshold
elemental ratio denotes the threshold of the C:nutrient ratio above which C will be in excess relative to the grazers’ somatic
demands. Excess C may be disposed of in various ways, all causing decreased growth efficiency in terms of C.
ser 2002). The origins of stoichiometry are rooted in
agricultural studies involving Liebig’s law of the minimum. In the stoichiometric approach, ratios of C, N,
and P are compared in consumers and prey and the
element in shortest supply relative to demand identified
(Hessen 1992). Limiting elements are expected to be
utilized for growth and transferred in food chains with
high efficiency, while nonlimiting elements, by definition present in excess, must be disposed of and may
be recycled. The validity of this approach has been
demonstrated for a wide variety of heterotrophic organisms including bacteria (Tezuka 1990, Anderson
1992), algae (Myklestad 1977, Berman-Frank and Dubinsky 1999), protozoa (Caron and Goldman 1988),
and zooplankton (Elser et al. 1988, Sterner and Hessen
1994, Elser and Urabe 1999).
The fate of excess C across trophic levels is crucial
for our understanding of how C is either buried for
short or long periods as detritus or remineralized to
CO2. Different organisms may cope with this ‘‘problem’’ in different ways (Fig. 1), but there is surprisingly
little knowledge about this. At the cellular level, excess
C as perceived by the consumer may either pass
through the gut unassimilated due to enzymatic regulation that decreases a ssimilation of C-rich compounds
(cf. DeMott et al. 1998), or it may be assimilated and
then excreted either as dissolved organic carbon or respired as CO2 (Darchambeu et al. 2003, Trier and Mattson 2003).
In its first formulation, Liebig’s law was restricted
to the supply of mineral elements to plants as fertilizers,
but later formulations have also attempted to include
other factors such as temperature. In a pure stoichiometric context, however, it is our view that the law of
the minimum should only be applied to the elements
that are used as components in growth; with this use,
the law of the minimum has a direct connection to more
recent ideas related to the chemical stoichiometry of
living biomass. Other factors that influence growth,
such as temperature, do this by changing the rates at
which all the elemental resources are utilized, but without necessarily altering the underlying stoichiometry
(but see Woods et al. [2003]). Ingestad and Ågren
(1995) suggested the interaction between these differ-
STOICHIOMETRIC ECOLOGY
May 2004
STOICHIOMETRY
OF
CELLS
AND
FIG. 2. Growth rate (percentage of maximum RGR) as a
function of element ratios for a terrestrial plant (Betula pendula [circles]) and two aquatic plants (Dunaliella tertiolecta
[squares] and Monochrysis lutheri [triangles]) when grown
under P limitation (solid symbols) or N limitation (open symbols). Data for B. pendula are from Ingestad (1979), Ingestad
and Lund (1979), and Ericsson and Ingestad (1988); data for
D. tertiolecta are from Goldman et al. (1979) and Goldman
and Peavey (1979), and for M. lutheri from Goldman et al.
(1979). Note that there are no data for N:C under N limitation
for D. tertiolecta or N:C data for M. lutheri.
ORGANISMS
What autotrophs do, what heterotrophs do, and what
happens when the two come together
Much has been learned by studying and identifying
the consequences of systematic variations in the C:N:P
stoichiometry of organisms. One key contrast is found
when considering photoautotrophs (hereafter, ‘‘autotrophs’’) vs. heterotrophs, especially Metazoans. While
animals generally tend to maintain body elemental
composition within relatively narrow bounds, autotrophs have a large flexibility in their C:nutrient ratios
and these ratios vary systematically with the growth
rate of the plant. Both N:C and P:C ratios increase when
the relative growth rate (RGR) of an autotroph is increasing, indicating, in relative terms, a larger relative
allocation to active N- and P-based compounds than
C-based structural compounds (Kooijman et al. 2004,
Vrede et al. 2004). This variability in stoichiometry is
characteristic of the overall plasticity of autotrophic
organisms and has its physiological base in the ability
of autotrophs to maintain a balance between the types
of compounds required for growth (Droop 1974). Another cause of variability in stoichiometry comes from
excess uptake of elements, i.e., autotrophs have considerable capacity to accumulate elements in excess of
what is not immediately needed for growth. The ability
to take up an excess of available elements can most clearly
be seen when RGRs of a terrestrial autotroph and an
aquatic autotroph are plotted as a function of nutrient : C
ratios (Fig. 2). The terrestrial plant Betula pendula took
up little N or P above what was required to achieve its
maximum RGR, whereas one aquatic autotroph, Dunaliella tertiolecta, took up N to an excess of 50% and
another one, Monochrysis lutheri, took up P to a concentration five times that required for maximum RGR.
The reasons for, and the consequences of, these two
sources of variability in element ratios are, of course,
Special Feature
ent kinds of factors be called orthogonal to indicate
that they can act on growth independently of each other.
Nevertheless the concept of a single limiting element
is open to question—the complexity of organism physiology means that there may be circumstances where
co-limitation of more than one resource is possible. For
example cultures of yeast can be co-limited by both
oxygen and glucose supply rates (Duboc and von
Stockar 1998), shifting between different degrees of
aerobicity in response to the oxygen:glucose supply
rate.
The concept of stoichiometric regulation is by no
means restricted in application to individual organisms.
The literature abounds with statements inferring elemental limitation of whole systems, such as N limitation in terrestrial and marine systems (Rastetter et al.
1991) and P limitation in freshwater systems (Schindler
1977). Human intervention in the global cycles of carbon, nitrogen, and phosphorus (Vitousek et al. 1997,
Falkowski et al. 2000) is now changing the relative
availability of the elements with potential global effects
on climate and ecosystems. For example, accompanying the well-known alterations in atmospheric CO2
due to combustion of fossil fuels, major changes in
nutrient loading have also occurred in the form of combustion processes and agricultural activity that in recent
decades has led to considerable increases in availability
of N and P to the biosphere (Galloway et al. 1995,
Vitousek et al. 1997, Falkowski et al. 2000). Thus, an
understanding of the factors that couple C to nutrient
fluxes to regulate the sequestration and net balance of
C in ecosystems is imperative (Giardina and Ryan
2000, Grace and Rayment 2000). In this article we
discuss the scaling up of stoichiometric theory for single organisms (the focus of most published work) to
whole ecosystems, focusing in particular on the impact
of nutrient-loading scenarios on C sequestration.
1181
Special Feature
1182
DAG O. HESSEN ET AL.
FIG. 3. Growth rate in Daphnia magna as a function of
algal C mass and algal C:P ratios. Algal food (Selenastrum
capricornutum) was provided at fixed concentrations of 0.5
mg C/L (top panel) or 1.0 mg C/L (bottom panel). The C:P
gradient was obtained by varying P concentrations and light
in chemostat cultures (Hessen et al. 2002).
entirely different and should not be confused. When
comparing C:element ratios in various autotrophs, in
different habitats, or in different seasons, it is therefore
necessary to know whether the element in question is
limiting or not.
In contrast to the great physiological plasticity in
growth form, tissue allocation, and thus biochemical
and elemental composition exhibited by autotrophic organisms, metazoans exhibit far less intraspecific plasticity with a relatively narrow range of interspecific
variation in C:N:P ratios, and a generally nutrient-rich
biomass compared with autotrophs (Andersen and Hessen 1991, Elser et al. 2000, Sterner and Elser 2002).
Thus, when consumers encounter autotroph biomass
with low nutrient content, constraints on animal growth
imposed by limiting nutrients seem inevitable. The net
effect of stoichiometrically imbalanced food on growth
rate is illustrated by growth rates of Daphnia magna
feeding on different quantities and qualities of the
green alga Selenastrum capricornutuum (Fig. 3). While
there is a drop in growth rate from medium (1 mg C/
L) to low (0.5 mg C/L) food levels, reflecting foodquantity effects, growth rate also declines with increas-
Ecology, Vol. 85, No. 5
ing C:P ratio for both food concentrations, illustrating
how limitation by C and P may work in tandem. Given
the frequency of high C:element ratios in autotrophs,
herbivores in both freshwater and terrestrial food webs
may suffer from such stoichiometric constraints (Elser
et al. 2001). Indeed, there is now convincing field evidence for direct P limitation of growth of the aquatic
herbivore Daphnia (Elser et al. 2000, DeMott et al.
2001) and of the terrestrial herbivore Manduca sexta
(Perkins et al., 2004).
Thus, with increasing nutrient imbalance between
autotroph biomass and consumer, there will be less C
diverted up the grazing food chain. This may also in
part explain the commonly poor correlations between
autotroph and grazer biomass. That is, C fixed by photosynthesis does thus not yield increased biomass at
higher trophic levels in any straightforward manner.
This is illustrated by the relationship between phytoplankton biomass in lakes vs. the abundance of Cladocera, the main herbivorous group of freshwater
plankton (Fig. 4). While some of this scatter is likely
caused by grazer–algae dynamics, changes in phytoplankton community structure with increased nutrient
inputs, and top-down control of cladoceran biomass
(i.e., by fish predation), the correlations remain equally
poor even when excluding lakes where fish predation
pressure is considered high (Hessen et al. 2003).
Stoichiometric imbalance also plays a key role in
affecting decomposer and detritivore guilds that encounter nutrient-poor plant detritus. Indeed, such imbalances are likely to be even more severe than in herbivory, at least in the initial stages of detrital processing, as it is common in terrestrial plants for considerable nutrients to be reclaimed from leaves and other
tissues prior to loss (Killingbeck 1993). Of particular
interest are the consequences of differences between
fungi and bacteria in their nutrient requirements. Fungi
FIG. 4. Scatterplot of cladoceran biomass vs. phytoplankton biomass in lakes (both originally in units of mg C/L, prior
to log transformation). Data are from a survey of 110 Norwegian lakes sampled four times during the ice-free season
(i.e., a total of 440 samples).
May 2004
STOICHIOMETRIC ECOLOGY
ENRICHMENT SCENARIOS AND STOICHIOMETRY
OF E COSYSTEMS
In whole ecosystems, the sequestration of C depends
on nutrient inputs, the extent to which these nutrients
control C flow along food chains, and the position and
strengths of the C sinks within the overall flow network.
Special Feature
appear to have low N:C or P:C ratios while the heterotrophic bacteria have typically high (yet somewhat
flexible) N:C and P:C ratios (Sterner and Elser 2002,
Vadstein 2000), implying that fungi and bacteria may
have different roles in mineralization. Such stoichiometric relations are important for maintaining ecosystem element cycles with the consequence that soils are
relatively richer in N than C relative to plants. Indeed,
decomposition of soil organic matter, a key process in
soil ecosystem functioning (Griffiths et al. 2000), may
be limited by C, N, or P depending on soil type, management, and vegetation. Stoichiometric effects likely
extend into other parts of soil food webs. As for invertebrate herbivores, invertebrate decomposers have
typically one order of magnitude higher N and P concentrations in their bodies than fresh plant litter (Swift
et al. 1979). By processing soil organic matter with a
high C:N ratio, decomposers require extra N to produce
new microbial biomass. In situations in which the pool
of available inorganic N is limited, the energy-conversion efficiency (yield) or the consumption rate of
the microbes, and hence decomposition, will decrease
(Ågren et al. 2001). A distinguishing feature of the
decomposition system is that the decomposers start
with a material rich in C and hence are N or P limited
but, as the substrate is used, C is released and the
limiting element incorporated until a critical C:element
ratio is reached where C becomes limiting. The decomposers can then release N and P and make these
nutrients available for uptake by plants again. These
critical break points in C:N and C:P ratios are welldefined stoichiometric quantities (Ågren and Bosatta
1998). Their values, however, are coupled to inorganicnutrient availability (Ågren et al. 2001). A consequence
of the use of inorganic N and P is that organic-matter
processing by decomposers may be much less constrained by the C:nutrient ratios than by herbivore linkages.
In summary, we have provided a brief overview of
factors influencing organismal stoichiometry and how
this bears on C processing in ecological systems. There
are many ongoing studies examining nutrient limitation
of autotrophs and heterotrophs, and the fate of resulting
excess C. Certainly there is still much to be learned
here. In particular, the Liebig-type approach to determining limiting factors is simplistic and becoming outdated. New approaches to understanding and modeling
stoichiometric interactions will need to consider in
greater detail the complex interplay between elements,
biochemical constraints, and the physiology of organisms (Kooijman 1998, Anderson et al. 2004).
1183
FIG. 5. Schematic diagram of flows of carbon in marine
and terrestrial ecosystems. Major C sinks are indicated in
boldface type.
Diagrams illustrating C flows and sinks in forest and
open-ocean ecosystems are shown in Fig. 5. Major
sinks in terrestrial systems are aboveground biomass,
roots, and soils. The living biomass supported by aquatic systems is low, but C stores in freshwater sediments
may be of greater importance than commonly assumed
(cf. Dean and Gorham 1998). The major sinks in the
open ocean are dissolved organic and inorganic C
(along with burial in sediments on long time scales).
Continental margins also play a major role in ocean
biogeochemical cycles, as they are burial sites of organic C derived from both local production and terrestrial sources. The added complexity of whole ecosystems in comparison to individual organisms gives
rise to the expectation that the stoichiometric control
of overall nutrient and C cycling is likely to be similarly
complex.
DAG O. HESSEN ET AL.
1184
Changes in C storage capacity in ecosystems can be
explained in terms of three biogeochemical factors
(Rastetter et al. 1992): the total supply of nutrients from
external sources, the distributions of C, N, and other
elements between different system components, and
alterations in the stoichiometric ratios of the components themselves. In this section we describe the biogeochemical pathways leading to C sequestration in
terrestrial and aquatic systems, and examine the importance of stoichiometry in determining fluxes to C
sinks, using loading scenarios (topical in the context
of climate change) for illustrative purposes. We propose that incorporating organism-level responses and
stoichiometric mechanisms operating at key points in
ecosystem flows are central in understanding these patterns.
Special Feature
Terrestrial systems
Vegetation and soils are major global C sinks, 640
and 1358 Gt C/yr, respectively (Cao and Woodward
1998). Many terrestrial ecosystems are considered to
be N limited in the sense that added N will produce a
plant growth response and increase C storage (e.g.,
Vitousek and Howarth 1991). Carbon sequestration depends on the distribution of C and nutrient elements
between vegetation and soils, because the C:nutrient
stoichiometry of vegetation and soils differs greatly
(Townsend et al. 1996). Furthermore, there are tight
couplings between ambient CO2 and N uptake of terrestrial vegetation because of the key role of N-rich
rubisco in C fixation (see Vrede et al. 2004). Under a
prevailing N limitation of terrestrial autotrophs (Gutschik 1981, Schlesinger 1991), increased deposition of
N may sequester more CO2 and elevated CO2 may allow
for increased uptake of N (Schindler and Bayley 1993),
provided that C:N ratios remain rather constant. Modeling studies of the response of the terrestrial biosphere
to elevated CO2 indicate increases in net primary productivity and also in vegetation C:N ratio (Rastetter et
al. 1997). Long-term effects of CO2 on forest growth
are intimately linked with the availability of N (Oren
et al. 2001). Peterson and Melilo (1985) calculated that,
assuming C:N ratios in vegetation and soils remain
unchanged and that 60% of N in precipitation is invested in net growth, enhanced N deposition resulting
from fossil-fuel combustion could lead to significantly
increased storage in forest ecosystems. Similarly, Mäkipää et al. (1999) modeled forest response to increased
N deposition and predicted an 11% increase in total C
stock (vegetation, litter, and soil organic matter). On
the other hand, a further increase in N loadings may
reduce the uptake efficiency due to combined ammonia
and nitrate acidification or shift to other limiting constituents. This ‘‘N saturation’’ causes increased leaching of N from soils due to ‘‘saturation’’ effects (Aber
et al. 1989, Stoddard 1994), with increased levels of
NO23 in surface water.
Ecology, Vol. 85, No. 5
The ability of terrestrial ecosystems to sequester C
also depends on individual plant traits due to the effects
of competition for light and nutrients in determining
biomass allocation in plants (Tilman 1988). Nutrient
deficiency enhances biomass partitioning to roots; conversely, increased nutrient supply induces decreased
root : shoot ratio (Ericsson 1995, Andrews et al. 2001).
The reduced water stress associated with elevated CO2
also tends to decrease root : shoot ratio, particularly in
plants not subject to severe nutrient limitation (Friedlingstein et al. 1999). The combination of high C:N
ratio (.300) and long lifetimes in wood (.100 years)
(Vitousek et al. 1988, Schimel et al. 1994) means that
any sustained decrease in root : shoot ratio could significantly increase overall storage of C and decrease
nutrient retention in forest ecosystems. Moreover it is
well known that the N content of leaves and other organs of plants grown under elevated CO2 are lower than
in plants cultivated at ambient CO2 (Ceulemans and
Mousseau 1994, Curtis 1996, Curtis and Wang 1998,
Norby et al. 2001), creating potential for increased C
sequestration. The extent to which the C:N ratio of
woody tissues might respond to CO2 is however less
clear. Most studies have only looked at effects on N
and, in a compilation of results of various studies, Cotrufo et al. (1998) report a mean plant-tissue N concentration in elevated CO2 of 86% of the concentration
in ambient CO2. Variability in the extent of nitrogencontent response is, however, large among studies but
there are some systematic responses: nitrogen-fixing
plants and C4 plants show little response in N content
to altered CO2 regime. Data regarding elements other
than N are too limited to allow any general conclusions.
However, a study by Overdieck (1993) indicates that
tissue N:P might decrease under elevated CO2 but the
ratio of N to other elements remains unchanged. The
extent to which the C:N ratio of woody tissues might
be responsive is however less clear.
Another aspect to consider in an ecosystem context
is retranslocation of elements before tissue is shed,
leading to large reduction in element:C ratio in the litter
produced. Although the biochemical processes involved in retranslocation might differ for N and P, there
seems to be no major difference in efficiency in their
retranslocation (Aerts 1996, Killingbeck 1996). Whether this pattern remains under an increasing atmospheric
CO2 concentration remains to be evaluated. However,
decreases in the N:C ratio of live tissue in CO2-fertilization experiments with plants do not reliably translate
to similar changes in plant litter (Norby and Cotrufo
1998). Thus, consequences for turnover rates of litter
as a result of a changing N:C ratio are not clear. Although it has frequently been shown that the N:C and
P:C ratios reflect the decomposability of litter (e.g.,
Melillo et al. 1982, Enriquez et al. 1993), these ratios
could also reflect the importance of differences in carbon chemistry to the decomposer community rather
STOICHIOMETRIC ECOLOGY
May 2004
Aquatic systems
Inorganic C is abundant in water, and so any increase
in concentration resulting from anthropogenic CO2
emissions might be expected on first glance to have
little or no impact on phytoplankton growth and subsequent food-web dynamics. Consider, for example, the
‘‘biological pump’’ in the ocean that transports C via
sinking particles and dissolved organic matter (DOM)
from the surface euphotic zone to below the permanent
thermocline, where it is eventually sequestered (in either organic or inorganic form). If photosynthesis is
not limited by C, then it seems that the biological pump
cannot sequester anthropogenic CO2. It should however
be noted that sequestration of anthropogenic CO2 does
occur as a result of the ‘‘solubility pump,’’ which is
the uptake of C by air–sea exchange driven by undersaturation of pCO2 with respect to the atmosphere and
subsequent transport of C-enriched waters to depth by
physical processes (Follows et al. 1996). Nevertheless
climate change will impact the biological pump
through its effect on the natural C cycle in the ocean.
For example, changes in circulation as a result of warming may alter the supply of nutrients to the surface
ocean as well as the light environment experienced by
phytoplankton. These changes in turn affect plankton
community structure and export of organic matter.
The effects of changing nutrient supply on C cycling
are commonly associated with freshwater systems because of their close proximity to human activity. Most
lakes are considered to be P limited, based on close
correlations between total (or particulate) P and particulate C (Fig. 6) or some other measure of biomass.
Adding more P (and subsequently N) leads to eutrophication; primary production will draw on CO2 dissolved in water, hence more atmospheric CO2 will enter
the aquatic system for conversion to algal biomass,
potentially sedimenting and thus being sequestered into
a C sink. In fact, freshwater sediments may be a globally significant, yet neglected organic-C sink (Dean and
Gorham 1998). While lake sediments may represent an
important storage pool for organic C, recent data suggest that many lakes commonly are net conduits of CO 2
(del Giorgio and Peters 1993, Cole et al. 1994) due to
allochthonous imports of organic C from their watersheds. While CO2 fluxes to some extent are governed
by the pCO2 gradients across the water–air boundary
layer, biological processes and stoichiometric balance
will also be major determinants of net CO2 balance.
Freshwater systems are also prone to large variations
in CO2, and even a moderate increase in CO2 could
yield increased C:P ratios in algae (Burkhart and Riebsell 1997, Urabe et al. 2003). This effect is demonstrated in the freshwater chlorophyte Selenastrum capricornutum in Fig. 7.
Loading of organic carbon diverted into heterotrophs
will increase the output of oxidized C, while increased
loading of P and N will increase primary production
and thus work in the opposite direction. Based on a
survey of published data, del Giorgio and Peters (1993)
found that the epilimnetic P:R (primary production to
community respiration) ratio rises above 1 only in rather productive lakes. Their empirically derived equations for production (P) and respiration (R) in lakes as
related to algal biomass (expressed as chlorophyll a,
in micrograms per liter) were: P 5 10.3 3 chl1.19 and
R 5 59 3 chl0.56. In their model P and R reached 1 at
14 mg chlorophyll/L, below which the lakes became
net sources of CO2. Since algal biomass and chlorophyll are proxies of phosphorus concentrations, this can
be recast as a rough expression of the net exchange of
CO2 or the P:R ratio as a function of the P concentration, illustrating the bearing of P load on net C sequestration (Fig. 8).
Carbon export in the ocean depends on the balance
between recycling of material in surface waters and
export to depth, the former dominating in most instances. Thus it is important to understand the stoichiometry
of processes controlling trophic dynamics in marine
systems, and in particular those closely related to the
production of particles and dissolved organic matter
that are exported. Much of the particulate export is in
detritus, consisting of nonliving phytoplankton aggregates (‘‘marine snow’’, Lampitt et al. 1993) and zooplankton fecal pellets. This material is clearly subject
Special Feature
than the importance of differences in stoichiometry per
se (Joffre and Ågren 2001).
Elevated CO2 also increases the direct input of organic C into the soil through plant exudation (van Veen
et al. 1991, Dukes and Field 2000). The extent to which
such effects will increase the mass of litter and soil C
pools depends on how this extra C supply affects microbial decomposition, which is constrained by the
need for microbes to maintain their C:nutrient balance
(Paul and Clark 1989). For example, if an increase in
the C:N ratio of soil organic matter results in N limitation of microbial populations, decomposition could
slow under elevated CO2 leading to increased mass in
litter and soil C pools. When considering the impact
of climate change on soil C storage it is however necessary to consider not only response to elevated CO 2,
but also the simultaneous effects of temperature (Gorissen et al. 1995). Climate warming is thought to reduce
soil C by increasing microbial demand for C through
increased respiration (Kirschbaum 2000). If the N
thereby remineralized is sequestered into aboveground
woody biomass, C stored in forest ecosystems as a
whole could increase (Rastetter et al. 1991). Soil C:N
ratios also have strong bearings on aquatic recipients,
since the export of dissolved organic C (DOC) to surface waters has a strong postive correlation with the
soil C:N ratio (Aitkenhead and McDowell 2000). Ecosystem export of DOC increased linearly from ,5 to
;100 kg C·ha21·yr21 over a soil C:N ratio of ;13–30
(by mass).
1185
Special Feature
1186
DAG O. HESSEN ET AL.
Ecology, Vol. 85, No. 5
FIG. 6. Regressions of particulate (seston) fractions of carbon (PC), nitrogen (PN), and phosphorus (PP) based on a
survey of 110 Norwegian lakes sampled four times during the ice-free season (i.e., a total of 440 samples).
FIG. 7. Phytoplankton C:P ratios along a CO2 concentration gradient in ambient medium. Algal cultures (Selenastrum
capricornutum) were cultured in flow-through chemostats at
low dilution rate (280 mL/d) with buffered medium to avoid
changes in pH. Different CO2 concentration levels were produced by bubbling equilibrated industrial-grade gas through
the chemostats. There were four replicates for each treatment.
Data are means and 1 SD.
FIG. 8. Net CO2 balance in a large number of lakes as
predicted from the model of del Giorgio and Peters (1993).
The dashed lines indicate the community production:respiration ratio (P:R) 5 1 and the corresponding concentration
of total P (TP). For further explanation see Enrichment scenarios. . . : Aquatic systems.
May 2004
STOICHIOMETRIC ECOLOGY
imate limiting factors is therefore far from clear-cut
(Falkowski 2000). Deposition of macronutrients such
as N, or micronutrients such as iron, from the atmosphere may have important impacts on the biological
pump over long time scales. For example, nutrient supply is thought to have mediated variations in the sequestration of CO2 in the oceans that have accompanied
glacial–interglacial cycles over the past two million
years (Falkowski et al. 2000, Ganeshram et al. 2000).
Understanding the role of iron in regulating C sequestration in the ocean presents a particularly difficult
challenge for the scientific community. There are many
poorly quantified processes relating to the cycling of
iron in the ocean, notably the fraction of iron of aeolian
origin that becomes soluble and bioavailable upon deposition in surface waters, and rates of biological uptake and scavenging (Wells et al. 1995, Johnson et al.
1997).
The stoichiometry of marine seston has been widely
assumed to closely conform to the Redfield ratio of
106C:16N:1P (Redfield et al. 1963), which has proved
to be particularly convenient for biogeochemical modelers. This assumption seems surprising given the diversity of processes described above. However, the C:N
ratio of particulate organic matter in surface waters
does not often deviate far from Redfield (e.g., Chen et
al. 1996) and nutrient ratios in the deep ocean, reflecting long-term ratios of remineralization, are also close
to the Redfield value (Anderson and Sarmiento 1994).
The relatively constant Redfield ratios of dissolved nutrients and C in the deep ocean suggest that particles
are exported in the same ratio, and that there is only a
minor contribution from C-rich dissolved organics
(Kähler and Koeve 2001). However, as already discussed, element ratios in phytoplankton can deviate
significantly from Redfield (e.g., Daly et al. 1999), and
it is well known that zooplankton (e.g., Gismervik
1997) and bacteria (e.g., Goldman et al. 1987) tend to
have C:N ratios lower than the Redfield value. Geider
and La Roche (2002) concluded that, both on the basis
of laboratory C:N:P data and theoretical approaches to
biochemical composition, there was little evidence to
support the contention that the Redfield ratio reflects
a physiological or biochemical constraint on the elemental composition of primary production, a conclusion also reinforced by Sterner and Elser (2002). Therefore, there is no a priori reason to believe that cycling
and export must occur in Redfield proportions. Moreover, the C:N ratio of sinking particles is generally
believed to increase with depth because of preferential
remineralization of nutrient elements relative to C
(Thomas et al. 1999). Although Redfield ratios have
become an accepted empirical observation, a conceptual understanding of their biological and ecological
basis is lacking (Falkowski 2000). Given that physiological constraints do not appear to be responsible, it
must be that ecological feedback mechanisms including the regulation of total ocean N content involving
Special Feature
to processing by heterotrophs, and therefore might be
expected to have elevated C:N ratios if nutrient elements are limiting to consumers. For example, it is
known that the C:N ratio of zooplankton fecal pellets
is variable and sometimes high (see Anderson 1994).
The C:N ratio of dissolved organic matter (DOM)
may often be significantly higher than the Redfield ratio
(e.g., Søndergaard et al. 2000). DOM in aquatic systems is produced via various processes such as phytoplankton exudation, zooplankton ‘‘sloppy feeding,’’
viral lysis of bacteria and phytoplankton, and dissolution of detritus (Nagata 2000). The resulting products
often exhibit a C-rich seasonal accumulation (Williams
1995), in which case export of this material is potentially an important route for C sequestration in the
ocean. High C:N in marine DOM is likely due to phytoplankton extracellular release of organic C, with no
associated N (Anderson and Williams 1998). This release may be a result of photosynthetic overflow under
nutrient-limiting conditions, and so the highest relative
rates are generally associated with oligotrophy (Teira
et al. 2001). Another potential C-rich export route is
via transparent exopolymer particles (TEP), which are
polymeric gel particles that form abiotically from dissolved and colloidal organic matter. Although TEP do
not sink gravitationally, they are exported as aggregates
(Passow et al. 2001). Production rates of DOM and
TEP are in general inversely related to nutrient supply.
Nutrient loading could therefore change the internal
system stoichiometry by decreasing this source of organic C—an example of where C sequestration is potentially inversely related to the supply of nutrients
limiting primary production.
In addition to production of organic C, many marine
organisms, including coccolithophorids, foramaniferans, corals, and molluscs, produce calcium carbonate
(CaCO3). The ecology of these organisms is relatively
poorly understood, and it is by no means clear how
climate change will affect their competition with other
algal groups. High abundance of these carbonate-forming biota has been associated with high N:P ratios (Tyrrell and Taylor 1996), and so any shifts in these two
elements, e.g., due to proliferation of nitrogen fixers
or deposition of atmospheric N, could favor them. It
is indeed somewhat enigmatic that N can be a major
limiting element in the biosphere given the enormous
amounts of N2 in the atmosphere and the existence of
N-fixing plants. Nitrogen is generally considered to be
the ‘‘proximate limiting nutrient’’ in marine systems,
representing local limitation according to Liebig’s law,
while phosphate supplied by continental weathering
and fluvial discharge is viewed as the ‘‘ultimate limiting nutrient’’ influencing system productivity on long
time scales (Tyrell 1999). Shortfalls of N can be compensated for in the long term through N2 fixation, although availability of trace elements, notably iron, and
light limitation may limit this process (Martin 1990).
This distinction between P and N as ultimate and prox-
1187
DAG O. HESSEN ET AL.
1188
nitrogen-fixing and denitrifying organisms are important. Nevertheless, any change in the C:nutrient ratio
of export of organic matter could have a major bearing
on C sequestration in the ocean (e.g., Heinze et al.
1991) and so this issue bears serious consideration in
any attempt to model the ocean’s role in the C cycle
simultaneous with an accounting of nutrient processing.
Special Feature
DISCUSSION
Ecosystem responses reflect the net sum of processes
at the cellular or organismal level and their interaction
with abiotic factors in time and space. In both organisms and ecosystems the production efficiency in terms
of C uptake and C sequestration depends strongly on
the balance between intake of and demand for key nutrient elements. Nutrients like N or P will be recycled
and reused by autotrophs, while excess C may either
be oxidized to inorganic forms by the organisms themselves or heterotrophic decomposers, or be more permanently buried in sediment or soils. These similarities
in feedbacks arise not because ecosystems mimic organismal allocation rules for C, N, or P, but because
the interacting biota themselves that comprise ecosystems must obey stoichiometric rules at the level of their
individual cells and bodies that greatly constrain the
range of values that ecosystem-level pools and processes can attain. Thus, in the simplest scenarios, the
fate or sequestration of C will be determined by the
absolute and relative demands for C, N, and P of each
organism relative to its resources. In food webs the
largest discrepancy in element ratios is found between
autotrophs and herbivores, causing low efficiency in
the use of C (i.e., much C ‘‘in excess’’). The evolutionary explanation for such a mismatch in element
ratios between autotrophs and grazers may, from the
autotroph’s perspective, be seen as an adaptation to
cause lower grazer fitness (at some expense to its own
growth, however; cf. Moran and Hamilton 1980, White
1993), while the grazers dilemma appears to be a tradeoff between rapid growth and the risk of nutrient limitation. This trade-off appears to exist because of a
general link between high levels of key nutrients (especially P), high levels of important macromolecules
(e.g., P-rich ribosomal RNA), and growth rate (Elser
et al. 2000, Vrede et al. 2002, 2004). Thus, high and
low body levels and demands for P (and to a lesser
extent N) in consumers may be seen as stoichiometric
manifestations of evolutionary adaptations along the
classic r–K axis. In any case, these inexorable constraints mean that herbivores rarely find optimal food,
and that stoichiometrically imbalanced autotroph production is allowed to go elsewhere.
To the extent that other organisms in the ecosystem
remain limited by inadequate sources of fixed energy,
one should expect strong selective pressure towards
utilization of this excess C. Autotrophs may allocate
this excess C to C-rich compounds such as structural
Ecology, Vol. 85, No. 5
matter or exudates for various purposes such as digestive screens (algae) or grazer repellents. Under pressure
from selective grazers, it could also pay to maintain
high C:N or C:P ratios in grazable tissue to reduce the
risk of its consumption and/or to depress grazers (cf.
White 1993). However, this strategy would not work
in communities with nonselective grazers (e.g., aquatic
systems dominated by filter feeders) unless some
group-selection argument is invoked for the phytoplankton community. Excess C in the form of carbohydrates, and exchange of carbohydrates and nutrients,
could also be seen as a stoichiometric basis for the
ecological stabilization and evolution of symbiosis (cf.
Berman-Frank and Dubinsky 1999, Sterner and Elser
2002, Kooijman et al. 2004).
Following the same line of reasoning, one could argue that the organisms that comprise ecosystems occupy different stoichiometric niches. The ability to explore extremely C-rich (and stoichiometrically unbalanced) food like woody material and phloem implies
that organisms such as fungi, termites, wood lice and
aphids have evolved to use major sources of energy
that for other organisms would be largely inadequate
or inaccessible due to nutrient imbalance. Such adaptations could involve gut symbionts and sophisticated
means to metabolically reclaim N (as in termites; Higashi et al. 1992), the ability to flux large amounts of
unutilized organic carbon (e.g., honeydew production
by aphids), or the development of reduced body demands for limiting elements (Fagan et al. 2002, Jaenike
and Markow 2003). This again would have consequences for growth rates and life-cycle strategies. A
stoichiometric niche could also mean that organisms
spatially prefer some more stoichiometrically adequate
areas to others. At a small scale this could mean preference for the tip of buds rather than mature leaves for
herbivorous insects; at a larger scale it could imply
migration of ungulates to seek the most nutritious herbs
(White 1993, van der Wal et al. 2000).
At an even larger scale, the length and stability of
food chains could also be determined by stoichiometric
principles (Hastings and Conrad 1979, Hairston and
Hairston 1993). As argued by Hastings and Conrad
(1979), an evolutionarily stable strategy for food chains
would imply that food chains should collapse to the
shortest possible length compatible with biochemical
and physiological constraints (usually three steps).
Thus, the assumption that is it the trophic structure that
controls the fraction of energy consumed at each trophic levels, rather than the energetics controlling trophic structure (Hairston and Hairston 1993) may indeed
be questioned when seen from a stoichiometric point
of view.
To extrapolate from organisms to ecosystems may
be a dubious process, but the processes described above
may also impinge on large-scale ecosystem properties.
The unavoidable imbalances between autotrophs and
consumers appear to have the net effect that overall
May 2004
STOICHIOMETRIC ECOLOGY
1985, Sterner and Elser 2002). Whether or not such
extrapolations are accurate upon closer examination, it
seems likely that a better understanding of the biological coupling of multiple chemical elements in evolving
biota will improve our ability to predict the future dynamics of the biosphere under an increasing anthropogenic influence and to understand the consequences
of those changes for diverse ecosystems.
ACKNOWLEDGMENTS
DOH acknowledges the European Science Foundation for
supporting the workshop leading up to this Special Feature.
T.R. Anderson is funded by the Natural Environment Research Council, U.K. J. Elser acknowledges NSF Integrated
Research Challenges for Environmental Biology Grant DEB9977047.
LITERATURE CITED
Aber, J. D., K. J. Nadelhoffer, P. Steudler, and J. M. Melillo.
1989. Nitrogen saturation in northern forest ecosystems:
hypothesis and implications. BioScience 39:378–386.
Aerts, R. 1996. Nutrient resorption from senescing leaves of
perrennials: are there general patterns? Journal of Ecology
84:597–608.
Ågren, G. I., and E. Bosatta. 1998. Theoretical ecosystem
Ecology—understanding element cycles. Cambridge University Press, Cambridge, UK.
Ågren, G. I., E. Bosatta, and A. M. Magill. 2001. Combining
theory and experiment to understand effects of inorganic
nitrogen on litter decomposition. Oecologia 128:94–98.
Aitkenhead, J. A., and W. H. McDowell. 2000. Soil C:N ratio
as a predictor of annual riverine DOC flux at local and
global scales. Global Biochemical Cycles 14:127–138.
Andersen, T., and D. O. Hessen. 1991. Carbon, nitrogen and
phosphorus content in common crustacean zooplankton
species. Limnology and Oceanography 36:807–814.
Anderson, L. A., and J. L. Sarmiento. 1994. Redfield ratios
of remineralization determined by nutrient data analysis.
Global Biogeochemical Cycles 8:65–80.
Anderson, T. R. 1992. Modelling the influence of food C:N
ratio and respiration on growth and nitrogen excretion in
marine zooplankton and bacteria. Journal of Plankton Research 14:1645–1671.
Anderson, T. R. 1994. Relating C:N ratios in zooplankton
food and faecal pellets using a biochemical model. Journal
of Experimental Marine Biology and Ecology 184:183–
199.
Anderson, T. R., M. Boersma, and D. Raubenheimer. 2004.
Stoichiometry: linking elements to biochemicals. Ecology
85:1193–1202.
Anderson, T. R., and P. J. le B. Williams. 1998. Modelling
the seasonal cycle of dissolved organic carbon at station
E1 in the English Channel. Estuarine, Coastal and Shelf
Science 46:93–109.
Andrews, M., J. A. Raven, and J. J. Sprent. 2001. Environmental effects on dry matter partitioning between shoot
and root of crop plants: relations with growth and shoot
protein concentration. Annals of Applied Biology 138:57–
68.
Berman-Frank, I., and Z. Dubinsky. 1999. Balanced growth
in aquatic plants: myth or reality? BioScience 49:29–37.
Burkhardt, S., and U. Riebesell. 1997. CO2 availability affects elemental composition (C:N:P) of the marine diatom
Skeletonema costatum. Marine Ecology Progress Series
155:67–76.
Cao, M., and F. I. Woodward. 1998. Net primary and ecosystem production and carbon stocks of terrestrial ecosystems and their responses to climate change. Global Change
Biology 4:185–198.
Special Feature
transfer of primary production to higher trophic levels
is impaired under conditions of nutrient-limited autotroph production. As a result, much C can be sequestered in standing biomass of nutritionally unsuitable
autotroph biomass with high C:nutrient ratio. This further affects the fate of production by increasing the
percentage of production entering detrital pools to be
entrained into slow detrital food webs, where the efficiency of C processing is low due to the detritus’ low
nutrient content. This is illustrated by the cross-ecosystem comparison of Cebrian (1999). Systems dominated by autotrophs with low C:element ratios and fast
turnover rates such as phytoplankton communities are
effectively grazed and little C is sequestered to recalcitrant detrital pools, while terrestrial systems dominated by high C:element ratios and low turnover sequester most of the fixed C as detritus.
The balance of carbon fixation by photosynthesis and
release by oxidation processes is the basis of the global
C cycle. The principles of this cycle have been known
for a long time, yet the various factors regulating this
balance and dynamics are still imperfectly understood
(Falkowski et al. 2000). In a global-change context, an
understanding of the parameters regulating the sequestration and net balance of C in ecosystems is imperative (Giardina and Ryan 2000, Grace and Rayment
2000). Although the cumulative effects of anthropogenic C emissions are implicated as the primary cause
of increased atmospheric CO2 concentrations, other
ecosystem processes still account for some 95% of the
fluxes. Importantly, cycling of C occurs in tandem with
nutrient elements such as N and P, both of which have
also been subject to major changes in biogeochemical
cycling (Falkowski et al. 2000). As described above,
the efficiencies with which different elements are cycled or sequestered in organisms and ecosystems depends on their relative availabilities. Thus, major
changes in fluxes of C to and from the atmosphere may
be induced by relatively subtle shifts in the C:nutrient
ratios of particular pools or in the relative importance
of different pools having different C:nutrient ratios.
Since global perturbations of nutrient cycles are much
larger than current anthropogenic impacts on C cycling
(Falkowski et al. 2000), a better understanding of how
nutrient cycles interact with C cycling is imperative.
While following the balance of multiple elements in
individual organisms and its implications for specific
ecological interactions are reasonably straightforward,
application of stoichiometric principles at larger scales
should also help greatly in understanding the fate of
fossil-fuel C and its interactions with anthropogenic
impacts on limiting nutrient availability. Following the
axiom of Liebig limitation, one simple implication of
these principles in a global context is that biospheric
storage of C is limited relative to current levels of C
release to the atmosphere because C:N and C:P ratios
in C sources (fossil fuels) are much larger than the
ratios in potential modern sinks (Peterson and Melillo
1189
Special Feature
1190
DAG O. HESSEN ET AL.
Caron, D. A., and J. C. Goldman. 1988. Dynamics of protistan carbon and nutrient cycling. Journal of Protozoology
35:247–249.
Cebrian, J. 1999. Patterns in the fate of production in plant
communities. American Naturalist 154:449–468.
Ceulemans, R., and M. Mousseau. 1994. Effects of elevated
CO2 on woody plants. New Phytologist 127:425–446.
Chen, C.-T. A., C.-M. Lim, B.-T. Huang, and L.-F. Chang.
1996. Stoichiometry of carbon, hydrogen, nitrogen, sulfur
and oxygen in the particulate matter of the western North
Pacific marginal seas. Marine Chemistry 54:179–190.
Cole, J. J., N. F. Caraco, G. W. Kling, and T. K. Kratz. 1994.
Carbon-dioxide supersaturation in the surface waters of
lakes. Science 265:1568–1570.
Cotrufo, M. F., P. Ineson, and A. Scott. 1998. Elevated CO2
reduces the nitrogen concentration of plant tissues. Global
Change Biology 4:43–54.
Curtis, P. S. 1996. A meta-analysis of leaf gas exchange and
nitrogen in trees grown under elevated carbon dioxide.
Plant, Cell and Environment 19:127–137.
Curtis, P. S., and X. Wang. 1998. A meta-analysis of elevated
CO2 effects on woody plant mass, form, and physiology.
Oecologia 113:299–313.
Darchambeau, F., P. Færovig, and D. O. Hessen. 2003. How
Daphnia copes with excess carbon in its food. Oecologia
136:336–346.
Daly, K. L., D. W. R. Wallace, W. O. Smith, Jr., A. Skoog,
R. Lara, M. Gosselin, E. Falck, and P. L. Yager. 1999. NonRedfield carbon and nitrogen cycling in the Arctic: effects
of ecosystem structure and dynamics. Journal of Geophysical Research 104:3185–3199.
Dean, W. E., and E. Gorham. 1998. Magnitude and significance of carbon burial in lakes, reservoirs, and peatlands.
Geology 26:535–538.
del Giorgio, P. A., and R. H. Peters. 1993. Balance between
phytoplankton production and plankton respiration in lakes.
Canadian Journal of Fisheries and Aquatic Sciences 50:
282–289.
DeMott, W. R., R. D. Gulati, and K. Siewertsen. 1998. Effects
of phosphorus-deficient diets on the carbon and phosphorus
balance of Daphnia magna. Limnology and Oceanography
43:1147–1161.
DeMott, W. R., R. D. Gulati, and E. Van Donk. 2001. Effects
of dietary phosphorus deficiency on the abundance, phosphorus balance, and growth of Daphnia cucullata in three
hypereutrophic Dutch lakes. Limnology and Oceanography
46:1871–1880.
Droop, M. R. 1974. The nutrient status of algal cells in continuous culture. Journal of the Marine Biological Association (UK) 54:825–855.
Duboc, P., and U. von Stockar. 1998. Dual limitations: kinetic
and stoichiometric analysis. Thermochimica Acta 309:121–
128.
Dukes, J. S., and C. B. Field. 2000. Diverse mechanisms for
CO2 effects on grassland litter decomposition. Global
Change Biology 6:145–154.
Elser, J. J., M. M. Elser, N. A. Mackay, and S. Carpenter.
1988. Zooplankton-mediated transitions between N-limited and P-limited algal growth. Limnology and Oceanography 33:1–14.
Elser, J. J., W. F. Fagan, R. F. Denno, D. R. Dobberfuhl, A.
Folarin, A. Huberty, S. Interlandi, S. S. Kilham, E. McCauley, K. L. Schulz, E. H. Siemann, and R. W. Sterner.
2000. Nutritional constraints in terrestrial and freshwater
food webs. Nature 408:578–580.
Elser, J. J., K. Hayakawa, and J. Urabe. 2001. Nutrient limitation reduces food quality for zooplankton: Daphnia response to seston phosphorus enrichment. Ecology 82:898–
903.
Ecology, Vol. 85, No. 5
Elser, J. J., and J. Urabe. 1999. The stoichiometry of consumer-driven nutrient recycling: theory, observations, and
consequences. Ecology 80:735–751.
Enriquez, S., C. M. Duarte, and K. Sand-Jensen. 1993. Patterns in decomposition rates among photosynthetic organisms—the importance of detritus C:N:P content. Oecologia
94:457–471.
Ericsson, T. 1995. Growth and shoot:root ratio of seedlings
in relation to nutrient availability. Plant and Soil 168/169:
205–214.
Ericsson, T., and T. Ingestad. 1988. Nutrition and growth of
birch seedlings at varied relative phosphorus addition rates.
Physiologia Plantarum 7:227–235.
Fagan, W. F., E. H. Siemann, R. F. Denno, C. Mitter, A.
Huberty, H. A. Woods, and J. J. Elser. 2002. Nitrogen in
insects: implications for trophic complexity and species
diversification. American Naturalist 160:784–802.
Falkowski, P. G. 2000. Rationalizing elemental ratios in unicellular algae. Journal of Phycology 36:3–6.
Falkowski, P. G., et al. 2000. The global carbon cycle: a test
of our knowledge of Earth as a system. Science 290:291–
296.
Follows, M. J., R. G. Williams, and J. C. Marshall. 1996.
The solubility pump of carbon in the subtropical gyre of
the North Atlantic. Journal of Marine Research 54:605–
630.
Friedlingstein, P., G. Joel, C. B. Field, and I. Y. Fung. 1999.
Toward an allocation scheme for global terrestrial carbon
models. Global Change Biology 5:755–770.
Galloway, J. N., W. H. Schlesinger, H. L. Levy, A. Michaels,
and J. L. Schnoor. 1995. Nitrogen fixation: an anthropogenic enhancement–environmental response. Global Biogeochemical Cycles 9:235–252.
Ganeshram, R. S., T. F. Pedersen, S. E. Calvert, G. W.
McNeill, and M. R. Fontugne. 2000. Glacial–interglacial
variability in denitrification in the world’s oceans: causes
and consequences. Paleoceanography 15:361–376.
Geider, R. J., and J. La Roche. 2002. Redfield revisited: variability of C:N:P in marine microalgae and its biochemical
basis. European Journal of Phycology 37:1–17.
Giardina, C. P., and M. G. Ryan. 2000. Biogeochemistry—
soil warming and organic carbon content—reply. Nature
408:790.
Gismervik, I. 1997. Stoichiometry of some marine planktonic
crustaceans. Journal of Plankton Research 19:279–285.
Goldman, J. C., D. A. Caron, and M. R. Dennett. 1987. Regulation of gross growth efficiency and ammonium regeneration in bacteria by substrate C:N ratio. Limnology and
Oceanography 32:1239–1252.
Goldman, J. C., J. J. McCarthy, and D. G. Peavey. 1979.
Growth rate influence on the chemical composition of phytoplankton in oceanic waters. Nature 279:210–214.
Goldman, J. C., and D. G. Peavey. 1979. Steady-state growth
and chemical composition of the marine chlorophyte Dunaliella tertiolecta in nitrogen-limited continuous cultures.
Applied Enviromental Microbiology 38:894–901.
Gorissen, A., J. H. van Ginkel, J. J. B. Keurentjes, and J. A.
van Veen. 1995. Grass root decomposition is retarded when
grass has been grown under elevated CO2. Soil Biol. Biochem. 27:117–120.
Grace, J., and M. Rayment. 2000. Respiration in the balance.
Nature 404:819–820.
Griffiths, B. S., K. Ritz, R. D. Bardgett, R. Cook, S. Christensen, F. Ekelund, S. J. Sørensen, E. Bååth, J. Bloem, P.
C. de Ruiter, J. Dolfing, and B. Nicolardot. 2000. Ecosystem response of pasture soil communities to fumigationinduced microbial diversity reductions: an examination of
the biodiversity–ecosystem function relationship. Oikos
90:279–294.
May 2004
STOICHIOMETRIC ECOLOGY
Mäkipää, R., T. Karjalainen, A. Pussinen, and S. Kellomäki.
1999. Effects of climate change and nitrogen deposition
on the carbon sequestration of a forest ecosystem in the
boreal zone. Canadian Journal of Forest Research 29:1490–
1501.
Martin, J. 1990. Glacial–interglacial CO2 change: the iron
hypothesis. Paleoceanography 5:1–13.
Melillo, J. M., J. D. Aber, and J. F. Muratore. 1982. Nitrogen
and lignin control of hardwood leaf litter decomposition
dynamics. Ecology 63:621–626.
Moran, N., and W. D. Hamilton. 1980. Low nutritive quality
as defence against herbivores. Journal of Theoretical Biology 86:247–254.
Myklestad, S. 1977. Production of carbohydrates by marine
planktonic diatoms. II. Influence of the N/P ratio in the
growth of cellular and extracellular carbohydrates by Chaetocerus affinis var. willei (Gran) Hustedt and Skeletonema
costatum (Grev.) Cleve. Journal of Experimental Marine
Biology and Ecology 29:161–179.
Nagata, T. 2000. Production mechanisms of dissolved organic
matter. Pages 121–152 in D. L. Kirchman, editor. Microbial
ecology of the oceans. Wiley-Liss, New York, New York,
USA.
Norby, R. J., and M. F. Cotrufo. 1998. Global change—a
question of litter quality. Nature 396:17–18.
Norby, R. J., M. F. Cotrufo, P. Ineson, E. G. O’Neill, and J.
G. Canadell. 2001. Elevated CO2, litter chemistry, and decomposition: a synthesis. Oecologia 127:153–165.
Oren, R., D. S. Ellsworth, K. H. Johnsen, N. Phillips, B. E.
Ewers, C. Maier, K. V. R. Schafer, H. McCarthy, G. Hendrey, S. G. McNulty, and G. G. Katul. 2001. Soil fertility
limits carbon sequestration by forest ecosystems in a CO2enriched atmosphere. Nature 411:469–472.
Overdieck, D. 1993. Elevated CO2 and the mineral content
of herbaceous and woody plants. Vegetatio 104/105:403–
411.
Passow, U., R. F. Shipe, A. Murray, D. K. Pak, M. A. Brzezinski, and A. L. Alldredge. 2001. Origin of transparent
exopolymer particles (TEP) and their role in sedimentation
of particulate matter. Continental Shelf Research 21:327–
346.
Paul, E. A., and F. E. Clark. 1989. Soil microbiology and
biogeochemistry. Academic Press, San Diego, California,
USA.
Perkins, M. C., H. A. Woods, J. F. Harrison, and J. J. Elser.
2004. Dietary phosphorus affects the growth of larval Manduca Sexta. Archives of Insect Biochemistry and Physiology, in press.
Peterson, B. J., and J. M. Melillo. 1985. The potential storage
of carbon caused by eutrophication of the biosphere. Tellus
37B:117–127.
Rastetter, E. B., G. I. Ågren, and G. R. Shaver. 1997. Responses of N-limited ecosystems to increased CO2: a balanced-nutrition, coupled-element-cycles model. Ecological
Applications 7:444–460.
Rastetter, E. B., R. B. McKane, G. R. Shaver, and J. M.
Melillo. 1992. Changes in C storage by terrestrial ecosystems: how C-N interactions restrict responses to CO2 and
temperature. Water, Air, and Soil Pollution 64:327–344.
Rastetter, E. B., M. G. Ryan, G. R. Shaver, J. M. Melillo, K.
J. Nadelhoffer, J. E. Hobbie, and J. D. Aber. 1991. A general biogeochemical model describing the responses of the
C and N cycles in terrestrial ecosystems to changes in CO2,
climate, and N deposition. Tree Physiology 9:101–126.
Redfield, A. C., B. H. Ketchum, and F. A. Richards. 1963.
The influence of organisms on the composition of sea water.
Pages 26–77 in M. N. Hill, editor. Comparative and descriptive oceanography. Wiley, New York, New York,
USA.
Special Feature
Gutschick, V. P. 1981. Evolved strategies in nitrogen acquisition by plants. American Naturalist 118:607–637.
Hairston, N. G., Jr., and N. G. Hairston, Sr. 1993. Cause–
effect relationships in energy flow, trophic structure, and
interspecific interactions. American Naturalist 142:379–
411.
Hairston, N. G., F. E. Smith, and L. B. Slobodkin. 1960.
Community structure, population control, and competition.
American Naturalist 94:421–425.
Hastings, H. M., and M. Conrad. 1979. Length and evolutionary stability of food webs. Nature 282:838–839.
Heinze, C., E. Maier-Reimer, and K. Winn. 1991. Glacial
pCO2 reduction by the world ocean: experiments with the
Hamburg carbon cycle model. Paleoceanography 6:395–
430.
Hessen, D. O. 1992. Nutrient element limitation of zooplankton production. American Naturalist 140:799–814.
Hessen, D. O., B. Faafeng, and P. Brettum. 2003. Autotroph:
herbivore biomass ratios; carbon deficits judged from
plankton data. Hydrobiologia 491:167–175.
Hessen, D. O., P. Færøvig, and T. Andersen. 2002. Light,
nutrients, and P:C ratios in algae: grazer performance related to food quality and quantity. Ecology 83:1886–1898.
Higashi, M., T. Abe, and T. P. Burns. 1992. Carbon–nitrogen
balance and termite ecology. Proceedings of the Royal Society of London B 249:303–308.
Hutchinson, G. E. 1959. Homage to Santa Rosalia or Why
are there so many kinds of animals? American Naturalist
93:145–159.
Ingestad, T. 1979. Nitrogen stress in birch seedlings. II. N,
K, P, Ca, and Mg nutrition. Physiologia Plantarum 45:149–
157.
Ingestad, T., and G. I. Ågren. 1995. Plant nutrition and
growth—basic principles. Plant and Soil 169:15–20.
Ingestad, T., and A.-B. Lund. 1979. Nitrogen stress in birch
seedlings. I. Growth technique and growth. Physiologia
Plantarum 45:137–148.
Jaenike, J., and T. A. Markow. 2003. Comparative elemental
stoichiometry of ecologically diverse Drosophila. Functional Ecology 17:115–120.
Joffre, R., and G. I. Ågren. 2001. From plant to soil: litter
production and decomposition. Pages 83–99 in J. Roy, B.
Saugier, and H. A. Mooney, editors. Terrestrial global productivity. Academic Press, San Diego, California, USA.
Johnson, K. S., R. M. Gordon, and K. H. Coale. 1997. What
controls dissolved iron concentrations in the world ocean?
Marine Chemistry 57:137–161.
Kähler, P., and W. Koeve. 2001. Marine dissolved organic
matter: can its C:N ratio explain carbon overconsumption?
Deep-Sea Research 48:49–62.
Killingbeck, K. T. 1993. Nutrient resorption in desert shrubs.
Revista Chilena de Historia Natural 66:345–355.
Killingbeck, K. T. 1996. Nutrients in senescing leaves: key
to the search for potential resorption and resorption proficiency. Ecology 77:1716–1727.
Kirschbaum, M. U. F. 2000. Will changes in soil organic
carbon act as a positive or negative feedback on global
warming? Biogeochemistry 48:21–51.
Kooijman, S. A. L. M. 1998. The synthesizing unit as model
for the stoichiometric fusion and branching of metabolic
fluxes. Biophysical Chemistry 73:179–188.
Kooijman, S. A. L. M., T. Anderson, and B. W. Kooi. 2004.
Dynamic energy budget representations of stoichiometric
constraints on population dynamics. Ecology 85:1230–
1243.
Lampitt, R. S., W. R. Hillier, and P. G. Challenor. 1993. Seasonal and diel variations in the open ocean concentration
of marine snow aggregates. Nature 362:737–739.
Lindeman, R. I. 1942. The trophic dynamic aspect of Ecology. Ecology 23:399–418.
1191
Special Feature
1192
DAG O. HESSEN ET AL.
Schimel, D. S. 1995. Terrestrial ecosystems and the carbon
cycle. Global Change Biology 1:77–91.
Schimel, D. S., B. H. Braswell, E. A. Holland, R. McKeown,
D. S. Ojima, T. H. Painter, W. J. Parton, and A. R. Townsend. 1994. Climatic, edaphic and biotic controls over storage and turnover of carbon in soils. Global Biogeochemical
Cycles 8:279–293.
Schindler, D. W. 1977. Evolution of phosphorus limitation
in lakes. Science 195:260–262.
Schindler, D. W., and S. E. Bayley. 1993. The biosphere as
an increasing sink for atmospheric carbon—estimates from
increased nitrogen deposition. Global Biogeochemical Cycles 7:717–733.
Schlesinger, W. H. 1991. Biogeochemistry—an analysis of
global change. Academic Press, San Diego, California,
USA.
Søndergaard, M., P. J. le B. Williams, G. Cauwet, B. Riemann,
C. Robinson, S. Terzic, E. M. S. Woodward, and J. Worm.
2000. Net accumulation and flux of dissolved organic carbon and dissolved organic nitrogen in marine plankton
communities. Limnology and Oceanography 45:1097–
1111.
Sterner, R. W., and J. J. Elser. 2002. Ecological stoichiometry: the biology of elements from molecules to the biosphere. Princeton University Press, Princeton, New Jersey,
USA.
Sterner, R. W., and D. O. Hessen. 1994. Algal nutrient limitation and the nutrition of aquatic herbivores. Annual Review of Ecology and Systematics 25:1–29.
Stoddard, J. L. 1994. Long term changes in watershed retention of nitrogen. Its causes and aquatic consequences.
Pages 223–284 in L. A. Baker, editor. Environmental chemistry of lakes and reservoirs. Advances in Chemistry Series
number 237. Americal Chemical Society, Washington,
D.C., USA.
Swift, M. J., O. W. Heal, and J. M. Anderson. 1979. Decomposition in terrestrial ecosystems. Blackwell, Oxford, UK.
Teira, E., M. J. Pazó, P. Serret, and E. Fernández. 2001. Dissolved organic carbon production by microbial populations
in the Atlantic Ocean. Limnology and Oceanography 46:
1370–1377.
Tezuka, Y. 1990. Bacterial regeneration of ammonium and
phosphate as affected by the carbon:nitrogen:phosphorus
ratio of organic substrates. Microbial Ecology 19:227–238.
Thomas, H., V. Ittekkot, C. Osterroht, and B. Schneider. 1999.
Preferential recycling of nutrients—the ocean’s way to increase new production and to pass nutrient limitation? Limnology and Oceanography 44:1999–2004.
Tilman, D. 1988. Plant strategies and the dynamics and structure of plant communities. Princeton University Press,
Princeton, New Jersey, USA.
Townsend, A. R., B. H. Braswell, E. A. Holland, and J. E.
Penner. 1996. Spatial and temporal patterns in terrestrial
carbon storage due to deposition of fossil fuel nitrogen.
Ecological Applications 6:806–814.
Trier, T. M., and W. J. Mattson. 2003. Diet-induced thermogenesis in insects: a developing concept in nutritional
ecology. Environmental Entomology 32:1–8.
Ecology, Vol. 85, No. 5
Tyrrell, T. 1999. The relative influences of nitrogen and phosphorus on oceanic primary production. Nature 400:525–
531.
Tyrrell, T., and A. H. Taylor. 1996. A modelling study of
Emiliania huxleyi in the NE Atlantic. Journal of Marine
Systems 9:83–112.
Urabe, J., J. Togari, and J. J. Elser. 2003. Stoichiometric
impacts of increased carbon dioxide on a planktonic herbivore. Global Change Biology 9:818–825.
Vadstein, O. 2000. Heterotrophic, planktonic bacteria and
cycling of phosphorus—phosphorus requirements, competitive ability, and food web interactions. Advances in
Microbial Ecology 16:115–167.
van der Wal, R., N. Madan, S. van Lieshout, C. Dormann,
R. Langvatn, and S. D. Albon. 2000. Trading forage quality
or quantity: plant phenology and patch choice by Svalbard
reindeer. Oecologia 123:108–115.
van Veen, J. A., E. Liljeroth, L. J. A. Lekkerkerk, and S. C.
van de Geijn. 1991. Carbon fluxes in plant–soil systems
at elevated atmospheric CO2 levels. Ecological Applications 1:175–181.
Vitousek, P. M., J. D. Aber, R. W. Howarth, G. E. Likens, P.
A. Matson, D. W. Schindler, W. H. Schlesinger, and D. G.
Tilman. 1997. Human alterations of the global nitrogen
cycle: sources and consequences. Ecological Applications
7:737–750.
Vitousek, P. M., T. Fahey, D. W. Johnson, and M. J. Swift.
1988. Element interactions in forest ecosystems: succession, allometry and input–output budgets. Biogeochemistry
5:7–34.
Vitousek, P. M., and R. W. Howarth. 1991. Nitrogen limitation on land and in the sea: how can it occur? Biogeochemistry 13:87–115.
Vrede, T., D. Dobberfuhl, S. A. L. M. Kooijman, and J. J.
Elser. 2004. Fundamental connections among organism
C:N:P stoichiometry, macromolecular composition, and
growth. Ecology 85:1217–1229.
Vrede, T., J. Persson, and G. Aronsen. 2002. The influence
on food quality (P:C ratio) on RNA:DNA ratio and somatic
growth of Daphnia. Limnology and Oceanography 47:487–
494.
Wells, M. L., N. M. Price, and K. W. Bruland. 1995. Iron
chemistry in seawater and its relationship to phytoplankton:
a workshop report. Marine Chemistry 48:157–182.
White, T. C. R. 1993. The inadequate environment: nitrogen
and the abundance of animals. Springer-Verlag, New York,
New York, USA.
Williams, P. J. leB. 1995. Seasonal accumulation of carbonrich dissolved organic material, its scale in comparison with
changes in particulate material and the consequential effect
on net C/N assimilation ratios. Marine Chemistry 51:17–
29.
Woods, H. A., W. Makino, J. B. Cotner, S. Hobbie, J. F.
Harrison, K. Acharya, and J. J. Elser. 2003. Temperature
shifts cause systematic changes in chemical composition
of ectothermic organisms. Functional Ecology 17:237–245.