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New Phytol. (1997), 137, 179–203
Tansley Review No. 95
"&N natural abundance in soil–plant
systems
B PETER HO$ GBERG
Section of Soil Science, Department of Forest Ecology,
Swedish University of Agricultural Sciences, S-901 83 Umeac , Sweden
(Received 2 September 1996)
Summary 179
I. Introduction 180
II. Units, causes of isotope effects,
stoichiometry, modelling 181
III. N dynamics and variations in "&N
abundance in soil–plant systems 183
1. General considerations 183
2. Specific processes 184
(a) N mineralization 184
(b) Ammonia volatilization 184
(c) Nitrification 185
(d ) Denitrification 185
(e) Ion exchange, diffusion and mass
flow 185
( f ) Plant uptake of N 185
(g) N uptake by mycorrhizal fungi and
mycorrhizal plants 186
Equilibrium and kinetic isotope fractionations during incomplete reactions result in minute differences in the
ratio between the two stable N isotopes, "&N and "%N, in various N pools. In ecosystems such variations (usually
expressed in per mil [δ"&N]deviations from the standard atmospheric N#) depend on isotopic signatures of inputs
and outputs, the input–output balance, N transformations and their specific isotope effects, and compartmentation
of N within the system. Products along a sequence of reactions, e.g. the N mineralization–N uptake pathway,
should, if fractionation factors were equal for the different reactions, become progressively depleted. However,
fractionation factors vary. For example, because nitrification discriminates against "&N in the substrate more than
does N mineralization, NH%
+ can become isotopically heavier than the organic N from which it is derived.
Levels of isotopic enrichment depend dynamically on the stoichiometry of reactions, as well as on specific abiotic
and biotic conditions. Thus, the δ"&N of a specific N pool is not a constant, and δ"&N of a N compound added to
the system is not a conservative, unchanging tracer. This fact, together with analytical problems of measuring δ"&N
in small and dynamic pools of N in the soil–plant system, and the complexity of the N cycle itself (for instance
the abundance of reversible reactions), limit the possibilities of making inferences based on observations of "&N
abundance in one or a few pools of N in a system. Nevertheless, measurements of δ"&N might offer the advantage
of giving insights into the N cycle without disturbing the system by adding "&N tracer.
Such attempts require, however, that the complex factors affecting δ"&N in plants be taken into account, viz. (i)
the source(s) of N (soil, precipitation, NOx, NH
$, N
#-fixation), (ii) the depth(s) in soil from which N is taken up,
(iii) the form(s) of soil-N used (organic N, NH%
+, NO$
−), (iv) influences of mycorrhizal symbioses and
fractionations during and after N uptake by plants, and (v) interactions between these factors and plant phenology.
Because of this complexity, data on δ"&N can only be used alone when certain requirements are met, e.g. when a
clearly discrete N source in terms of amount and isotopic signature is studied. For example, it is recommended
that N in non-N#-fixing species should differ more than 5^ from N derived by N
#-fixation, and that several non-
N#-fixing references are used, when data on δ"&N are used to estimate N
#-fixation in poorly described ecosystems.
(h) N#-fixation 187
(i) N metabolism in plants 187
( j) The role of animals 189
IV. Applications 189
1. Estimation of contributions of different
soil N and other non-N#
sources to
plant N uptake 189
2. Estimation of N#-fixation by the "&N
natural abundance method 191
3. Interpretation of δ"&N profiles in soils
(with comments on horizontal spatial
variability) 193
4. Assessment of N balances of ecosystems 195
V. Conclusions and suggestions for future
research 197
Acknowledgements 198
References 198
180 P. HoX gberg
As well as giving information on N source effects, δ"&N can give insights into N cycle rates. For example, high
levels of N deposition onto previously N-limited systems leads to increased nitrification, which produces "&N-
enriched NH%
+ and "&N-depleted NO$
−. As many forest plants prefer NH%
+ they become enriched in "&N in such
circumstances. This change in plant δ"&N will subsequently also occur in the soil surface horizon after litter-fall,
and might be a useful indicator of N saturation, especially since there is usually an increase in δ"&N with depth
in soils of N-limited forests.
Generally, interpretation of "&N measurements requires additional independent data and modelling, and
benefits from a controlled experimental setting. Modelling will be greatly assisted by the development of methods
to measure the δ"&N of small dynamic pools of N in soils. Direct comparisons with parallel low tracer level "&N
studies will be necessary to further develop the interpretation of variations in δ"&N in soil–plant systems. Another
promising approach is to study ratios of "&N: "%N together with other pairs of stable isotopes, e.g. "$C: "#C or
")O: "'O, in the same ion or molecules. This approach can help to tackle the challenge of distinguishing isotopic
source effects from fractionations within the system studied.
Key words: "&N abundance, nitrogen, plants, soils.
.
The ratio between the two stable isotopes of
nitrogen, "&N and "%N, varies in the biosphere as a
result of isotope fractionation in physical, chemical
and biological processes. Atmospheric N#, which has
a "&N abundance of 0±3663 atom%, is the accepted
standard in this context (Junk & Svec, 1958;
Mariotti, 1983). Variations in the ratio of other N
pools usually fall within the narrow interval of
®0±0040–0±0060 atom% from this standard (e.g.
Moore, 1974; Le! tolle, 1980; Macko & Ostrom,
1994; Nadelhoffer & Fry, 1994), and some variability
of interest lies within less than one tenth of this
amplitude. Because of the focus on variability
associated with the third and fourth decimal places,
δ units, which represent parts per thousand (^) of
the abundance of the "&N:"%N ratio in atmospheric
N#, are far more commonly used than atom% (see
Section II).
Early reports on variations in "&N abundance in
nature were made by, amongst others,, Schoen-
heimer & Rittenberg (1929), Hoering (1955), Parwel,
Ryhage & Wickman (1957), Wellman, Cook &
Krouse (1968) and Delwiche & Steyn (1970). It was
realized that regular variations in stable N ratios
could potentially provide useful, sometimes unique,
information about sources of N used by plants, and
fluxes of N in ecosystems. Studies during the 1970s
and 1980s largely focused on tracing the fate of
fertilizer N in intensive agriculture (e.g. Kohl,
Shearer & Commoner, 1971; Meints, Boone &
Kurtz, 1975a), and on estimating the fractional
contribution of N#-fixation to the N of N
#-fixing
plants (e.g. Amarger, Mariotti & Mariotti, 1977;
Delwiche et al., 1979; Shearer & Kohl, 1986).
Contemporary critics cautioned, in the former case,
that internal isotope fractionations within eco-
systems would preclude meaningful interpretation of
source effects, i.e. they pointed out that the isotopic
signature is not always a good conservative tracer
(e.g. Hauck et al., 1972). A much discussed problem
was that of analysis of minute differences in isotopic
composition of small pools of N, and another concern
was over the large variability of soil δ"&N. These
problems remain significant, and are consequently a
recurrent theme in this text. However, further
evidence of consistent variations in δ"&N accumulated
and provided arguments for continued work as
earlier reviewed by Le! tolle (1980), Hu$ bner (1986),
Shearer & Kohl (1986, 1993), Handley & Raven
(1992), Macko & Ostrom (1994), Nadelhoffer & Fry
(1994), and Handley & Scrimgeour (1997).
The recent development of analysing isotopic
pairs of more than one element in a molecule or ion
can help to distinguish isotopic source effects from
effects of isotope fractionation, for instance by
measuring both δ"&N and δ")O in NO$
− (Amberger &
Schmidt, 1987; Aravena, Evans & Cherry, 1993).
Using this technique Durka et al. (1994) attempted
to distinguish between the contributions to ground-
water of NO$
− from rainwater and from nitrification
in forest soils (see IV.4). Similar technical ap-
proaches, coupled to modelling of isotope fraction-
ations, will most probably dominate coming ad-
vances within this field.
Some decades ago only a few laboratories, notably
in Australia, France, Germany, Japan and North
America, had isotope ratio mass spectrometers
(IRMS) of the calibre needed in studies of "&N
natural abundance. Such instruments were largely
based on the original design by Nier (1947). They
were capable of very precise measurements, but the
necessary preparation of samples by Kjeldahl wet
digestion, and subsequent analysis of N#
released
after reaction with hypobromite in Rittenberg tubes
was tedious and slow (Haystead, 1983; Robinson &
Smith, 1991). The maximum throughput possible
was c. 25 samples d−". Many modern instruments are
based on the principle of combusting dry samples in
O#according to the Dumas method, and introducing
the products into a continuous flow of a carrier gas
admitted to the mass spectrometer (CF-IRMS, e.g.
Barrie & Lemley, 1989). This allows analysis of
about 100 samples d−". Losing the potentially higher
precision of traditional dual-inlet IRMS is perhaps
"&N natural abundance in soil–plant systems 181
less important, because the biological variation
between, e.g., individuals of a plant species at a site
is usually larger than the analytical error. Moreover,
if the N concentration of samples can be measured or
estimated before isotopic analysis, the analyst can
minimize variations in N content between samples
and thereby acquire a high precision with CF-IRMS
(³0±1^ of repeated samples, Handley et al.,
1993). In our laboratory, we frequently obtain a
precision of ³0±2^ by using CF-IRMS and
adjusting the sample size, and thus its N content,
based purely on rough a priori estimates of %N.
CF-IRMS systems are rapidly becoming wide-
spread, and commercial analytical services are now
available, which make "&N natural abundance studies
an option for many more terrestrial ecologists. This
review, which attempts to guide that group of people,
as well as to provide substance for further discussions
between those more specialized within the field,
focuses on the use of variations in "&N abundance in
studies of N in soil–plant systems. Isotope fraction-
ations within plants, although of relevance, will not
be treated in detail as they were recently covered by
Raven (1987), Handley & Raven (1992) and
Yoneyama (1995).
. , ,
,
As stated above, natural "&N abundance is commonly
expressed in δ units, which denote parts per thousand
deviations, ^, from the ratio "&N:"%N in atmospheric
N#, which is 0±0036765 and corresponds to
0±3663 atom% "&N (Junk & Svec, 1958; Mariotti,
1983):
δ"&N(^)¯ ((Rsample
}Rstandard
)®1)¬1000, (1)
where R denotes the ratio "&N:"%N which, however,
is derived from the ratio between masses 29 and 28
of charged N#
molecules reaching the Faraday cups
at the end of the flight tube in the mass spectrometer.
The "&N abundance A in atom% is:
A¯100}(1}(1((δsample
}1000)1)¬Rreference
)). (2)
Isotope effects are only seen when reactions are
incomplete, i.e. when not all N atoms in a substrate
go into the product of a reaction (Fig. 1). Within a
specific system there can be a number of pools or
compartments of N-containing molecules or ions
with distinctly different "&N abundances. The pro-
portions between these N pools might change, as
well as their individual δ"&N, but the overall weighted
average δ"&N of the system remains the same, unless
N is added or lost, and the N in those fluxes has an
isotopic signature different from the mean value of
the system.
Variations in stable isotope ratios are the result of
equilibrium and kinetic isotope effects. A larger
activation energy is required to dissociate an iso-
Substrate
Instantaneousproduct
Cumulativeproduct
0 50 100
Substrate consumption (%)
å
d15N
(‰
)
Figure 1. Relative changes in δ"&N of components during
a complete reaction in a closed system (redrawn from
Mariotti et al., 1981). In an open reaction system, where
the supply of substrate is infinite, the instantaneously
formed, as well as the cumulative, product would fall on
the same horizontal parallel line below the substrate-line,
and the difference between the substrate- and the product-
lines would be ε. Many reactions in nature would fall
between these two extreme examples.
topically heavy chemical species than a light one.
Hence, an isotopically light atom or ion will be
bonded less strongly at equilibrium (Bigeleisen,
1965). Equilibrium isotope effects in reactions of the
type A%B may be described by α :
α¯ δA}δ
B. (3)
As an example we may use the equilibrium between
NH$
and NH%
+ in aqueous solution:
"%NH$"&NH
%
+ % "&NH$"%NH
%
+. (4)
In this case NH%
+ is more enriched with "&N than
NH$at equilibrium, i.e. α"1 (e.g. Kirschenbaum et
al., 1947). This reaction commonly preceeds another
equilibrium reaction with a well-documented strong
isotope fractionation, i.e. ammonia volatilization (see
III.2.(b)). Ion exchange is another equilibrium
reaction which might involve isotope fractionation
(see III.2.(e)).
The other type of isotope fractionation is kinetic
fractionation, which occurs because heavier mole-
cules or ions react more slowly than isotopically
lighter analogues (Fig. 1). Note that Figure 1
182 P. HoX gberg
Table 1. Fractionation factors, α, for various processes
in the N cycle
Process Fractionation
factor
N mineralization E1±000
(org N!NH%
+)
NH%
+ %NH$
in solution 1±020–1±027*
NH$
volatilization 1±029
Diffusion of NH%
+, NH$, E1±000
NO$
− in solution
Nitrification 1±015–1±035
Dentrification 1±000–1±033
N assimilation 1±000–1±020†N
#-fixation 0±998–1±002
Metabolic steps in plants 0±980–1±020
*Equilibrium fractionation factor, other examples rep-
resent kinetic fractionations.
†Values!1±002 are probably more appropriate in most
natural situations.
As in Handley & Raven (1992) α (eqn (5)), if not given
directly in the original reference or compilation of
references, has been calculated as α¯ ((δ"&Nsubstrate
-
δ"&Nproduct
)}1000)1. Data largely from compilations by
Hu$ bner (1986), Shearer & Kohl (1986) and Handley &
Raven (1992). The table gives an overview and omits some
exceptional values discussed in the text.
describes a closed reaction, which occurs when the
supply of substrate is limited. The difference
between this and an open system with unlimited
supply of the substrate is described in the text to
Figure 1. Kinetic isotope fractionations are often
also described by α, which relates, in this case, to the
ratio between the rates, kLand k
H, of a process for the
light and heavy isotopes, respectively:
α¯kL}k
H. (5)
Some authors, however, use β instead of α and define
α as the inverse of eqn (5) (e.g. Mariotti et al., 1981).
Hence, it is important to make clear definitions. In a
simple case, a unidirectional reaction in which
substrate is not limiting, α is constant. Fractionation
factors here range from ®0±98 to 1±06 (Table 1;
III.2(a)–( j)). When substrate is in limited supply,
the isotopic composition of the instantaneously
formed product, as well as that of the cumulative
product, varies during the reaction (Fig. 1). It is thus
difficult, during a reaction, to isolate the instan-
taneously formed product and to measure its δ"&N.
It is also often of interest to focus on ∆δ, more
commonly referred to as ∆ (or ε), i.e. the dis-
crimination, which is the difference between δ of the
substrate (δs) and δ of the product (δ
p) :
∆s/p
¯ ((δ®δp)}(1δ
p}1000)) (6)
Generally, the denominator is approximated by 1
here, and hence:
∆s/p
¯ δs®δ
p. (7)
A convenient approximation of the fractionation
factor is α¯ (∆}1000)1 (Handley & Raven, 1992).
Isotopic fractionation can thus also be described
by the enrichment ε, which describes the enrichment
of the product relative to that of the substrate, and
which is also expressed per mil (^), and may be
positive or negative. It is another term for dis-
crimination (∆) :
ε¯ (α®1)¬1000. (8)
Note that a fractionation factor, α, of 1±020 leads to
an enrichment (actually a depletion), ε, of the product
of ®20^ relative to the substrate.
Using a modification of the classical Rayleigh
equation, it is possible to calculate ε for a reaction
with a finite supply of substrate if the fraction of
unreacted substrate, f, and δsand
s!, i.e. the δ of the
substrate remaining and at time zero, respectively,
are known (Mariotti et al., 1981):
ε¯1000¬loge(((1}1000)¬δ
s1)}
((1}1000)¬δs!1))}log
ef. (9)
This equation, or derivations of it, e.g.
δs¯ δ
s!ε log
ef, (10)
have been elegantly used in analyses of the decrease
in NO$
− concentration caused by denitrification in
aquifers (e.g. Mariotti, Landreau & Simon, 1988). In
a soil system, however, consisting of various solids in
addition to an aqueous solution, organisms and a gas
phase, it is difficult to confine the actual substrate
(see III), and hence to apply eqn (9) or eqn (10).
Many different values of the fractionation factors
α or β are found in the literature (e.g. Hu$ bner, 1986;
Shearer & Kohl, 1986; Handley & Raven, 1992;
Table 1), but one should not adopt, by default,
values from other studies without proper consider-
ation. Firstly, related organisms might differ signifi-
cantly. For example, there is sometimes a slight
discrimination against "&N during N#-fixation, and
this varies depending on the strain of bacterium in
legume–Rhizobium symbioses (see III 2(h)). Sec-
ondly, effects of abiotic factors, or interactions
between these and biotic factors, are not easily
predicted, e.g. in different soil types, in different
biomes etc., where rates and processes of N trans-
formations are very different. Even at a single site,
conditions can vary considerably spatially and tem-
porally. Thirdly, bi- and multidirectional reactions,
as well as reversible reactions, which are the rule
rather than the exception in soils, limit the ap-
plicability of simple mathematical approaches
(Peterson & Fry, 1987), and more advanced mod-
elling will certainly be necessary to deal with complex
multicomponent reaction systems.
In some experimental settings it is, however, often
valuable to apply simple mass–balance equations,
which allow the calculation of, for example, the δ"&N
of a N source without actually measuring it. Consider
a case, in which the isotopic signature and amount of
seed N (δ"&Nseed
and Nseed
), as well as the isotopic
"&N natural abundance in soil–plant systems 183
composition and amount of total N in the plant
(δ"&Nplant
and Nplant
), are known. The δ"&N of the N
source (δ"&Navail
) used in addition to seed N by the
plant can then be calculated:
δ"&Navail
¯ ((δ"&Nplant
¬Nplant
)®(δ"&Nseed
¬Nseed
))}
(Nplant
®Nseed
). (11)
When mixing-models of this type are used, the
estimated δ"&N of a small N source (Navail
) is likely to
be particularly imprecise.
A number of models have been developed to cope
with the greater challenge of describing variations in
δ"&N in dynamic interactions in soil–plant systems
(Focht, 1973; Shearer et al., 1974; Herman &
Rundel, 1989; Winkler & Gebauer, 1993; Garten &
van Miegroet, 1994; van Damm & van Breemen,
1994). These compartment-flow models resemble
conventional N cycle models, but have the added
feature of tallying "&N:"%N ratios. They differ from
"&N tracer models in that fluxes are associated with
certain fractionation factors, α or β. In series of
reactions, the fractionation during the rate-limiting
step determines the net overall fractionation for the
entire reaction sequence. If α were the same for all
reactions, products of sequential reactions should
become progressively depleted, but this is not the
case. For example, NH%
+ can become heavier than
the organic N it is derived from in a nitrifying system
because of the large fractionation during nitrification
(Shearer et al. 1974; Garten & van Miegroet 1994;
see III. 2(c)). Refinement of δ"&N models will
require interactive comparisons with "&N tracer
studies and models developed from them. In ad-
dition, their validation would be greatly assisted by
technical developments allowing reliable and precise
measurements of δ"&N values of small but rapidly
turning-over pools of N in soils.
Those interested further in isotope stoichiometry
and the chemical and physical basis of isotope
fractionation are referred to texts by Tong &
Yankwich (1957), Craig, Miller & Wasserberg
(1964), Bigeleisen (1965), Melander & Saunders
(1980), Mariotti et al. (1981), Shearer & Kohl (1993)
and Hoefs (1997).
. "&
–
1. General considerations
Most of the N in most soils is bound in forms not
immediately available to plants (Jansson, 1958;
Binkley & Hart, 1989). Hence, the δ"&N of soil total-
N is, in general, not a good approximation of the
δ"&N of N available to plants. At most a few % of soil
total N becomes available during a year. Recent "&N
tracer work, based on calculations of pool dilutions
and high temporal resolution (Davidson et al., 1991),
suggests a turnover of inorganic N pools within one
or a few days in both temperate grasslands and acid
forest soils (Davidson, Stark & Firestone, 1990;
Davidson, Hart & Firestone, 1992; Hart et al.,
1994). Turnover times in agricultural soils should be
comparable or even shorter. Conventional soil
incubations suggest a considerably slower rate of net
N mineralization and nitrification than this, es-
pecially in forest soils (Davidson et al., 1992). This is
because in such incubation studies the focus is on net
mineralization and net nitrification, and excess N is
only produced when micro-organisms run out of
available C (Hart et al., 1994), whereas "&N pool
dilution studies also recognize microbial immobi-
lization. There is now evidence that soil inorganic N
pools turn over within a few days. Free amino acids
and proteins might possibly turn over at similar rates
(Kielland, 1995), whereas microbial N and other
labile N pools turn over within weeks or months.
Using a "&N tracer approach Bjarnason (1988) found,
in an agricultural soil, that remineralization of added
NH%
+ took place after 2 wk, and NO$
− from labelled
N was produced in 4 wk. Long-lived exceptions are
the microbial N compounds which are precursors of
stable organic N. This recalcitrant pool, which is
formed by biological as well as chemical immobi-
lization of N, forms the major portion of soil total-N
and might, like the C it is bound to, have turnover
times of hundreds of years (Paul & Clark, 1989).
The complex dynamics of soil N make detailed
studies of natural abundance of "&N of soil N pools
difficult. Attempts to isolate the very small biologi-
cally active N pools in soils are likely to disturb the
studied system. Hence, the measured "&N of these
pools might represent artefacts rather than the true
isotopic signature of the N available to plant roots.
For example, Lindau & Spalding (1984) reported
that the δ of KCl-extractable NO$
− was apparently
affected by the sample:extractant ratio; a decrease of
this ratio from 1:1 to 1:10 increased the measured
δ"&N by 6±2^. Variations in substrate supply, abiotic
conditions and composition of organism assemblages
and their demand for N all dynamically affect and
change the isotopic signatures of each of the various
N species (see III.2( f )).
The classical view has been that plants principally
use inorganic N, although capacity to use simpler
organic N sources might not be uncommon. How-
ever, as early as the 1880s Frank (1885) suggested
that ectomycorrhizal (ECM) species should have
access to organic N in forest soils. His suggestion has
now been confirmed (e.g. Abuzinadah, Finlay &
Read, 1986), and evidently plants with ericoid
mycorrhizas also have the capacity to use proteins,
amino acids and even chitin (Bajwa, Abuarghub &
Read, 1985; Leake & Read, 1990). Furthermore, it
was recently reported that a non-mycorrhizal (NON-
MYC) sedge typical of high-latitude systems, Erio-
phorum vaginatum, could take up "%C-labelled meth-
ylamine, glycine, glutamate and aspartate (Chapin,
184 P. HoX gberg
d15N
(‰
)
10
0
–10
Time (months)
NH3
applied
Figure 2. Sequential changes in δ"&N of NH%
+ (E) and
NO$
− (D) after addition of anhydrous ammonia to an
agricultural field (redrawn from Feigin et al., 1974).
Moilanen & Kielland, 1993), and it might well be
that the importance of organic N sources has wrongly
been overlooked in many systems in which they are,
in fact, used by plants.
In theory, different sources of N might have
different δ"&N, but, as discussed above, these signa-
tures are far from constant and discrete. For
example, measured δ"&N of extractable NO$
−NH%
+
in a fallowed soil can decrease"10^ within half a
year (Turner et al., 1987; cf. also Fig. 2). Atmo-
spheric N#
is an exception, but, again, as described
above (see also III.2(h)), the condition of the plant
host and the identity of the bacterial symbiont can
also influence the isotopic signature of N derived
from this source.
There is no direct way of assessing the fractional
contributions of organic N, NH%
+ and NO$
− to plant
N uptake in the field. However, the presence of
NO$
− in plant tissue (Hesselman, 1917) or nitrate
reductase activity (NRA) above constitutive levels
("0±2–0±3 µmol NO#
− g−" (fresh matter) h−" in in
vivo tests (Lee & Stewart, 1978)) imply that NO$
− is,
or has recently been, taken up from the soil (unless
there is enough NOxin the air to make it a significant
N source, cf. Wellburn, 1990). Higher in vivo NRA
in shoots, e.g. "1 µmol NO#
− g−" (fresh matter) h−",
provides a strong indication that NO$
− is an
important N source, since NH%
+ in the growth
medium might suppress uptake of NO$
− (Scheromm
& Plassard, 1988; Lee & Drew, 1989; Marschner,
Ha$ ussling & George, 1991). Moreover, competition
between roots and microbes has an important
influence on plant N source availability in soils (e.g.
Riha, Campbell & Wolfe, 1986; Zak et al., 1990),
and will hence also affect the δ"&N of N taken up by
different organisms.
Horizontal and vertical variability in δ"&N adds to
the complexity (see IV.3), especially in natural
ecosystems, where soils are not tilled, unlike soil in
most agricultural fields, and where the mixing of
soils by animal activity is sometimes insignificant.
Animal activity can, on the other hand, also con-
tribute to heterogeneity. For example, Ho$ gberg &
Alexander (1995) found that Combretum molle, a tree
species found exclusively in association with termite
mounds, was "3^ enriched relative to other non-
N#-fixing species in African miombo woodland.
Trees, with their very extensive root systems, pose
particular problems, because they scavenge a large
volume of soil with many different microsites. They
also reflect a formidable mixture of N source effects
due to the temporal variability in δ"&N of available N
sources.
In conclusion, the δ"&N of soil total-N is domi-
nated by the isotopic signature of stable N, which is
not likely to change over decades (Johannisson &
Ho$ gberg, 1994). By contrast, biologically active
pools might change quite dramatically over short
time periods, but are difficult to analyse more
directly. Specific, active N pools do not have discrete
constant δ"&N. Plants are integrators of δ"&N of
available N sources, but δ"&N in foliage might be
affected by the signature of N stored in the plants in
addition to the signature of recent soil-derived N.
Possibly, fine roots might give the most reliable
information on the δ"&N of available N in the soil (see
III.2( f ) and III.2(i)).
2. Specific processes
(a) N mineralization. There is very little evidence
that fractionation of N isotopes during mineral-
ization of N from larger molecules in soils should be
of significance, with the exception of the theoretical
derivation made by Focht (1973), who calculated a
theoretical fractionation of 1±0046 for tryptophan,
the heaviest natural amino acid. Isotope effects can
affect a minor functional group of a heavy molecule
(cf. Schowen & Schowen, 1995), but cannot be
detected if the isotopic signature of the whole
molecule is measured. Furthermore, the small
differences (!2^) in δ"&N between soil total-N and
root N in surface horizons in systems which are
thought to have little nitrification (Nadelhoffer &
Fry, 1994; Ho$ gberg et al., 1996), also suggest that N
isotope fractionation during N mineralization is
small. More conclusive evidence is required before
firm conclusions can be drawn. As a note, the large
fractionation for N mineralization mentioned by
Le! tolle (1980) refers to the sequence organic N!NH
%
+ !NO$
−, and is not confined to the frac-
tionation during the first step in that sequence.
(b) Ammonia volatilization. Volatilization of NH$
involves several steps in which isotopic fractionation
can occur: the equilibrium NH%
+ !NH$in solution
(cf. eqn (4)), diffusion of NH$
to the site of
volatilization, volatilization of NH$, and diffusion of
NH$away from the site of volatilization, unless it is
"&N natural abundance in soil–plant systems 185
removed by turbulent air flow (in which case the
transport does not involve any fractionation). The
compounded effect of these processes on the net
fractionation can be large, as the equilibrium
fractionation and the kinetic fractionation associated
with the volatilization per se have each been reported
to have α values "1±02 (Table 1). The net
fractionation, as measured by absorption of NH$by
an ‘infinite’ sink, e.g. H#SO
%-acidified paper, can
yield NH%
+ which is c. ®40^ relative to the
substrate (e.g. Handley et al., 1996). However,
figures vary considerably according to reaction
stoichiometry, e.g. depending on which step in the
sequence is rate-limiting, and, therefore, on pH of
the substrate and other factors.
It should also be borne in mind that in situations
where the substrate supply is limited, as for example
when N-rich manure is applied to a field, the δ"&N of
the NH$
volatilized changes with time (cf. Fig. 1).
Where volatilization of NH$is a significant process,
it will leave the remaining N enriched. Hence, the
δ"&N of animal manure, which has lost NH$
and
might have a δ"&N"10^ or even "20^, can
sometimes be used to trace the fate of manure N in
the soil–plant system (Selles & Karamanos, 1986;
Kerley & Jarvis, 1996), but it also adds to the
heterogeneity in surface-soil δ"&N of pastures (Steele
& Daniel, 1978).
During senescence, plants and fungi, (in particular
those with high N concentrations, or more precisely,
those with high concentrations of soluble proteins
and free amino acids) might be a significant source of
NH$
(e.g. Wetselaar & Farquhar, 1980; Ingelo$ g &
Nohrstedt, 1993, see III.2.(i)). Ammonia volatil-
ization from fire can also be significant (Raison,
1979).
(c) Nitrification. Nitrification has been associated
with fairly large isotope effects (α) ranging from
1±015 to 1±036 in studies of pure cultures of
Nitrosomonas europaea carrying out the first step of
the reaction: NH%
+ !NO#
− (e.g. Delwiche & Steyn,
1970; Mariotti et al., 1981; Yoshida, 1988). Frac-
tionation during nitrification by organisms other
than Nitrosomonas, e.g. Nitrospira or heterotrophic
organisms, has not been studied. The second step of
nitrification, NO#
− !NO$
−, is not rate-limiting, and
should not, therefore, lead to a further fractionation.
In soils, complete nitrification (NH%
+ !NO#
− !NO
$
−) has been estimated to have a fractionation of
1±012–1±029 (Shearer & Kohl, 1986). A particularly
instructive example of the isotope effect of nitrifi-
cation was provided by Feigin et al. (1974), who
described the development of a difference ofE20^between NH
%
+ and NO$
− after addition of anhydrous
NH$
to an agricultural field (Fig. 2).
(d ) Denitrification. Fractionation during denitrifi-
cation has been found to be highly variable, with
fractionation factors of 1±000–1±033 (Wellman et al.,
1968; Delwiche & Steyn, 1970; Mariotti et al., 1981;
Bryan et al., 1983; Yoshida et al., 1989). Possible
explanations of this variability have invoked dif-
ferences in concentrations of electron donors and
acceptors, and variations in temperature (Shearer &
Kohl, 1986; Kohl & Shearer, 1995). Another
potential cause of variations in soils could be the
dispersion effect on isotope fractionation during
denitrification (Kawanashi et al., 1993). Incidentally,
the fractionation factor of 1±060, reported by Yoshida
(1988), during production of N#O by nitrification is
remarkably high.
Denitrification is important in aquatic systems and
wet terrestrial systems, but is thought to be less
significant in most well-drained soils (e.g. Tiedje et
al., 1982; Goodroad & Keeney, 1984; Robertson &
Tiedje, 1984). It occurs along with nitrification in
highly localized environments in the landscape (e.g.
Groffman et al., 1993) and is also localized in
extremely anaerobic micro-environments in the soil
(Parkin, 1987). However, Lloyd (1993) has recently
stated that denitrification is more widespread in
aerobic environments than previously thought, al-
though the quantitative importance of denitrification
under aerobic conditions remains to be assessed. It is
clear from the above that more information is
required about factors determining fractionation
during denitrification, and the approach of using
determinations of both δ")O and δ"&N in reaction
components provides new possibilities in this di-
rection (Shearer & Kohl, 1988).
(e) Ion exchange, diffusion and mass flow. Several
studies report a small fractionation (!1±005) during
ion exchange (Hu$ bner, 1986; Shearer & Kohl, 1986).
When a solution of NH%Cl, for instance, was
exchanged with cations on a clay, the immediate
reaction (0–2 h) was an increase in δ"&N of the
solution, followed by a decline to ®1±5^ of the
initial value (Karamanos & Rennie, 1978). However,
such equilibrium fractionations are variable and
dynamic, and thus differ depending on the ex-
perimental set-up. Although fractionations during
ion exchange have some potential to affect soil profile
development of δ"&N profiles in soil (Nadelhoffer &
Fry, 1994), they are likely to be small compared with
some kinetic biological fractionations (Table 1). The
same applies to diffusion of inorganic N in soil
solution. There is no isotopic fractionation during
mass flow.
( f ) Plant uptake of N. Mariotti et al. (1980a)
reported a small discrimination against "&N during
uptake of NO$
− by 38 species of plants (∆¯®0±25³0±10 (xa³1), range ®2±2 to 0±6^) in a
laboratory study, and suggested that the fraction-
ation was affected by the concentration of NO$
− in
the medium. Kohl & Shearer (1980) observed a
186 P. HoX gberg
larger discrimination against "&N in non-nodulating
soybeans, ryegrass and marigold (∆¯®3±2, ®4±4and ®4±0^, respectively), but they used a higher
concentration of NO$
− in the medium (7±5 compared
with 5 m). In another study, of pearl millet
(Pennisetum spp.), Mariotti et al. (1982) observed no
fractionation when plants were grown in 0±5 m
NO$
−, and a fractionation of ®3^ when the plants
were supplied with 12 m NO$
− in the rooting
medium. They suggested that NO$
− reduction was
the major step of fractionation. Yoneyama & Kaneko
(1989) reported there was no fractionation in a study
of Brassica campestris (but it is unclear whether they
subtracted the isotopic source effect of seed N or
not; cf. eqn (11)). In contrast to the above,
Bergersen, Peoples & Turner (1988) reported a
discrimination against "&N initially (up to 20 d),
followed by a discrimination against "%N (days
27–55), viz. an enrichment of "&N of 2^, in non-
N#-fixing soybeans grown in 5 m nitrate. They did
not use a solution culture system as used in the other
studies, and cautiously noted that exchange processes
on the sand–vermiculite substrate could have af-
fected their results.
As regards NH%
+, marine micro-organisms were
shown to discriminate less against "&N during uptake
of NH%
+ at high rates than at lower rates of uptake
(Hoch, Fogel & Kirchman, 1994). They reported a
variation in ∆ of between ®5 and ®20^ based on
measurements of the concentration and isotopic
composition of NH%
+ remaining in solution (and on
the assumption that NH%
+ was not remobilized by
the organisms). Yoneyama et al. (1991) found a
discrimination of c. ®4±4^ in two varieties of rice
given 1±4 m ammonia (NH$NH
%
+), whereas the
discrimination was around ®12^ when the con-
centration of ammonia was 7±1 m. Evans et al.
(1996) found no discrimination during uptake of
NO$
− or NH%
+ at the low concentration of 50 µ.
The capacity for uptake of NH%
+ is considerably
higher in N-deficient plants than in those given an
ample supply of N, which forms the basis for a root
bioassay to determine the N status of plants (e.g.
Jones, Quarmby & Harrison, 1991). The capacity for
uptake of NO$
− is also related to supply, and first
increases at low supply rates (phase I), thereafter
decreases (phase II), and finally (phase III) attains a
value at non-limiting supply (Larsson, 1994). These
phases are under regulation of influx, and there
might be an increased density of carriers in phase I,
which is metabolite down-regulated in phases II and
III.
There is an interaction between plant demand for
N, and competition between plants and micro-
organisms. In N-limited systems, where there are
constraints on autotrophic nitrification other than
substrate supply, e.g. low soil pH, NH%
+ should be
the most important inorganic N species available to
plants. Concentrations of inorganic N"1 m are
rare in soil solutions in most natural systems (e.g.
Evans et al., 1996); elevated concentrations occur
locally, or at most in very short transient flushes, e.g.
in the spring in temperate climates (e.g. Zak et al.,
1990; Groffman et al., 1993) and after the first rains
in tropical seasonal climates (Birch, 1958). Hence, in
most natural N-limited systems, uptake of N must
be very efficient, resulting in virtually all available
inorganic N being taken up, and, therefore, in no or
negligible fractionation (Nadelhoffer & Fry, 1994;
Evans et al., 1996).
The data from laboratory experiments show,
nevertheless, that discrimination against "&N during
uptake of inorganic N can occur, especially when the
concentration of N in the medium is high in relation
to plant demand. Further investigations of this
phenomenon should, therefore, better describe the
relationships between isotope fractionation and the
balance between N supply and demand, by using, for
example, the growth technique proposed by Ingestad
(1982). It is possible that some of the intra- and
interspecific variability in fractionation discussed
above did not reflect intrinsic differences in frac-
tionation during uptake, but rather occurred as a
result of differences in the balance between N supply
and demand for N under specific growth conditions.
There is also a need for data on the possible
discrimination against "&N during uptake of organic
N.
(g) N uptake by mycorrhizal fungi and mycorrhizal
plants. The studies discussed above were conducted
on NON-MYC plants, a condition thought to be
uncommon for many species under natural con-
ditions (Newman & Reddell, 1987; Smith & Read,
1997). It is important, therefore, to establish the
degree to which mycorrhizal fungi alter the δ"&N of
N taken up from the soil. Bardin, Domenach &
Chalamet (1977) reported from a laboratory study
that δ"&N in ECM Pinus halepensis was ®2^ relative
to that in NON-MYC plants, whereas Handley et al.
(1993) found no difference between ECM and NON-
MYC Eucalyptus globulus but found a significant
difference of 0±7^ between arbuscular mycorrhizal
(AM) and NON-MYC Ricinus communis. It is
possible, however, that the isotope effects reported
by Bardin et al. (1977) and Handley et al. (1993)
reflected differences in dilution of seed N rather than
differences in isotope fractionation during uptake of
N (Ho$ gberg et al., 1994). Further detailed ex-
perimentation is, therefore, warranted.
Gebauer & Dietrich (1993) found that carpophores
of ECM fungi were clearly more enriched in "&N
than other ecosystem components studied (trees,
field-layer species, saprophytic fungi, litter), apart
from subsurface soils. Other studies have now
confirmed this pattern (Handley et al., 1996;
Ho$ gberg et al., 1996; Taylor et al., 1997; G.
Gebauer, pers. comm.). Handley et al. (1996) studied
"&N natural abundance in soil–plant systems 187
materials from contrasting environments, viz. a
tropical montane rain forest in Queensland, Aus-
tralia, and coastal forest in Scotland. There were
striking similarities between the two materials ; the
average δ"&N for fungal and other system com-
partment samples was 2±7^ in Scotland compared
to 2±6^ in Australia, and fungi (wood decomposers
excluded) were enriched (8 and 4^, respectively)
relative to other components. The study also demon-
strated a (1^) higher δ"&N in fungal caps relative to
stipes, which was also the case (Taylor et al., 1997)
with ECM fungi in northern Sweden (where the
difference was 2^). As decaying fungal carpophores
are known to release substantial amounts of NH$
(Ingelo$ g & Nohrstedt, 1993), Handley et al. (1996),
in a separate experiment, trapped NH$released from
Agrocybe sp. The NH$
volatilized was indeed
depleted (c. ®40^). However, calculations showed
that it would take a considerable loss of N from the
carpophores to explain their high δ"&N value;
moreover, these high values are also found in young
carpophores. Taylor et al. (1997) tested whether,
because of the risk of NH$
volatilization, different
methods of drying fungal material affected their δ"&N
values, but found no difference between material
dried at 40, 80 and 105 °C.
Ho$ gberg et al. (1996) found that ECM roots of
Norway spruce and beech (Fagus sylvatica) collected
across Europe were roughly 2^ enriched relative to
NON-MYC roots. Fungal sheaths stripped from
ECMs of beech were 2±4–6±4^ enriched relative to
the remaining root core. A simple mixing-model
based on the 2^ difference between ECM and
NON-MYC roots, and the contribution of fungal N
to ECMs, suggested that fungal N should be 3–11^enriched relative to host plant N. This may be
compared with the difference observed at A/ heden,
northern Sweden, of 5–19^, where δ"&N in a range
in carpophores of ECM fungi were found to vary
from ®0±8 to 12.7^, whereas in the foliage of
potential host plants it varied from ®6 to ®5^(Taylor et al., 1997). Hence, the δ"&N of fungal
material surrounding roots was similar to that of the
carpophores in these studies.
Why then is there δ"&N enrichment in fungal
tissue, through which soil N passes into the plant
root? Taylor et al. (1997), recently found that
proteins and amino acids, which together account for
the major portion of fungal N (c. 90% according to
F. Martin, pers. comm.), were c. 10^ enriched
relative to chitin N in ECM carpophores. Their data
implied that the chitin N was only marginally
enriched relative to host plant N. However, fungal
protein and amino acids were appreciably enriched
relative to plant N, which is puzzling, since it is
believed that a glutamine–glutamate shuttle carries
out a bidirectional transport of N between fungus
and host plant (Martin & Botton, 1993), although the
net flux is in the direction of the plant. Differences in
δ"&N between fungal and plant N could be caused by
fractionations during metabolic processes
(cf. III.2(i)) or result from selective retention of
specific N compounds by the fungus.
Of particular interest are orchids, which obtain N
from their mycorrhizal fungi by transfer from, or
possibly by lysis of, hyphae (Smith & Read, 1997).
According to preliminary observations (G. Gebauer,
pers. comm.) some orchids have δ"&N values within
the higher range found amongst ECM fungi at sites
where other plants have typically lower values. Yet
another taxon of interest are the achlorophyllous
Monotropaceae, which parasitize ECM fungi associ-
ated with other hosts (Bjo$ rkman, 1960; Cullings,
Szaro & Bruns, 1996). Delwiche et al. (1979)
reported that two members of that family had c. 10^higher "&N abundance than other plants.
The implication of the above is that symbiotic
fungi can alter the δ"&N of the N they transfer from
the soil to their host plants. The importance of this
has not yet been assessed, nor do we know the nature
of the mechanisms responsible. There is also a
possibility that mycorrhizal fungi have access to
complex organic N pools, with a high δ"&N, which
are not available to NON-MYC roots.
(h) N#-fixation. Delwiche & Steyn (1970) reported a
fractionation of 1±004 in pure cultures of the free-
living diazotroph Azotobacter vinelandii, whereas
Hoering & Ford (1960) and Mariotti et al. (1980a)
found in studies of the same species smaller isotope
effects of 1±0022 and 1±0024, respectively. Shearer &
Kohl (1986) suggested that this discrepancy might
relate to methodological differences. Data from other
species of Azotobacter and symbiotic N#-fixers
indicate that the fractionation is generally small, and
there are even reports of a discrimination in the
opposite direction, i.e. against "%N (Table 1). Several
authors have demonstrated that the fractionation
during N#-fixation is influenced by the bacterial
strain (Steele et al., 1983; Bergersen et al., 1986;
Ledgard, 1989), and Ledgard (1989) provided
evidence that nutrient supply and soil moisture also
influence the fractionation factor.
According to data assembled by Peoples et al.
(1989) from studies of 12 genotypes of nodulated
legumes grown on media free of combined N, shoots
were only slightly depleted in "&N (®0±65³0±2^, P
"0±01), whereas whole plants were not depleted at
all (®0±07³0±12^). Because of the need to account
for fractionation during N#-fixation when estimates
of fixation are based on the "&N natural abundance
method (see IV.3; eqn (12)), many authors have felt
inclined to include by default some arbitrary values
for fractionation from the literature. Comments on
this practice will be given below (see IV.3).
(i) N metabolism in plants. The reader has already
been referred to texts by Raven (1987), Handley &
188 P. HoX gberg
Raven (1992) and Yoneyama (1995), which give
detailed accounts on N isotope fractionations within
plants, and discuss how variations in δ"&N can be
used to elucidate metabolic events. Do fractionations
during such metabolic events confound interpre-
tations of N source use based on plant δ"&N?
Processes such as deamination and transamination
have been associated in other systems with isotope
fractionations of &101 (Hermes, Weiss & Cleland,
1985 and Macko et al., 1986, respectively), and N#-
fixing root nodules are in some cases likewise
considerably enriched relative to the rest of the plant
(Reinero et al., 1983; Shearer & Kohl, 1986;
Yoneyama 1988; Yoneyama & Sasakawa, 1991; Kohl
& Shearer, 1995). There are also reports of a
significant fractionation during NO$
− reduction in
plants (Mariotti et al., 1982, Ledgard, Woo &
Bergersen, 1985b ; Yoneyama & Kaneko, 1989).
These fractionations are comparable to some con-
sidered to be important in the N cycle (Table 1).
However, field ecologists measure the isotopic
signature of total N in plants, and not that of its
constituents, and it has long been recommended to
sample the largest plant N pools, notably foliage.
The field ecologist interested in N-cycling will be
misled by overlooking metabolic processes within
plants if these lead to variations in δ"&N between
different plant parts large enough to interfere with
the interpretation of N source effects. Cautious
interpretation also applies to organs, such as fruits,
receiving N by translocation from the primary sink
(leaves), and with senescent parts likely to have
volatile losses of NH$. Of particular interest is the
site of NO$
− reduction. Some plants reduce NO$
−
primarily in roots, whereas others do so primarily in
the shoot. This balance can change with the supply
of NO$
− (Andrews, 1986), and can result in variable
expression of fractionation during NO$
− reduction in
roots. For example, Evans et al. (1996) found that in
tomato, leaves could be as much as 5±8^ enriched
relative to roots when NO$
− was the N source. When
NH%
+ was the N source, there was no difference in
δ"&N between leaves and roots.
In theory, it is possible that δ"&N can change after
sampling, especially as DeNiro & Hastorf (1985)
attributed a 10–35^ higher "&N abundance found in
archaeological uncarbonized plant specimens, rela-
tive to modern samples, to an isotopic effect of
diagenesis. However, dried conifer needles kept at
room temperature under dry conditions have not
changed their %N during 20 yr of storage, nor do
data on δ"&N of these needles indicate effects of
storage, despite large differences in %N, and,
therefore, in different potentials for ammonia volatil-
ization (Johannisson C. & Ho$ gberg P., unpublished).
Most studies of plant material in the field find
differences between plant parts of 2^ (e.g. Shearer
& Kohl, 1986). For example, Peoples et al. (1989)
found, using data from a variety of studies of
nodulated legumes grown on media free of combined
N, that δ"&N in shoots was on average ®0±66³0±13^(P!0±001) lower than in whole plants, and whole
plants were not enriched relative to the source, N#
(see III. 2(h)). In a detailed assessment of a Scottish
old field Handley & Scrimgeour (1997) found
differences between plant parts of at most 3^. They
also reported average seasonal variations of 2^ in
above-ground parts of broom, but this variability
also included variations among plants. Gebauer &
Schulze (1991) found differences of 1^ amongst
different needle age classes and different canopy
positions in Norway spruce trees; there being a
marginal tendency towards a minor decline followed
by a minor increase in δ"&N with increasing needle
age. Na$ sholm (1994) reported that senescent needles
had a δ"&N of up to 1^ higher than green needles,
and suggested NH$volatilization during senescence
as a possible explanation. Koopmans (1996) reported
differences between needle age classes and year of
sampling of two coniferous forests of 2^ ; other
parts of the trees, except cones (which were 1±5–4^enriched), were not much different from needles.
Ho$ gberg et al. (1996) compared the δ"&N of roots and
needles in N-limited and experimentally N-saturated
Scots pine forest and found that roots were slightly
enriched relative to needles under N-limited con-
ditions, but not under N-saturated conditions. The
difference was c. 2^ for the most superficial roots,
which had the highest %N and are likely to
contribute most to shoot N. This difference can be
explained by the contribution of N from the ECM
fungi to the δ"&N of root N (see III.2(g)). Ho$ gberg
(1986) sampled foliage from a 2 ha plot of Tanzanian
miombo woodland in May 1981 and again in May
1984 (from the same populations of deciduous trees,
but not necessarily the same individuals). Foliar
samples from five individual trees of each of seven
species were taken, and the differences in δ"&N
between years for each species was only 0±30³0±12^on average and varied between 0±03 and 0±83^(Ho$ gberg, 1990a). Samples were also taken from five
individuals of four species from the same plot (but
not necessarily from the same trees) before this, in
May 1980, but those samples were mixed into one
composite sample per species. The intra-species
difference for the four species sampled over all 3 yr
varied at the most between 0±12 and 1±21^.
Interactions between within-plant variability and
sampling strategy set the limits for our interpre-
tations. Based on the above considerations, in field
studies one should be cautious when discussing in
detail differences in foliage "&N between species of
2^, whether they are statistically significant or not.
I recommend the application of the δ"&N natural
abundance method for quantification of N#-fixation
in ecosystem studies only when the δ"&N of foliage of
reference species deviates"5^ from that of N
derived by N#-fixation (see IV.2), if there are no
"&N natural abundance in soil–plant systems 189
complementary data on N pool sizes, patterns of N
transformations, root distribution etc. In the coming
years we will probably see many reports of dif-
ferences in δ"&N between various metabolic constitu-
ents and plant parts. Such studies will have bearings
on field studies only if the N supply mimics the low
levels generally found under most natural conditions,
and will then help to constrain the limits of
interpretations based on δ"&N of plants.
( j) The role of animals. Herbivores, amongst other
effects, redistribute N within the system, and hence
have a variety of potential effects on δ"&N in
soil–plant systems. There are also considerable
fractionations during animal metabolism (e.g.
Gaebler et al., 1963; Gaebler, Vitti & Vukmirovich,
1966), and N leaving the animal via urine is depleted
in "&N, thus the animal itself will become enriched
(although a patch of recently deposited urine will
likely be a site of NH$
volatilization and, therefore,
subsequent enrichment). As a result, there is an
average increase of c. 3–5^ per trophic level
(Minagawa & Wada, 1984). This is of importance for
those interested in estimating the contribution of
animal-N to carnivorous plants (see IV.1). Most
research on δ"&N of animals has aimed at tracing
their food sources (e.g. Peterson, Howarth & Garritt,
1985; Heaton et al., 1986), for which purpose
muliple stable isotope approaches are most useful.
.
1. Estimation of contributions of different soil N and
other non-N#
sources to plant N uptake
Because of the difficulties in assessing the con-
tributions of various soil N sources to plant N uptake
in the field by traditional methods (see III.1 and
III.2( f )) and the large within-site variability in
δ"&N, researchers have hoped that use of specific N
sources can be deduced from the δ"&N of plants. In
extreme cases, e.g. high-latitude systems, variations
between plant species of up to 10^ have been
observed within sites (Schulze, Chapin & Gebauer,
1994; Michelsen et al., 1996; Nadelhoffer et al.,
1996). As will be discussed here, it is, however,
difficult to justify inferences concerning plant use of
specific soil N sources based solely on δ"&N.
Additional and independent non-isotopic or "&N
tracer data are frequently needed to aid in the
interpretation of δ"&N. In fact, in many situations,
"&N tracer methods are likely to be the most powerful.
However, there are some examples, apart from use of
the "&N natural abundance method to assess N#-
fixation (IV.2), where the isotopic signal might
become distinct enough to reveal a contribution of N
from a specific source.
Schulze et al. (1991a) used "&N abundance meas-
urements to assess the contribution of arthropod N
to carnivorous Drosera in Banksia woodlands in SW
Australia, basing the study on the fact that the insect
prey was enriched in "&N (up to 5–10^) relative to
non-carnivorous reference plants or (up to 5^)
relative to mutants of Drosera lacking the insect-
capturing glands. On average, carnivorous Drosera
had, as expected, the highest δ"&N, but not by much;
in fact the most apparently appropriate reference,
the glandless Drosera mutants had only 0±2–1±7^lower values.
Treseder, Davidson & Ehleringer (1995) similarly
examined an assocation between ants and a CAM
epiphyte, Dischidia major, in kerangas forest in
Sarawak. Leaves of this epiphyte roll up to form a
cavity, which is inhabited by Philidris ants. Treseder
et al. (1995) were able to demonstrate that these
leaves obtained 40% of their C by fixing CO#
respired by ants (which fed on C$plants with a δ"$C
different from the CAM epiphyte). They also
estimated that the leaves obtained roughly one third
of their N from ant-debris-derived N based on a
mixing-model of δ"&N values (cf. eqn (11)) of ant
debris (c. 1^), leaves of D. major (c. ®2±3^) and
leaves of the non-ant-inhabited D. nummularia
(c.®3±5^), which is epiphytic on the same host tree
species.
Data on non-ant-inhabited epiphytes collected in
Brazil, Australia and the Pacific (Stewart et al., 1995)
clearly showed that the epiphytes (with a δ"&N of
from ®2.7 to 0^) were depleted (x¯3±4³0±8^)
relative to their non-N#-fixing host trees (with a δ"&N
of from ®1±1 to 3±5^). There are four possible
explanations of this pattern: (i) the epiphytes receive
atmospheric N with a different isotopic signature
than the soil N taken up by the trees, (ii) the
epiphytes are associated with N#-fixing organisms,
(iii) the epiphytes receive N leached (and with an
altered isotopic signal) or volatilized from tree
canopies, and (iv) discrimination against "&N is an
intrinsic function of epiphyte physiology.
By comparison, Schulze et al. (1991c) showed that
mistletoe plants parasitic on trees in the Namib
desert were on average only 0±6^ depleted relative
to their tree hosts, which covered a range from ®3 to
10^. Their data suggested there were no, or only
very small, inputs of N to mistletoe plants other than
soil-derived N obtained via the trees.
Plant canopies can take up gaseous N pollutants
and N in wet deposition. Ammonia is more readily
taken up than NOx
(Wellburn, 1990; Pearson &
Stewart, 1993). The fractional contribution of can-
opy N uptake is highly variable, and difficult to
estimate in field studies. G. Gebauer (pers. comm.)
states that the input from wet and dry deposition can
constitute up to 10–30% of total N taken up by trees.
There has been a considerable interest in the
possibilities of using δ"&N signatures of pollutant N
as a tracer (e.g. Freyer, 1978; Heaton, 1986, 1987;
Garten, 1992). These studies, as well as recent
190 P. HoX gberg
unpublished studies, show that δ"&N of N deposition
varies considerably both spatially and temporally.
Given the similarly large simultaneous variability in
isotopic signature of available soil N it seems
problematic to estimate the fractional contribution
of canopy N uptake based on natural abundance
data, although such data might certainly be helpful
in tracing the origins of N inputs such as those
sampled by rainfall collectors and denuders. Such
data will also help to determine the fate of deposited
N in canopies when comparisons are made with N in
throughfall collectors.
To assess contributions from different soil N
sources based on δ"&N data is in itself quite complex.
However, nitrification leaves the remaining NH%
+
enriched in "&N, and this effect can be substantial
(see III. 2(b)), which might open possibilities in this
context. If nitrification is rapid and leaves little
NH%
+ behind, however, plants will shift to NO$
− as
the N source because of the consequentially low
concentration of NH%
+ in the soil solution and, as a
result, plant δ"&N should, in theory, become lower.
In confirmation of the former suggestion Marschner
et al. (1991) demonstrated that roots of Norway
spruce (Picea abies) used NH%
+ exclusively in a
solution despite the presence of NO$
−, but shifted to
NO$
− as [NH%
+] dropped below 0±1 m. They also
pointed out that in a soil with high cation exchange
capacity, the use of NO$
− should be further favoured
because of the relative immobility of NH%
+.
There are many seemingly contrasting reports in
this context. Pate, Stewart & Unkovich (1993)
studied 24 non-N#-fixing spp. in a Banksia woodland
in SW Australia, and found that δ"&N correlated
positively with shoot in vivo NRA (r¯0±67, P!0±001). Anion-resin-extracted soil NO
$
−, and xylem
sap NO$
−, had the same δ"&N as the leaves of Ptilotus
polystachus, the species with the highest NRA.
However, Pate et al. (1993) did not report the δ"&N of
NH%
+ or other potential sources of N, and NO$
− was
not necessarily the isotopically heaviest component
of available soil-N, although in dry ecosystems such
as this woodland, denitrification during wetter
periods might be an important process (e.g. Skujins,
1981), which should leave enriched NO$
− in the soil.
Nadelhoffer et al. (1996), by contrast, found that of
plants sharing the same site those with the highest in
vivo shoot NRA in an arctic ecosystem had the
lowest δ"&N. Also, in an on-going detailed study of a
swamp forest in Sweden, preliminary data indicate
that field layer herbs with high in vivo shoot NRA
have the lowest δ"&N (L. Ho$ gbom, M. Ohlson & P.
Ho$ gberg, unpublished). In this swamp forest (which
is described by Ohlson & Ho$ gbom (1993)) the
mobile NO$
− should be freely available, and the root
systems of the different species are mainly confined
to the upper 10 cm of soil. In another detailed study
of temperate coniferous forest in Sweden, the
researchers (L. Ho$ gbom, U. Nilsson & G. O$ rlander,
unpublished) have followed shoot NRA and δ"&N in
the grass Deschampsia flexuosa before and after forest
clear-felling at four sites. Their data set comprises
three samples taken each year for 6 yr collected at
five subplots, at each of the four locations. They
found that both NRA and δ"&N, as well as the
concentration of NO$
− in the soil solution, increased
rapidly after clear-felling. Natural abundance of
δ"&N in the grass peaked after 2–4 yr, with values
3±5–7^ higher than in the undisturbed forest (cf.
Fig. 4), while NRA continued to stay high for at least
6 yr after clear-felling. D. flexuosa prefers NH%
+
(Gigon & Rorison, 1972), and it is possible that the
initial increase in δ"&N resulted from uptake of NH%
+
enriched in δ"&N because of fractionation during
nitrification, followed by a drop in δ"&N as NO$
−
became the dominant source along with the build-up
of a nitrifying population of micro-organisms. In the
experimental site at Norrliden, where N had been
added annually at four levels (N0–N3) as urea or
NH%NO
$, δ"&N was positively correlated with in vivo
NRA in D. flexuosa (Ho$ gberg, 1990b). The maximal
isotope effect due to additions of N, the difference
δN$
®δN!
in the grass, was smaller in the NH%NO
$-
treated plots (4^), than in the urea-treated plots
(12^), although plant NRA was comparable, which
suggests that the enrichment in the grass was largely
a result of fractionations during processes in the soil
(and NH$
volatilization in the case of urea appli-
cations) rather than within plants (based on as-
sumption that fertilizer N was not enriched in the
first place: see IV.4). These examples underline the
need for additional independent non-isotopic or
δ"&N tracer data and}or modelling of δ"&N natural
abundance in interpretations of the contributions of
different N sources to plant δ"&N.
There can be considerable variation in δ"&N with
soil depth (see IV.3), and this could affect the
isotopic signature of plant N (e.g. Ledgard, Freney
& Simpson, 1984). In profiles of temperate co-
niferous forest soils with a typical upper organic mor
layer, there is commonly a decrease in δ"&N of
5–10^ with increasing soil depth within the upper
dm of soil (see IV.3). Gebauer & Schulze (1991) and
Ho$ gberg et al. (1996) demonstrated that in the
uppermost layer, δ"&N of roots was!1^ depleted
relative to soil total-N (only at the A/ heden site in
northern Sweden was the difference larger), whereas
a few cm further down this difference had increased
to"4^. Data from both groups indicated a pre-
dominance of N uptake from superficial soil layers
by the European conifers studied. Based on the
limited evidence available there seems to be limited
retranslocation of N from roots during senescence
(Nambiar, 1987), and data on "&N abundance of roots
suggest there is little transfer of N from roots in one
soil horizon to roots in another. In arctic tundra, low
δ"&N occurred in shallow-rooted ericaceous spp. and
Betula nana, whereas high values occurred in deeply
"&N natural abundance in soil–plant systems 191
rooted sedges (Nadelhoffer et al., 1996). In the forest
experiment at Norrliden, Ho$ gberg et al. (1996)
compared δ"&N of fine roots of different species, by
horizon, on severely N-limited control plots and
experimentally N-saturated plots. Under conditions
of N-limitation there were differences of at most
1±8^ between species (ECM Scots pine, AM
Deschampsia flexuosa and Vaccinium spp. with ericoid
mycorrhiza) within each horizon, whereas on N-
saturated plots differences between roots of the
different species could be up to 5^. These observa-
tions tallied with data on plant foliage, and will be
discussed in more detail below (see IV.4). In the type
of site discussed here, the data suggested that, under
conditions of N-limitation, larger differences among
plant taxa for δ"&N are likely to be caused by
differences in rooting patterns, rather than by
differences in the use of various N species in the
same loci of soil.
As summarized by Nadelhoffer et al. (1996), the
δ"&N abundance of plants depends on (i) the source
of plant N (e.g. soil, precipitation, gaseous N
compounds, N#-fixation), (ii) the depth in soil from
which N is taken up, (iii) the form of N used (e.g.
NH%
+, NO$
−, organic N sources), and (iv) the
influence of mycorrhizal symbioses and fraction-
ations during and after N uptake by plants. To this
one can add interactions between these factors and
plant phenology. Again, the isotopic signature of a N
source is not a constant, but is dependent on its
origin and the character of N transformations in the
specific system. Thus, data on plant δ"&N cannot be
used directly in comparisons between ecosystems,
but may assist in interpretations of plant N source
use in comparisons within ecosystems, notably in
experimental settings and in combination with other
data and modelling.
2. Estimation of N#-fixation by the "&N natural
abundance method
The process of N#-fixation is highly variable both
temporally and spatially. In some methods used to
quantify N#-fixation, it is hoped that a meaningful
average rate will be derived by measuring a character
which is believed to integrate the effect of the process
over time, viz. the N difference method, the "&N
tracer dilution method and the "&N natural abun-
dance method (Bergersen, 1980; Peoples et al.,
1989). Another category of methods represent point-
measurements in time, and therefore need to be
repeated to incorporate temporal variability, viz. the
acetylene reduction method and the xylem–solutes
method (Bergersen, 1980; Peoples et al., 1989). Each
of these methods has its own merits and dis-
advantages, and over a wide range of environmental
conditions and logistic constraints it is difficult to
identify any one of them as the superior. The "&N
dilution method is no doubt, in theory, the most
precise, but is problematic to apply to deep-rooted
perennials in the field, because it is difficult to label
a deep soil profile uniformly with tracer "&N (Peoples
et al., 1989), and because the "&N enrichment of
available N pools will vary considerably over time
after the addition of the tracer (Witty, 1983).
Perennial plants also pose a problem because of their
large N content before labelling.
The "&N natural abundance method (e.g. Amarger
et al., 1977; Shearer & Kohl, 1986) is based on the
same principle as the "&N tracer dilution method, i.e.
the "&N abundance of a N#-fixing species (
fix), which
obtains N from atmospheric N#
in addition to
combined soil N sources, is compared with that of
one (or several, cf. Ledgard et al., 1985a, b) non-N#-
fixing reference species (ref
), which rely solely on
soil-derived N (the contribution of N via the
atmosphere directly to the plant canopy, e.g. from
NH$or NO
x, is ignored or thought to be the same for
the two plants). The calculation of the fraction N
derived from N#-fixation (N
dfa) is made as follows
(Amarger et al., 1977):
Ndfa
¯ (δ"&Nref
®δ"&Nfix
)}(δ"&Nref
®B), (12)
where B is the δ"&N of the N#-fixing plant when
totally dependent on N#, and is included to account
for the fractionation during the process of fixation
(see III.2(h)). Reference species may have positive
or negative δ"&N (e.g. Vitousek, Shearer & Kohl,
1989); eqn (12) will work either way. The method is
often described as if reference species need to have
positive δ"&N, but this is not the case. The use of B
is discussed below. Some researchers argue (L. L.
Handley, pers. comm.) that the basis of the method
is doubtful as there might be soil-derived N sources
with the same isotopic signature as N#derived from
fixation. It is likely that this happens, and is why one
should use several reference species, and be careful
when interpreting small differences between the two
groups of species. The ultimate proof that an
organism is N#-fixing is to test if it assimilates "&N
#.
The apparent merit of the "&N natural abundance
method is that nothing has to be added or disturbed,
and it provides opportunities to indicate whether
deep-rooted plants and large perennials are diazo-
trophic or not. These are cases where the "&N
dilution method and the acetylene reduction method
are not practically feasible, as the former requires
"&N labelling of the rooting zone, and the latter
requires access to root nodules. For example, root
nodules of Prosopis can be confined to a narrow zone
just above the groundwater table at several metres
depth (Felker & Clark, 1982), and roots of Prosopis
spp. and other phraetophytes can reach"50 m
below the soil surface (Phillips, 1963; Canadell et al.,
1996). In miombo woodland in Africa it took up to
5 d of excavation to confirm nodulation in a tree
species, although these excavations were largely
192 P. HoX gberg
d15N
(‰
)
2
0
–2
(a)
(b)
6
4
2
0
4
2
0
6(c )
0 2 4
%N
Figure 3. Samples of surveys attempting to identify N#-
fixing spp. (E) in mixtures with non-N#-fixing spp. (D)
based on differences in "&N natural abundance. Data from
(a) a miombo woodland in Tanzania (Ho$ gberg, 1986,
1990a), (b) a miombo woodland in Zambia (Ho$ gberg &
Alexander, 1995), and (c) a lowland rain forest in
Cameroon (Ho$ gberg & Alexander, 1995).
made in the upper 0±5 m of the soil profile (Ho$ gberg,
1986). Negative evidence from such efforts did not
prove unequivocally that a species is non-nodu-
lating; only a minute portion of the fine roots had in
practice been searched, and there was a risk that
nodules were disconnected from roots during the
work. On the other hand, the "&N natural abundance
method requires that reference and diazotrophic
species have (i) similar root distributions, (ii) similar
temporal N uptake patterns, and (iii) the same
preferences for the various species of inorganic and
organic N in the soil. The disadvantages of the
method relate to these three issues, and are height-
ened in situations where the difference in δ"&N
among potential reference species is large, for no
obvious reason, compared to the difference between
the two groups of species (Shearer & Kohl, 1986;
Hansen & Pate, 1987; Ho$ gberg, 1990a ; Handley,
Odee & Scrimgeour, 1994).
Thus, the usefulness of the "&N natural abundance
methods is directly related to the characteristics of
the species at the site investigated. This is illustrated
in Figure 3, which presents results from a number of
surveys, displayed as plots of δ"&N vs. %N of foliage,
which may better separate N#-fixing from non-N
#-
fixing plants than histograms of δ"&N values only,
especially under N-limiting conditions, when non-
N#-fixing species are likely to have low %N
(Ho$ gberg, 1986; Ho$ gberg & Alexander, 1995).
However, an argument against the use of %N in this
way is that it has been proposed that legumes,
irrespective of whether they form N#-fixing symbi-
oses or not, tend to have a high %N and to be
adapted to a N-demanding lifestyle (McKey, 1994).
The cases displayed in Figure 3 are:
(a) A miombo woodland in Tanzania (Ho$ gberg,
1986, 1990a), where N#-fixing spp. (all nodulated by
active rhizobia as shown by C#H
#-reduction tests)
had δ"&N values close to or slightly below 0^, and a
%N of 1±7–3±0, whereas non-N#-fixing reference
species had either higher (ECM spp.), or similar or
lower (AM spp.) δ"&N, and similar or lower %N.
Here, the "&N natural abundance method can only
give a weak indication of N#-fixation in some of the
putative N#-fixers; there is no basis for a quantifi-
cation in this case. The AM spp. are the proper
reference species because the N#-fixing spp. also
have AM. At this site, C#H
#-reduction tests con-
ducted after tedious root excavation work proved to
be a better indicator of N#-fixation.
(b) A miombo woodland in Zambia (Ho$ gberg &
Alexander, 1995), where putative N#-fixing spp. had
δ"&N values grouped around 0^, and comparatively
high %N. Baphia bequaertii, a species likely to fix N#
because of its taxonomic position, shared these
characteristics, as did Cassia abbreviata, a species
highly unlikely to form root nodules. Here, the data
indicate that putative N#-fixers probably do fix N
#,
but also show that non-N#-fixing spp. might share
their foliar characteristics. The basis for a quanti-
fication may thus be questioned. As an aside,
however, during a survey in the USA (Virginia &
Delwiche, 1982), foliage δ"&N of the non-legume
Chamaebatia foliolosa (Rosaceae) indicated it could
be N#-fixing, which was confirmed by excavation
work finding active ("&N#-reducing) root nodules
(Heisey et al., 1980).
(c) A lowland rainforest in Cameroon (Ho$ gberg &
Alexander, 1995) where there were no major dif-
ferences between non-N#-fixing spp. and putative
N#-fixing spp. Here, because of the relatively high
δ"&N of all of the reference spp. it should have been
possible to identify a species with substantial N#-
fixation, and to make semi-quantitative statements
about the fractional contribution of fixed N.
Very many surveys of natural communities have
demonstrated results similar to those displayed in
Figure 3, e.g. Delwiche et al. (1979), Virginia &
"&N natural abundance in soil–plant systems 193
Delwiche (1982), Shearer et al. (1983), Hansen &
Pate (1987), Yoneyama et al. (1990, 1993), Schulze et
al. (1991b), Handley et al. (1994), and Sprent et al.
(1996). Some authors have attempted to quantify N#-
fixation by using eqn (12). Clearly, quantification in
ecosystems where complementary data on rooting
patterns and N transformation patterns are lacking
requires a large (at least 5 or ®5^) difference in
δ"&N between non-N#-fixing reference species and N
derived by N#-fixation (B), and that there be no large
unexplained variability in δ"&N amongst reference
spp. Preferably one should have several reference
species, which should ideally differ in their known or
anticipated rooting patterns and temporal N uptake
patterns. One should also be cautious about the value
of B in this context (cf. eqn (12)), as it has been
shown to vary with bacterial strain as well as abiotic
factors (see III.2(h)). It seems impossible to con-
strain the value of B where there are deep-rooted
species which might harbour many rhizobial strains
in their root nodules. At most one could test the
potential importance of the most extreme values of B
reported (cf. Table 1).
Pot experiments and agricultural settings have
enabled comparisons to be made between the δ"&N
and the "&N dilution methods. In pot experiments,
where both methods, notably the "&N dilution
method, should be very precise, differences in Ndfa
were!10% in studies summarized by Shearer &
Kohl (1986). Bremer & van Kessel (1990) compared,
in a field study, the "&N natural abundance method
with the "&N dilution method. They conducted this
study at several sites over a 3-yr period using up to
four reference species and two N#-fixing species.
Estimates of N#-fixation (%N
dfa) did not differ in 18
out of 21 comparisons of mean values; the largest
difference was 33%, but the difference was only"10% in eight cases, averaged only 9±2³1±8%, and
was not directional. In a more limited but similar
field study Bergersen & Turner (1983) found a
largest difference of 33% between the two methods
and an average difference in four cases of 13³7%. It
should be emphasized here that although the "&N
dilution method is theoretically more precise, it is by
no means clear that in complex field situations it
provides a more accurate estimate of Ndfa
than does
the "&N natural abundance method.
When analysis of "&N natural abundance was first
discussed as a means to follow the N cycle there was
concern that spatial variability of soils would be a
severe methodological constraint (Cheng, Bremner
& Edwards, 1964; Shearer, Kohl & Chien, 1978;
Broadbent et al., 1980; Kohl et al., 1981) . However,
if potentially N#-fixing plants and the non-N
#-fixing
reference plants are sampled only where they grow in
close proximity, this problem should be minimized
(Shearer & Kohl, 1986).
In conclusion, it is sometimes possible to obtain
quantitative data on Ndfa
based on the δ"&N natural
abundance method, but frequently in the field the
difference in δ"&N between non-N#-fixing species
and B is too small, and one cannot be certain about
the quality of the reference species, i.e. whether or
not they meet the requirements (i)–(iii). More often,
the method will at best only provide an indication
that a species might be N#-fixing. However, in
surveys of lesser known types of vegetation this may
be very valuable, and indeed sometimes the only
practical way of assessing whether N#-fixation
occurs.
Binkley, Sollins & McGill (1985) were the first to
try to trace a transfer of fixed N from N#-fixing
species (alders) to co-existing non-N#-fixing species
(conifers) by using the δ"&N of plants and soil
inorganic N pools ; they concluded that fraction-
ations of N isotopes in the soil make this difficult. By
contrast, van Kessel et al. (1994b) claimed that a
decline over time in δ"&N of understory plants under
N#-fixing Leucaena leucocephala shrubs was evidence
of a transfer of fixed N. It might well be that a
transfer of fixed N was responsible, but the decline
in δ"&N of the non-N#-fixing plants could also have
been caused by other processes, e.g. a decline in
nitrification (which may follow after the transient
disturbance connected with stand establishment).
Such a decline seems to be a common feature in
forest ecosystems after disturbances such as clear-
cutting, fire etc. (see IV.1), and should affect δ"&N of
plant N and thereafter δ"&N of soil profiles via litter-
fall (see IV.3). The control in a study like this should
be δ"&N of understory plants under one or several
non-N#-fixing tree species, but one also has to
consider that N-inputs from N#-fixing species might
stimulate nitrification, which can affect the δ"&N of
available N in unexpected ways.
3. Interpretation of δ"&N profiles in soils (with
comments on horizontal spatial variability)
In most ecosystems studied, plants have been found
to have a δ"&N lower than that of soil total-N
(Ledgard et al., 1984; Shearer & Kohl, 1986;
Nadelhoffer & Fry, 1994). Redeposition of "&N-
depleted plant N onto the soil surface by litter-fall
explains why δ"&N of soil surfaces in many forest
ecosystems is lower than further down in the soil
(Nadelhoffer & Fry, 1988, 1994). In many agri-
cultural systems plants are also depleted in "&N
relative to soil total-N (e.g. Meints et al., 1975a ; van
Kessel, Farrell & Pennock, 1994a), but above-
ground parts are taken away at harvests or mixed
into deeper soil layers by ploughing. In other
systems, soil animals carry out a similar mixing, and
in these cases there is not a "&N-depleted surface
horizon. The increase in δ"&N from the surface
downwards in the upper dm of soil in forests is
between 5 and 10^, or even more (Riga, van Praag &
Brigode, 1971; Mariotti et al., 1980b ; Wada,
194 P. HoX gberg
Imaizumi & Takai, 1984; Nadelhoffer & Fry, 1988;
Gebauer & Schulze, 1991; No$ mmik et al., 1994;
Ho$ gberg et al., 1996; Koopmans, 1996; Piccolo et
al., 1996). Nadelhoffer & Fry (1988) ruled out
selective preservation of "&N-enriched compounds
during decomposition, as well as illuviation and
changes in δ"&N of N sources, as causes of the profiles
observed in forest soils. Accordingly, fractionation
against "&N during the mineralization–plant uptake
pathway and deposition of this "&N-depleted N onto
the soil is probably the major cause of the "&N-
depleted surface layer.
Another process to consider is ammonia volatil-
ization from litter during decomposition (Turner,
Bergersen & Tantala, 1983). As removal of N by this
process, plant uptake, or leaching of NO$
− produced
by nitrification tends to enrich the remaining N with
"&N, the δ"&N of litter is related to its stage of decay.
Ehleringer et al. (1992) took the lack of change in
δ"&N with depth in deep litter profiles as evidence of
lack of decomposition under Prosopis shrubs in the
Atacama desert, a suggestion corroborated by data
on δ"$C and "%C-datings. Yet another process to
consider in this context is the possible synthesis in
situ of "&N-enriched compounds by microbes during
decomposition. As discussed above (III.2(g)), fun-
gal N can be considerably enriched relative to other
ecosystem components, and it is also an important
precursor of recalcitrant N in soils (e.g. Paul &
Clark, 1989). As recalcitrant N accumulates with
increasing soil depth this process might contribute to
the increase in δ"&N down the profile. This, of
course, has implications for the interpretation of the
difference between δ"&N of soil total-N and root N.
Tiessen et al. (1984) found that the old heavy
stabilized fraction of soil N, which comprises
aggregates of organic matter and clay particles, had
δ"&N values"12^, whereas the lighter sand fraction
was"5^ lighter, in native and cultivated prairie
soils.
Forest fires consume the upper δ"&N-depleted
surface layer, which forces plants to find N in lower
horizons and leads to an increase in δ"&N of plants, to
which an increase in nitrification after the fire
(Raison, 1979) may contribute further by providing
"&N-enriched NH%
+ (Fig. 4). In an experimental
study in our laboratory, δ"&N of Deschampsia flexuosa
and Vaccinium vitis-idaea increased up to 3–4^ with
increasing burn depth (P. Wikstro$ m, pers. comm.).
Other disturbances, such as clearfelling of forest,
may cause similar changes (see IV.1), and changes in
N cycle patterns will ultimately feed back on soil
profile development (see IV.4 and Figs 4, 5).
It could be suggested that in N-limited, N-
aggrading forests, low rates of nitrification mean no
isotopic enrichment of NH%
+. If NH%-N is preferred
by the trees, uptake might result in progressively
depleted N at the soil surface, as isotopically depleted
litter will be the starting point for the N mineral-
d15N
(‰
)
Time
Major disturbancee.g. fire or clearfelling
Minordisturbancee.g. thinning
d15N d15N
So
il d
ep
th
So
il d
ep
th
Figure 4. Hypothetical development of δ"&N of foliage of
northern temperate coniferous forest trees over time. The
δ"&N of soil profiles at two points in time are shown in
inserts. The time scale spans c. 50 yr.
ization–plant uptake pathway (Fig. 4). Data on
current needles in 20-yr-long time series from plots
receiving no fertilizer N, and where N deposition
levels have been low, show a slow decline in δ"&N
over time (Ho$ gberg, 1991; Ho$ gberg et al., 1995;
Johannisson, 1996). As these plots were non-
fertilized control plots located in between heavily N-
fertilized plots, a contamination by "&N-depleted N
(as NO$
−) leached from experimentally N-saturated
plots is, however, also possible (Ho$ gberg, Tamm &
Ho$ gberg, 1992; Ho$ gberg et al., 1995; see IV.4).
This suggestion is supported by the fact that the
decline on control plots was most pronounced (c.
®4^ over 15 yr) in the Stra/ san trial, which has the
steepest slope of the trial sites and is, therefore, the
most likely to have a flow of solutes between plots.
Poulson, Chamberlain & Friedland (1995) reported,
in a detailed study of δ"&N of tree rings in wood, a
decline of the same order; this decline could be
caused by a decline in δ"&N of N inputs to the site,
metabolic processes within trees, or the feedbacks to
profile development described above.
By contrast, in situations where high N inputs
promote nitrification, and thus δ"&N-enriched NH%
+
production, plants preferentially using NH%
+ as N
source will progressively enrich the surface soil
(Ho$ gberg et al., 1996; Johannisson, 1996; Fig. 5; see
IV.4). Under such circumstances NO$
− depleted in
"&N is readily lost from the upper part of the soil
profile, but might be partly retained further down.
This process might explain why the increase in δ"&N
with depth, near the surface, can turn into a decrease
further down in the soil (e.g. Riga et al., 1971;
Karamanos & Rennie, 1980b ; Karamanos, Voroney
& Rennie, 1981; Piccolo et al., 1996).
In conclusion, the δ"&N abundance of undisturbed
forest soil profiles can provide information about the
N cycle in forest ecosystems: comparatively low "&N
abundance in the surface layer appears to indicate N
limitation and low rates of nitrification, whereas a
"&N natural abundance in soil–plant systems 195
d15N
(‰
)
Time
N-limitation:little nitrication
N03-
becomesthedominantN source
d15N d15N
So
il d
ep
th
So
il d
ep
th
N-saturation:nitrificationproducing15N-enriched NH
4+
Figure 5. Hypothetical development of δ"&N of foliage of
forest trees during a phase of N saturation due to high rates
of N-deposition. The δ"&N of soil profiles at two points in
time are shown in inserts. The time scale spans c. 50 yr.
higher δ"&N in the surface layer than in deeper layers
appears to indicate high rates of nitrification, which
under humid conditions correlate with loss of N
from the system (Ho$ gberg et al., 1996; Johannisson,
1996; Nohrstedt et al., 1996; Na$ sholm et al., 1997).
If the "&N-depleted NO$
− stayed in the horizon
where nitrification occurred, the isotopic mass
balance would not change, and there would not be a
shift in the isotopic composition. Variations in δ"&N
of soil profiles might also result from physical mixing
of soil layers, and their removal by fires.
Spatial variability in δ"&N of soils can be con-
siderable, as has been shown in detailed studies of
arable lands. For example, Sutherland, van Kessel &
Pennock, (1991) studied variability in plants and
soils at two scales, using an 11¬11 m grid and an
110¬110 m grid with 144 sample points each, on an
irrigated field with durum wheat. At the smaller
scale, variability appeared to be random, whereas at
the larger scale high δ"&N values in soils and plants
were associated with depressions, where denitri-
fication activity was high. The variability in topsoil,
0–10 cm deep, (6±2–10±3^) was much less than in
plants (1±6–24±4^) reflecting the much more dy-
namic nature of available-N than total-N. Garten
(1993) found low variability in mineral soil δ"&N in a
deciduous forest watershed in SE USA, but larger
variations in plants; high values being found in
plants in valley bottoms, where net nitrification rates
were high. In a subsequent survey of a larger area,
Garten & van Miegroet (1994) confirmed the
correlation between "&N of plants and net nitri-
fication rates. These results from forests probably
reflect variations in nitrification, fractionation
against "&N during nitrification and a preference for
NH%
+ amongst the plants (see III, IV.1 and 4).
Karamanos and Rennie (1980a) found that the
δ"&N of NO$
− moving with the groundwater flow
towards discharge areas in agricultural fields was
comparatively low. In the discharge areas the δ"&N of
both NO$
− and soil total-N was clearly higher,
suggesting denitrification. Detailed data of this kind
are lacking for forest ecosystems.
4. Assessment of N balances of ecosystems
Early studies found that soils, notably below the
surface horizon, were frequently enriched in "&N
relative to the atmosphere (e.g. Cheng, Bremer &
Edwards, 1964; Delwiche & Steyn, 1970). It has
been suggested (Handley & Raven, 1992) that this is
caused by the roughly10 times larger isotope effect of
denitrification compared with that of N#-fixation (cf.
Table 1). For example, if the flux into the biosphere
via N#-fixation is 10 times that out of the biosphere
via denitrification, then the biosphere will have the
same δ"&N as atmospheric N#; if losses via denitri-
fication are comparatively larger, or if the frac-
tionation during denitrification is comparatively
stronger, then the biosphere will become more
enriched than the atmosphere. In reality, the picture
is more complex since large losses from the biosphere
occur via NH$
volatilization, via N#
and N#O in
denitrification, and via NOx
from combustion of
fossil fuels, whereas larger inputs occur through
biological N#-fixation (c. 50%), industrial fixation of
fertilizer N (c. 25%) and re-assimilation of NH$and
NOx
(Jenkinson, 1990). However, in the past, N#-
fixation and denitrification were the major deter-
minants of the gross balance between N in the
atmosphere and in the biosphere.
Seabird rookeries are amongst the most naturally
N-enriched ecosystems, and they have very high
δ"&N values (Mizutani, Hasegawa & Wada, 1986;
Mizutani & Wada, 1988; Mizutani, Kabaya & Wada,
1991,). This is caused by a combination of the high
values δ"&N values in the fish prey, and N losses at
the rookeries, e.g. by NH$
volatilization, and nitri-
fication followed by loss of NO$
−. There is also an
interesting trend of increasing δ"&N enrichment of
soils (ε¯ δsoil
®δbirddroppings
) with increasing latitude
(Mizutani et al., 1991b). This correlation can be
explained by a larger kinetic isotope effect during
NH$
volatilization at lower temperatures (Wada,
Shibata & Torii, 1981) and possibly also by more
leaching of ("&N-depleted) NO$
− in colder, wet
situations (cf. the presence of guano NO$
− deposits
in the tropics and sub-tropics). (Incidentally, Wada
et al. (1981) speculated that fractionation during
diffusion of gaseous N pollutants from lower lati-
tudes would lead to depletion of "&N in the very small
amounts of N deposited at high latitudes. Pre-
liminary work at Svalbard, 79° N, however, does not
indicate that the plants in the more common
ecosystems there use N with lower δ"&N than plants
in low latitude systems (L. Ho$ gbom, I. J. Alexander,
M. Ho$ gberg & P. Ho$ gberg, unpublished)). In the
extreme case of Antarctic seabird rookeries, the
value of ε (¯ δsoil
®δbirddroppings
) exceeded 25^
196 P. HoX gberg
(Mizutani et al., 1991b). It was proposed that N
isotope ratios could help to identify deserted seabird
rookeries (Mizutani et al., 1991a), but for most
researchers analysis of phosphate would be just as
effective and cheaper (Arrhenius, 1931; Dauncey,
1952).
There is growing concern over potential effects of
deposited N on natural ecosystems (e.g. Aber et al.,
1989), and "&N abundance can provide insights into
the N balance of an ecosystem, as in the examples
above. The legitimacy of tracing of N inputs and
outputs has been a major issue of debate within the
field of "&N abundance studies. The debate started
with a contribution by Kohl et al. (1971) and was
immediately followed by a critical remark from
Hauck et al. (1972, see also reply from Kohl, Shearer
& Commoner, 1972), and later more detailed
assessments (Edwards, 1973; Focht, 1973; Hauck,
1973). This discussion concerned the use of fertilizer
N in agriculture and the ground and stream water
pollution that could result from it.
Kohl et al. (1971) established a negative cor-
relation between concentration of NO$
− and δ"&N of
NO$
− in stream water in an agricultural area, and
used a mixing model to calculate the contribution of
fertilizer N to stream water NO$
− concentration. As
discussed above, δ"&N of fertilizer N is not a
conserved tracer within the soil (e.g. Hauck et al.,
1972, see also III.1). Hence, although excessive use
of fertilizer N will no doubt lead ultimately to
increased levels of stream water NO$
−, data on δ"&N
cannot alone reveal the source of NO$
− or allow its
quantification. An interesting advance within this
field is the measurement of both δ"&N and δ")O in
NO$
− (Amberger & Schmidt, 1987; Aravena et al.,
1993), especially since the δ")O of deposited NO$
− can
differ from that of NO$
− produced by nitrification in
soil by c. 50^. In the latter process two O atoms
originate from soil H#O and one O atom from O
#gas
in the soil. Durka et al. (1994) used this promising
approach to attempt to distinguish between these
two sources of NO$
− in stream water in forested
catchments in Germany. However, the δ")O of NO$
−
produced by nitrification might vary more widely
than assumed because of variations in the con-
tribution of respired CO#
to the δ")O signal of H#O
at different soil depths (C. Kendall, pers. comm.).
At low rates of additions, fertilizer N contributes
directly as a source effect to the δ"&N of soil–plant
systems. Meints et al. (1975 b) compared unenriched
and δ"&N-enriched fertilizer as tracers for N fertilizer
uptake, and found that unenriched fertilizer under-
estimated uptake of fertilizer N in four out of six
cases. Kohl, Shearer & Commoner (1973) and
Shearer & Legg (1975) demonstrated a decrease in
δ"&N towards that of fertilizer N with increasing rate
of N addition to corn and wheat, respectively. At
higher rates of continuous additions this simple
pattern breaks down as the added N starts to have
profound effects on soil N transformations and the
input–output balance. For example, in agricultural
and forested systems with annual additions of N
(Meints et al., 1975a ; Ho$ gberg, 1990a, b, respect-
ively), low rates of additions lowered (or did not
change) the δ"&N of plants, whereas high addition
rates increased the δ"&N. In the above examples of
agricultural systems, the effect on δ"&N of soil total-
N were, as expected, small–due to the large amounts
of N already present in the soil. Meints et al. (1975a)
suggested that the increase in δ"&N of plants at high
rates of N addition could, amongst other things,
result from the fractionating processes associated
with N losses, and this speculation was repeated by
Ho$ gberg (1990b, 1991).
In the forest experiment at Norrliden, annual
additions of N at a rate of c. 30 kg N ha−" (treatment
N1) led to a decrease in δ"&N abundance of current
year needles over a period of 2 decades (Ho$ gberg,
1991; Johannisson, 1996), and this mild N treatment
did not differ much from the control in this respect.
By contrast, high additions of N (N2 and N3; 2 and
3 times N1, respectively) led to an increase in δ"&N of
current needles. This isotope effect was small
(!2^)whenNH%NO
$wasaddedcomparedwith the
increase when urea was added ("5^). Potentially,
fractionations during NH$
volatilization, nitrifi-
cation followed by leaching or denitrification, and
denitrification itself can explain these isotope effects.
Ammonia volatilization appears to play a minor role
as the first 2 yr of fertilization with urea at high rates
(180 kg N ha−" yr−" in treatment N3 during those
years) did not change the δ"&N of needles, although
their %N increased by c. 50% (Johannisson, 1996;
Na$ sholm et al., 1997). (Incidentally, No$ mmik et al.
(1994) found an increase (from c. ®7 to ®1^) in
current year needles the second year after urea
fertilization, but this did not necessarily involve NH$
volatilization as the δ"&N of the fertilizer was ®1±1^,
i.e. it could simply reflect the isotopic source effect.)
This observation, as well as the decline in δ"&N of
needles in treatment N1 at Norrliden over the first
15–20 yr suggests that the positive isotope effect was
not a result of a positive source being added to the
system. The small variations between years suggest
that the isotopic composition of fertilizer N was
relatively constant (Johannisson, 1996). (The iso-
topic signature of fertilizer N can vary depending on
the process of manufacturing (Hu$ bner, 1986), and
the NO$
− in NH%NO
$is generally a few per mil
enriched relative to the NH%
+.) As pointed out by
Johannisson (1996), the positive isotope effect on
N2- and N3-treated plots in the Norrliden trial
probably results mainly from nitrification, and the
preference for uptake of NH%
+ by the trees. The
larger isotope effect in the urea treatment was
presumably caused by higher rates of nitrification
there; levels of extractable NO$
− have consistently
been found to be the same in plots given either of the
"&N natural abundance in soil–plant systems 197
two N source treatments at levels N2–N3, although
half of the N added as NH%NO
$was already in the
form of NO$
−. Moreover, at these rates the biological
demand for N is exceeded, which leaves ("&N-
depleted) NO$
− to be leached and lost from the
system (Johannisson, 1996). There was a positive
correlation between the change in δ"&N in current
year needles from 1970–1989 and the calculated
fractional losses of added N (Ho$ gberg & Johan-
nisson, 1993), and this shift could also be traced in
the soil profile (Johannisson, 1996). Nitrogen-
saturated forests in Central Europe had profiles
similar to those of the N3 plots at Norrliden, i.e. with
a comparatively high δ"&N in the surface layer, but
not always the more typical clear increase with depth
(Ho$ gberg et al., 1996). Moreover, in surveys of non-
N-fertilized forests, Garten (1993) and Garten & van
Miegroet (1994) found a correlation between net
nitrification potential and ε, here defined as ε¯δ"&N
leaves®δ"&N
soil. Thus, data from both forest
experiments and surveys of forests suggested that
the use of foliage δ"&N, ε or profile studies of soil δ"&N
might be developed as tools in surveys to monitor N
saturation of forests.
Nohrstedt et al. (1996) found, in a study of the N
cycle in three Norway spruce forests in SW Sweden,
that δ"&N and concentrations of arginine in foliage
were positively correlated with leakage of nitrate.
Arginine is widely used as an indicator of N excess in
coniferous trees (e.g. van Dijk & Roelofs, 1988;
Na$ sholm & Ericsson, 1990; Ericsson et al., 1993),
and in a larger survey Na$ sholm et al. (1997) found a
correlation between arginine and ε. In a complemen-
tary study of material from the Norrliden ex-
periment they demonstrated that the two parameters
are not necessarily directly related, but can be
correlated because they reflect two aspects of N-
saturated forests ; N excess in trees and nitri-
fication, respectively. In their survey of 23 sites
Na$ sholm et al. (1997) encountered elevated concen-
trations of arginine and comparatively high ε at the
five sites having detectable leaching of NO$
−.
However, the same was true for three sites without
detected leakage of NO$
−, which perhaps illustrates
the weakness of the lysimetry approach used (only
three tension lysimeters were used at each site).
Emmett et al. (1997) observed a correlation between
ε and N content in throughfall when data from four
European sites were used; a fifth site, the Welsh
Aber forest, did not conform to this pattern, differing
in several respects, e.g. deep ploughing of a relatively
N-rich subsoil, which could have directly affected ε.
However, data from Aber contributed to the positive
correlation between nitrification rates and ε. Koop-
mans (1996) pointed out the need to account for
variability in isotopic signature of N inputs in studies
of this kind; the mean value for NH%-N in bulk
precipitation was ®0±6^ at one site, but 10±8^ at
another.
In conclusion, changes in the input–output bal-
ance can affect the δ"&N of ecosystems. Soil total-N
represents a large source effect in this context, and
changes slowly. When an ecosystem is subjected to
high levels of N input, more rapid changes in δ"&N
will take place in the active inorganic N pools ; NH%
+
will become enriched during nitrification and am-
monia volatilization, whereas NO$
−, although de-
pleted because of fractionation against δ"&NH%
+
during nitrification, will become enriched during
denitrification. Experimental data and surveys have
indicated that ε of forest plants can be used as an
indicator in regional surveys of the stage of N-
saturation of forests, but this tool is better suited to
studies of dose-responses in more confined ex-
perimental settings. I have previously stated
(Ho$ gberg et al., 1995) that plants with a high
capacity for uptake of NO$
− would be ideal for
studies of this kind. This is not correct, since the
higher enrichment occurs in the NH%
+ pool. Hence,
the conifers with their preference for uptake of this
N form, or other plant species with this same
preference, are good candidates for such studies.
.
Above all, this review has highlighted the complex
factors affecting δ"&N in plants, and problems of
interpreting this isotopic signature. This situation is
in marked contrast to interpretation of "$C measure-
ments, where there is a substantial difference, for
instance, between C3 and C4 plants (Smith &
Epstein, 1971), and a well-founded biophysical
understanding of fractionation during C-fixation in
C3 plants (Farquhar, O’Leary & Berry, 1982). It is
clear, however, that variability in δ"&N is not random.
The challenge is, therefore, to interpret δ"&N signa-
tures. It should perhaps be emphasized that even
%N, which is reported on a routine basis, is also a
ratio (N content}mass of sample), and is not always
a straightforward diagnosis of plant N nutritional
status. A certain %N might, for example, imply
either N-deficiency or N-sufficiency (e.g. Timmer &
Armstrong, 1987). Thus, %N could sometimes be
argued to be of no more value than δ"&N, apart from
the fact that in the case of %N we can base our
interpretations on a much larger body of exper-
imental and empiricial evidence. Ten years ago there
was a remarkable difference in technical complexity
between analysing total-N and analysing δ"&N. Now,
the simplicity of use of modern CF-IRMS means the
user will rapidly obtain large data sets on %N and
δ"&N, but is left with major problems of interpret-
ation.
In the near future we will probably see an
intensified discussion about the role of external
source effects vs. internal fractionations as causes of
variability in δ"&N in plants. The way ahead is to
198 P. HoX gberg
conduct careful experiments in the field and the
laboratory (where complete isotopic mass balances
are possible, and should be reported), and to combine
δ"&N approaches with δ"&N tracer and non-isotopic
methods in these experiments. Multiple stable
isotope approaches are especially useful, as they can
help to disentangle isotopic source effects from
fractionations within the system. Interpretation will
not be possible without modelling. In particular,
there is a need to develop models which can carry out
interactive comparisons between data sets derived
from contrasting approaches. Moreover, methods to
analyse δ"&N in small and dynamic soil N pools need
to be further developed, if models and interpre-
tations are to be critically tested.
I would like to thank the editors of The New Phytologist for
the invitation to write this review, and for their patience
when it was delayed. Present and past members of my
group (especially Mona Ho$ gberg, Lars Ho$ gbom, Christian
Johannisson, Helga Schinkel and Ha/ kan Wallmark) have
given many helpful contributions and allowed me to use
their unpublished data. Reviewers of this paper and
previous papers have given important comments. Ian
Alexander, Charles Garten, Gerhard Gebauer, Linda
Handley, Tony Haystead, Carol Kendall, Knute
Nadelhoffer, Torgny Na$ sholm, Ernst-Detlef Schulze and
George Stewart have shared their knowledge and some-
times given access to unpublished data. Dave Myrold
provided important insights into modelling. My work has
been funded by the Swedish Research Council for Forestry
and Agriculture, the Swedish Environmental Research
Agency, the EEC (project NiPhys, grant no: EV5V-
CT92–0433), the Swedish Natural Sciences Research
Council, the Swedish Agency for Research Cooperation
with Developing Countries, and the Swedish Forestry
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