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

Research Foundation.

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