Land Use Change and Mountain Biodiversity - Chapter 7 potx

89

7

Fire, Plant Species Richness,
and Aerial Biomass
Distribution in Mountain
Grasslands of Northwest
Argentina

Roxana Aragón, Julietta Carilla, and Luciana Cristóbal

INTRODUCTION

Grazing and fire are the most common distur-
bances in many grassland ecosystems around
the world (McNaughton et al. 1993; Vogl 1974
in Oesterheld et al. 1999, De Baro et al. 1998),
and they both affect biodiversity and plant com-
munity dynamics. Grazing and fire influence
species composition and richness, determine
dominant life-forms and therefore the general
structure of the community (Belsky 1992; Diaz
et al. 1992; Milchunas and Lauenroth 1993;
Collins et al. 1998). They can also regulate eco-
system processes such as nutrient cycling
(Hobbs et al. 1991) and plant productivity
(McNaughton 1985; Rusch and Oesterheld
1997). Importantly, grazing and fire often occur
together, and they interact deeply.

Grazing and fire are both consumers of
plant production. Herbivores feeding on forage
can determine the fuel load. Fire, in turn, con-
sumes accumulated biomass that could be used
by herbivores (Oesterheld et al. 1999). Grazing
can influence fire frequency and intensity, and
fire determines what is left for herbivores, not
only in terms of quantity but also in terms of
forage quality (Hobbs et al. 1991). In addition,
these disturbances provide open space for col-
onization that, in turn, can modify species
diversity, promote the establishment of certain
species, and change the general structure of the
community (Collins 1987; Pucheta et al. 1998;
Valone and Kelt 1999). Grazing and fire occur
naturally in many grasslands and savannas, and
they also are part of many management prac-
tices. In addition, burning in grasslands and
savannas has an important worldwide effect
because it is one of the major sources of atmo-
spheric methane and CO

2

, especially in tropical
areas (Crutzen et al. 1985 in Hobbs et al. 1991).
Livestock raising is one of the most impor-
tant land uses in many montane grasslands
(Eckholm 1975). Particularly in Andean grass-
lands, extensive cattle grazing is often com-

bined with burning of the natural vegetation
(Schmidt and Verweij 1992 in Hofstede et al.
1995, Grau and Brown 2000). Fire promotes
resprouting and is believed to encourage the
development of more palatable life-forms (Grau
and Brown 2000). However, grazing and fire
can also increase soil susceptibility to erosion,
reduce species or functional richness (Lloret
and Vila 2003), and modify community com-
position (Pucheta et al. 1998; Diaz et al. 1992).
Eventually, their positive or negative effects
depend on an array of factors such as grazing
intensity, fire frequency, and climate.
Mountain grasslands are one of the most
species-rich habitats of northwest Argentina.
They are important in regulating the hydric
regime and in providing economic resources
(e.g. cattle ranching and scenic values). In spite
of their ecological and economical importance,

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90

Land Use Change and Mountain Biodiversity

mountain grasslands are scarcely represented in
the protected areas of Argentina, and little is
known about their functioning. The study site

of this work, the valley of Los Toldos, is located
in the upper Bermejo River basin and is con-
sidered an area of high conservation priority at
a national level (Brown et al. 2001). The dom-
inant land use is for grazing by cattle, and this
is combined with periodic fires. As was
observed in other neotropical mountains, recent
works suggest a decrease in land use intensity
in this area (Grau et al.

submitted)

. The reduc-
tion in the density of animals may produce
changes in fire frequency and intensity that can,
in turn, affect plant communities in different
ways. In this chapter, we describe a study on
how fires affect vegetation structure in the
mountain grasslands of northwest Argentina
that are used for grazing. More specifically, this
study intends to investigate the effect that the
time since the last fire event may have on plant
species richness, vegetation structure, and bio-
mass dynamics.

METHODS
S

TUDY

A

REA

The study was performed at the valley of Los
Toldos (22

°

30 S, 64

°

50 W), Santa Victoria,
Salta, Argentina. The study area consists of a
mosaic of mountain grasslands and

Alnus
acuminata

forest patches at an altitude of about
1700 masl. This area lies in the upper altitudinal
level of the phytogeographic province of the
Argentinean Yungas (subtropical montane for-
est) (Cabrera 1976). The original vegetation
seems to have been dominated by forest
patches, but a long history of grazing in the
valley probably shaped the current vegetation
physiognomy (Malizia 2003). The mean annual
temperature is 15

°

C, and the average precipita-
tion is 1300 mm (Ramadori 1995). The precip-
itation is highly seasonal, with most of the rain
falling during the summer months (Bianchi
1981).
The main disturbances at this altitudinal
range are grazing, fire, and landslides (Grau,
2005), and livestock raising is the most com-
mon land use. Cattle grazing is extensive with
no fences limiting individual properties.
There are no data on grazing intensity in Los
Toldos, but information provided by national
agricultural censuses for Santa Victoria shows
a decrease in the population of domestic ani-
mals during the 20th century (Grau et al. sub-
mitted). These data, together with information
provided by local people, suggest that grazing
intensity in Los Toldos is currently low
(between 0.5 and 1 cow per 10 hectares). The
pastoral system involves transhumance, a sea-
sonal movement of cattle from the highlands
to midaltitude and piedmont forests (Grau and
Brown 2000). Cattle are driven up to the high-
land grasslands at the beginning of the sum-
mer period and, in March, they are brought
back to lower ranges (piedmont forest). Dur-
ing the summer period (from November to

March), the animals feed mainly in grassland
patches, but also browse in the

Alnus

forest
understory. Summer grazing by cattle is usu-
ally prepared for by burning the vegetation in
spring. The extent and frequency of burning
seem to depend on the proximity to settle-
ments and on the weather conditions (wind,
temperature, and soil humidity) when the fire
is started. As a result of these management
practices, the landscape consists of a mosaic
of vegetation patches, differing in the time
since the last burning event occurred.

S

AMPLING

D

ESIGN



AND

D

ATA

A

NALYSIS

In November 2000, we conducted a survey in
the study site, looking for evidence of previous
fire events. Based on this survey and on infor-
mation provided by local inhabitants, we iden-
tified three types of vegetation patches that dif-
fered in the time since the last fire event
occurred:
1. Areas burned during the ongoing
growing season (the last fire event
probably occurred during spring
2000). These areas showed evident
signs of fire, such as abundant char-
coal, ashes, and burned vegetation.
2. Areas burned during the previous
growing season (spring 1999), with

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Fire, Plant Species Richness and Biomass in Mountain Grasslands of NW Argentina

91

some evidence of fire (mainly the
remains of charcoal).
3. Areas not burned recently. In this
case, the last fire event apparently
took place at least 5 years ago (spring
1995 or earlier). This information
was checked with local residents.
We selected three patches of each vegeta-
tion type (nine in total). The time since the last
fire event was regarded as “treatment.” Hereaf-
ter, we will refer to the different treatments as:
<1 year (areas burned during the ongoing grow-
ing season); >1 year (areas burned during the
previous season), and >5 years (areas not
burned for at least 5 years). Unfortunately, since
burning is a common practice in this area, we
did not have any plots that had no fire and could
have served as a control. All the vegetation
patches included in our sample were no more
than 3 km apart from each other, had areas of
less than 500 m

2

, and were in similar topo-
graphic positions. Because of the absence of
fences, vegetation patches had potentially sim-
ilar grazing pressure.
In December 2000, we conducted plant
relevés in 1 m

×

1 m plots with five plots per
patch. The plots were placed every 10 m in a
50-m transect. The transects were placed at ran-
dom in each patch (i.e. 1 transect in each patch).
Each plot was divided into four 0.5 m

×

0.5 m
quadrats, and all the plant species present were
recorded. In addition, we collected all the aerial
biomass in ten 0.2 m

×

0.2 m plots in each
patch. Whenever possible, the plots were placed
in two 50-m transects that were separated by
10 m. If the vegetation patches were not big
enough, we placed plots in shorter transects, but
always used the same number of plots. We col-
lected biomass in December 2000 and in Janu-
ary, February, March, and August 2001 (before
the next burning event). The biomass was clas-
sified into live biomass, standing dead, and lit-
ter. Live biomass was further classified for dif-
ferent life-forms (i.e. graminoids, tussock

grasses, erect species, rosettes, prostrate spe-
cies, ferns, and woody species). All the material
was classified, dried to constant weight at 70

°

C,
and weighted.
We used ANOVA tests for the comparisons
between treatments (both for total biomass and
proportions). In the case of species richness and
total number of species, we used Kruskal–Wal-
lis tests, a nonparametric technique, because the
assumptions required for parametric tests were
not met. Differences in the biomass collected
throughout the year were tested through
repeated measures ANOVA. Species frequency
was computed as the number of plots per treat-
ment in which each species was recorded.
Equativity was computed as:

E

m

= –

Σ

(

p

i

log

p

i

)/log N
where

p

i

is the proportion of the species
recorded in transect

m

that belong to life-
form

i

, and

N

is the total number of different
life-forms in that transect. Small values of

E

imply that one or a few life-forms are dominant
in the community; in other words, this index
is an indication of evenness (O’Neill et al.
1988). To measure compositional similarity
among plots, we performed a detrended corre-
spondence analysis (DCA) with downweight-
ing of rare species. Only the species that were
recorded in at least two plots were considered.
We computed a nonparametric Kendall’s tau
correlation between plot scores in the ordina-
tion space and the time since the last fire event.
The analyses were performed in Statistica
(StatSoft 1993) and PCORD (McCune and
Mefford 1997).

RESULTS

We recorded a total of 149 species in the study
area. In Table 7.1, we have included only the
species that had a frequency 0.4 in at least one
of the treatments. Among these 45 species, 32
were common to all the treatments. The number
of species per square meter did not differ

between the treatments (Kruskal-Wallis test,
KW 3.29,

p

= .19) (Table 7.1), and the most
common species were present in all the patches
independently of their fire history. The most
common grasses were

Elionurus muticus

and

Paspalum notatum; Stevia yaconensis

was the
most frequent woody species, and the

Cuphea

sp. was the dominant prostrate species.
The ordination of plots in the DCA was not
clearly linked to the treatments. The first axis

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92

Land Use Change and Mountain Biodiversity

of the ordination explained approximately 30%
of the overall variance in species data

λ

1

=
0.259, total inertia = 0.859), and plot scores
were not significantly correlated with the time
since the last fire event (Kendall’s tau = 0.48

p

= .07). However, there seemed to be some
minor changes in species composition in
response to the treatments because we found
that some species were differentially recorded
in certain plots.

Anemone decapetala

and

Tes-
saria fastigiata

were recorded only in plots that

were recently burned.

Eupatorium bupleurifo-
lium

and

Ophioglossum

sp. were more abundant
in the >5-year treatment, whereas

Baccharis
tridentata

and

Setaria

sp. were predominantly
recorded in <1-year treatment (Table 7.1). In
addition, the equativity of life-forms showed a
slight tendency to decrease in the patches that
were not burned for 5 years (KW 5.42,

p

= .06)
(Table 7.1). The decrease in equativity in areas
that were not burned for 5 years was related to

the increasing dominance of woody species in
comparison to other life-forms that were less
frequently recorded, such as erect species and
rosettes.
The total aerial biomass was significantly
higher in the patches that were not recently
burned (F = 69.59,

p

< .001). The biomass in
the >1-year treatment was almost twice as high
as the biomass in the <1-year treatment, and the
biomass in the >5-year treatment was more than
3 times the biomass in <1-year treatment
(431.48 ± 27.95, 738.09 ± 63.64, and 1303.25
± 58.79 g m

−

2

for <1-year, >1-year, and >5-year
treatments, respectively, Figure 7.1). There was
no difference between <1-year and >1-year
treatments with respect to total live biomass,
but live biomass was highest in the >5-year
plots (F = 22.02,

p

< .01). There was no differ-
ence in total standing dead material (F = 3.33,

p

= .10), but the total amount of litter differed
between the treatments (F = 71.17,

p

<.001).
Patches that were burned in the ongoing grow-
ing season (<1-year) had considerably less litter
than both >1-year and >5-year patches (Figure
7.1).
The relative contribution of the different
biomass categories differed between the treat-
ments. Live biomass had a high contribution to
the total biomass in the patches that were
burned during the ongoing growing season (<1-
yr) (60 ± 0.5%), whereas the proportion of litter
was minimum in this treatment (15, 27, and
41% in <1-year, >1-year, and > 5-year treat-
ments, respectively) (F = 5.14,

p

= .04 for
ANOVA on live biomass and F = 73.33,

p

<
0.001 for ANOVA on litter) (Figure 7.2). The
contribution of standing dead material was
reduced in the patches that were not burned for
5 years (22, 29, and 10%, respectively) (F =
9.77,

p

< .01).
The proportion of live biomass differed
between <1-year and >1-year treatments, but
there was no difference between these two treat-
ments and the >5-year treatment. But impor-
tantly, although the overall proportion of live
biomass was similar between the <1-year and
>5-year treatments, the relative contribution of
the different life-forms to the total of live bio-
mass was quite distinct. Patches that were
recently burned (<1 year) had a high proportion
of erect species and ferns compared to the other
treatments (Table 7.2). The proportion of tus-
sock grasses plus graminoids did not differ
between <1-year and >1-year treatments but
their contribution was significantly smaller in
the patches that were not burned for 5 years
(28 and 33% in <1-year and >1-year and 15%

in >5-year treatments) (Table 7.2). This differ-
ence was mainly due to tussock grasses that
were reduced in >5-year patches (23, 25, and
9%, respectively, in the <1-year, >1-year, and
>5-year treatments). The biomass of grami-
noids was similar in all three treatments. Woody
species accounted for 72 ± 9% of the live bio-
mass in the >5-year treatment (Table 7.2).
The seasonal dynamics of the total live bio-
mass, standing dead, and litter showed some
similarities between the treatments. Biomass
assigned to the standing dead compartment
showed a peak in August in all three treatments
(Table 7.3). Similarly, litter had its maximum
value in August in the <1-year and >5-year
treatments, but we did not detect any seasonal
trend in the >1-year patches. Live biomass
showed a significant decrease in August in the
>1-year patches, and a small peak in March and
December; however, no similar trend was
detected in the other two treatments (Table 7.3).
Interestingly, the relative contribution of live
biomass throughout the year strongly differed
between the treatments. Patches that were

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Fire, Plant Species Richness and Biomass in Mountain Grasslands of NW Argentina

93

TABLE 7.1



Species with frequencies

≥

0.4 in at least one of the different treatments

a

<1 year >1 year >5 years

p

Values

b

Mean number of species per m

2

25.20 25.67 20.27 .19
Total number of species 89 96 83 .11
Equativity of life-forms 0.88 0.93 0.77 .06

Species Life-forms <1 year >1 year >5 years

Achyrocline

sp. Rosette 0 0.40 0.20

Agrostis

sp. Graminoid 0.13 0.40 0.07

Anemone decapetala

Prostrate 0.40 0 0

Baccharis coridifolia

Woody 0.73 0.73 0.53

Baccharis rupestris

Woody 0.60 0.60 0.27

Baccharis tridentata

Woody 0.73 0.40 0.13

Chaptalia modesta

Rosette 0.53 0.20 0.2

Chevreulia acuminata

Rosette 0.20 0.53 0.40

Clitoria cordobensis

Prostrate 0.67 0.07 0.13

Croton

sp. Woody 0 0.33 0.47

Cuphea

sp. Prostrate 1 1 1

Cynodon

sp. Graminoid 0.33 0.67 0.13

Cyperaceae

Graminoid 0.40 0 0.20

Desmodium affine

Prostrate 0.40 0.73 0.67

Desmodium

sp. Prostrate 0.40 0.40 0.80

Elionurus muticus

Tussock grass 1 1 0.67

Eupatorium bupleurifolium

Woody 0.13 0.40 0.93
Euphorbiaceae Prostrate 0.87 0.13 0.27

Gamochaeta

sp. Rosette 0.40 0 0.27

Hybanthus parviflorus

Erect 0.27 0.67 0.20

Hypericum

sp. Erect 0.40 0.53 0.07

Hyptis mutabilis

Woody 0.20 0.40 0.60

Juncus

sp. Graminoid 0.67 0.40 0.20

Unknown 3 (Myrtaceae)

Woody 0.20 0.40 0.27

Lepechinia vesiculosa

Woody 0 0.80 0.67

Lobelia nana

Prostrate 0.67 0.53 0.67
Malvaceae Prostrate 0.33 0.40 0.33

Ophioglossum

sp. Fern 0 0.26 0.47

Panicum ovuliferum

Graminoid 0 0.47 0.33

Panicum

sp. Graminoid 0 0.60 0.33

Paspalum notatum

Graminoid 0.53 0.47 0.47

Paspalum

sp. Tussock grass 0.87 0.47 0.67

Plantago

sp. Woody 0.60 0.07 0.20

Polygala pulchella

Erect 0.60 0.20 0

Polygala

sp. Erect 0.20 0.40 0

Pteridium aquilinum

Fern 0.47 0.27 0.07

Ranunculus praemorsus

Erect 0.20 0.40 0.20

Richardia

sp. Erect 0.53 0.13 0.13

Setaria

sp. Graminoid 0.53 0.07 0

Stenandrium dulce

Rosette 0.13 0.53 0.20

Stevia alpina

Woody 0.80 0.67 0.13

Stevia yaconensis

Woody 1 1 0.93

Tagetes filifolia

Prostrate 0.40 0.20 0.07

Tessaria fastigiata

Woody 0.40 0 0

Tibouchina

sp. Woody 0.13 0.20 0.40

a

<1 year: patches that were burned in the ongoing growing season, >1 year: burned the previous season, or >5 years:
not burned for 5 years; mean number of species per m

2

. Total number of species and equativity of life-forms per
treatment is also shown.

b

p

values correspond to Kruskal–Wallis tests.

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94

Land Use Change and Mountain Biodiversity

FIGURE 7.1

Variation of biomass according to different treatments. Live biomass, standing dead, litter, and
total biomass (g m

–2

) in patches that were burned in the ongoing growing season (<1 year), burned in the
previous season (>1 year), or not burned for 5 years (>5 years). Different letters denote significant differences
at

p

< .05 according to an ANOVA test, and ns indicates no significant differences. The comparisons were
made between treatments within each biomass category.

FIGURE 7.2

Relative contribution of live biomass, standing dead, and litter to the total biomass in the different
treatments: patches that were burned in the ongoing growing season (<1 year), in the previous season (>1 year),
or not burned for 5 years (>5 years). Different letters stand for significant differences at

p

< .05 according to
an ANOVA test. The comparisons were made between treatments within each biomass category.
a
a
b
ns
ns
ns
a
b
c
a
b
c
0
200
400

600
800
100 0
120 0
140 0
160 0
< 1 > 1 > 5 y e ar s
Time since last fire event
(y
ears
)

Dry weight (gm
-2
)
Live biomass
Standing dead
Litter
Total biomass

c
b
a
b
a
a
cd
bd
ac
0

0. 2
0. 4
0. 6
0. 8
1
<1 >1 >5
Time since last fire event (years)
Biomass (proportion)
Live biomass
Standing dead
Litter

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Fire, Plant Species Richness and Biomass in Mountain Grasslands of NW Argentina

95

burned during the ongoing growing season
(<1 year) had 70 to 80% of their biomass as
live biomass during December and January
(Figure 7.3), whereas in the >1-year and >5-
year treatments, this proportion hardly
approached 50%. The contribution of live bio-
mass to the total was higher in the <1-year
patches almost throughout the year, which may
represent substantial changes in the seasonal
pattern of forage availability.

DISCUSSION

Time since the last fire event affected the total
aerial biomass, the proportion of live biomass,
standing dead, and litter, and the contribution
of the different life-forms, both in terms of bio-
mass and life-form frequency. Nevertheless,
species richness was similar among all treat-
ments, and species composition showed only
small variations. Our results differ from those
of Collins (1987) and Pucheta et al. (1998), who

TABLE 7.2
Relative contribution (mean and standard deviation) of the different life-forms to the
live biomass

Life-Forms <1 year

a

>1 year

a

>5 year

a

p

-Values

b

Erect species 0.18 ± 0.02 0.08 ± 0.04 0.07 ± 0.03 .06
Ferns 0.10 ± 0.09 0.03 ± 0.03 0.02 ± 0.01 .67
Graminoids 0.05 ± 0.01 0.09 ± 0.01 0.06 ± 0.06 .30
Prostrate species 0.05 ± 0.02 0.05 ± 0.02 0.03 ± 0.01 .39
Rosettes 0.01 ± 0.01 0.01 ± 0.01 0.01 ± 0.01 .20
Tussock grasses 0.23 ± 0.04 0.26 ± 0.03 0.09 ± 0.09 .06
Woody species 0.37 ± 0.09 0.49 ± 0.05 0.72 ± 0.09 .03

a

<1 year: patches that were burned in the ongoing growing season, >1: burned the previous season or >5 years: not
burned for 5 years.

b
p values correspond to Kruskal–Wallis tests.
TABLE 7.3
Seasonal dynamics of total live biomass, standing dead, and litter (mean and standard error)
a
Biomass
Compartment December January February March August F
b
p
<1 year
Live 207.64 ± 16.18 371.77 ± 95.92 209.05 ± 29.84 262.85 ± 41.31 185.11 ± 34.51 1.81 0.21
Standing dead 23.18 ± 5.22 85.87 ± 43.11 100.80 ± 37.90 98.09 ± 23.25 252.39 ± 37.77 6.12 0.01
Litter 32.40 ± 11.27 49.55 ± 19.27 73.15 ± 18.21 38.72 ± 7.03 167.80 ± 19.38 10.80 0.002

>1 year
Live 391.53 ± 66.42 288.44 ± 44.93 327.87 ± 25.26 445.95 ± 30.70 184.76 ± 50.86 13.22 0.001
Standing dead 196.02 ± 47.97 159.19 ± 20.78 172.43 ± 32.48 184.92 ± 59.82 462.35 ± 56.28 20.01 0.003
Litter 235.88 ± 80.12 175.20 ± 18.92 172.84 ± 20.62 209.41 ± 39.03 313.61 ± 62.52 1.38 0.32
>5 years
Live 666.15 ± 52.82 542.86 ± 55.39 557.21 ± 94.12 611.96 ± 46.36 662.14 ± 80.68 1.80 0.22
Standing dead 142.35 ± 38.75 84.80 ± 35.89 92.57± 29.44 110.06 ± 29.17 298.33 ± 117.10 3.90 0.04
Litter 603.98 ± 112.18 409.74 ± 10.33 549.87 ± 65.30 367.83 ± 82.96 816.37 ± 84.52 5.88 0.05
a
In areas that were burned in the ongoing growing season (<1 year), burned the previous season (>1 year), or not burned
in 5 years (>5 years).
b
Boldened cells indicate significant differences at p < .05.
3523_book.fm Page 95 Tuesday, November 22, 2005 11:23 AM
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96 Land Use Change and Mountain Biodiversity
found that disturbances such as grazing and fire
increased both species richness and diversity.
In many cases, the increment in the number of
species results from the colonization by exotic
species or from the predominance of small-
sized species, which are tolerant to disturbance
(Belsky 1992; Pucheta et al. 1998). In our study
site, we did not record exotic species, and
because all our patches have a long history of
grazing, most of these species may indeed be
tolerant to disturbances. Fire and grazing may
produce the same kind of selective pressure,
and they can both favor fast-growing or small-
sized species, especially tussock grasses and

annuals. For these reasons, fire suppression may
not cause compositional changes in areas such
as our study site, where grazing occurs simul-
taneously.
Although species richness remained similar
in the different treatments, we detected changes
in the life-form spectrum and in the distribution
of aerial biomass. The reduction in functional
or species diversity as a consequence of a
decrease in disturbance frequency has been
observed in many cases (e.g. Pucheta et al.
1998; Valone and Kelt 1999) and is often attrib-
uted to a strengthening in species competition.
In the grasslands of Los Toldos, fire suppression
caused an increase in the dominance of woody
species; many of these species were present in
burned plots, but they became more abundant
and of a bigger size in plots that were not
recently burned. Tussock grasses were favored
by fire, but their contribution was reduced in
>5-year treatment. This change in the domi-
nance of woody species alters site flammability
that might reduce fire frequency in these plots
in the future.
In addition to the changes in life-form con-
tribution, there was a change in the distribution
of aerial biomass. Fire reduced the total bio-
mass by more than two-thirds (1303 gm
–2
in

areas not burned for 5 years compared to 431
gm
–2
in areas that were recently burned), and
the amount of litter was reduced in a similar
way. This reduction in aboveground biomass
that is associated with changes in the contribu-
tion of different life-forms results in changes in
vegetation structure that may alter soil cover.
Modification of soil cover can, in turn, affect
FIGURE 7.3 Relative contribution of live biomass to the total biomass throughout the year in patches that
were burned in the ongoing growing season (<1 year), in the previous season (>1 year), or not burned for
5 years (>5 year).
0
0.2
0.4
0.6
0.8
1
Dec Jan Feb Mar Aug
Biomass (proportion)
<1 year
>1 year
>5 years
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Copyright © 2006 Taylor & Francis Group, LLC
Fire, Plant Species Richness and Biomass in Mountain Grasslands of NW Argentina 97
erosion hazards that may have further implica-
tions on the hydrology and nutrient dynamics
of the system (Hofstede et al. 1995). The dif-

ferential allocation of biomass to the distinct
biomass compartments and, especially, the vari-
ation in the amount of dead material that
reaches the soil, can alter the decomposition
rate and, consequently, the nutrient pools
(Hobbs et al. 1991). Unfortunately, due to the
lack of sound information, at present we can
only hypothesize about these effects in the
grasslands of Los Toldos. On the other hand,
the short-term effect of fire on forage availabil-
ity to herbivores in this site is easier to appre-
ciate.
Patches that were burned in springtime had
more than 70% of their total biomass as live
biomass in the following summer (December
and January). Therefore, fire modifies the sea-
sonal dynamics of aerial biomass and changes
forage availability at least for the summer
period, when livestock is brought up to these
mountain grasslands. The availability of green
forage, especially in the form of highly palat-
able grasses, is particularly important for cattle
after a period when they have had access only
to low-quality winter forage. This means that
cattle obtain, in proportion, more green biomass
per bite in the patches that were recently
burned. This can explain why these patches are
often preferred (Coppock and Detling 1986;
Hobbs et al. 1991). Consequently, the propor-
tion of live biomass can have important effects

on livestock energy budgets and determine their
local movements. Importantly, fire promotes
more palatable life-forms (grasses instead of
woody species), and this makes the effect of
fire even more meaningful to herbivores.
Our results indicate that changes in the fire
frequency strongly affect vegetation dynamics
in the montane grasslands of northwest Argen-
tina. However, it is worth pointing out some
limitations of this study. First, we were unable
to find areas that were not burned and, there-
fore, we lacked a true control for our treatments.
Our conclusions refer to the effects of a change
in the fire frequency from once a year to once
in 5 years. We do not know if there is a threshold
after which a reduction in fire frequency pro-
duces no further changes in plant communities,
so we cannot say if our >5-year treatment
patches represent a transitional or a steady state.
Second, our sample size was rather small, espe-
cially with regard to species composition. This
is why we gave major emphasis to the results
referred to biomass distribution, the variability
of which seems to have been sufficiently
accounted for by our samples. Third, an
assumption of the present study is that our sam-
ple patches experience a similar grazing pres-
sure. Even though there are no fences or other
obstacles, and livestock have free access to all
patches in the study area, which are also very

close to one another, animals, as mentioned ear-
lier, may prefer recently burned grasslands. As
a consequence, these patches may receive
higher grazing pressure. All these limitations
have to be taken into account when considering
our conclusions. To overcome these inherent
difficulties, we are currently carrying out a con-
trolled experimental study with a bigger sample
size in the Los Toldos grasslands, which aims
at separating the effects of fire and grazing. The
preliminary results of this new experimental
setup, which has been running for more than
2 years, seem to support the findings reported
here.
SUMMARY
Fire and grazing are the most common distur-
bances in the mountain grasslands of northwest
Argentina. They can affect species composition
and richness, determine dominant life-form,
and the general structure of the community.
This work aims to determine the effect of burn-
ing on species richness, vegetation structure,
and aerial biomass distribution in the grasslands
of northwest Argentina that are subjected to
grazing. We performed a comparative study at
Los Toldos, Salta, Argentina (22º30 S,
64º50 W) at 1700 masl and surveyed patches
that differed in the time since the last fire event.
We considered three treatments: patches that
were burned during the ongoing growing sea-

son (in spring 2000), burned the previous sea-
son, or not burned for at least 5 years. Treat-
ments did not cause differences in species
richness, and caused only small changes in spe-
cies composition. The equativity of life-forms
showed a tendency to decrease with fire sup-
pression, with woody species becoming more
3523_book.fm Page 97 Tuesday, November 22, 2005 11:23 AM
Copyright © 2006 Taylor & Francis Group, LLC
98 Land Use Change and Mountain Biodiversity
dominant in plots that were not recently burned.
Total biomass and the proportions of live bio-
mass, standing dead, and litter varied among
treatments. Fire caused a reduction in total bio-
mass, but increased the contribution of live bio-
mass and encouraged the development of more
palatable growth forms (mainly grasses).
Patches that were burned during the ongoing
growing season had 80% of their biomass as
live biomass in December and January. In these
months, livestock are moved from forests at
lower altitudinal levels to these highland grass-
lands. This modification in the seasonal dynam-
ics of aerial biomass may represent a substantial
change in the pattern of forage availability,
especially at this time of the year.
ACKNOWLEDGMENTS
We are grateful to phytogeography students of
Universidad Nacional de Tucumán for the assis-
tance during fieldwork. The manuscript bene-

fited from suggestions from three anonymous
reviewers and from colleagues from LIEY.
International Foundation for Science and Fun-
dación PROYUNGAS provided financial sup-
port for this study.
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