SAJB-00643; No of Pages 10
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South African Journal of Botany xx (2011) xxx – xxx
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Early post-fire plant succession in Peninsula Sandstone Fynbos: The first
three years after disturbance
M.C. Rutherford a,b,⁎, L.W. Powrie a , L.B. Husted a , R.C. Turner c
a
Applied Biodiversity Research Division, Kirstenbosch Research Centre, South African National Biodiversity Institute, Private Bag X7,
Claremont 7735, South Africa
b
Department of Botany and Zoology, Stellenbosch University, Private Bag X1, Matieland 7602, South Africa
c
School of Botany and Zoology, University of KwaZulu-Natal, Private Bag X01, Scottsville, Pietermaritzburg 3209, South Africa
Received 16 July 2010; accepted 10 February 2011
Abstract
The early post-fire plant succession in fynbos vegetation in the Mediterranean-type climate area of South Africa was studied. Relatively little
has been published on this early stage of plant succession in fynbos. Annual sampling over the first three post-fire years confirmed a steady, but
relatively slow increase in plant canopy cover of shrubs and graminoids (mainly Restionaceae), whereas cover of geophytes and other herbs
peaked in the first year and declined significantly, thereafter. Cover of annual plants increased each year, which may relate to the persistence of a
relatively open vegetation cover by the third year. The responses of reseeder and resprouter species of the Restionaceae to the post-fire
environment appeared to be habitat dependent. Cover of the reseeders increased rapidly in seep areas, but their recovery was distinctly delayed in
dryland areas outside the seeps. Re-establishment of the many reseeder Erica species appeared to be delayed until the second post-fire year. Seed
banks of these species were possibly negatively impacted by the fire, and required dispersal of seed from unburnt areas for recruitment. In contrast
to some current generalisations, species richness appeared to increase after the fire; less certainly from the first to the second year, but more
certainly from the second to the third year. Therefore, this study does not support a short-term monotonic decline in species richness after fire in
fynbos.
© 2011 SAAB. Published by Elsevier B.V. All rights reserved.
Keywords: Burn; Canopy cover; Diversity; Ephemeral; Mediterranean-type; Reseeder; Resprouter; South Africa
1. Introduction
One of the many fire-prone vegetation types of the world
(Bond and Keeley, 2005) is the fynbos of the Cape Floristic
Region — CFR (Bond and Van Wilgen, 1996). The region is
characterized by exceptional levels of plant diversity (Van der
Niet and Johnson, 2009) and endemism, that is unparalleled
globally (Cowling et al., 2009). Response of fynbos to fire is
complex, and unique compared to other biomes of southern
Africa in that fynbos composition can be drastically altered after
a single fire (Bond, 1997). Fire is thus a major disturbance
factor that shapes the structure and composition of fynbos.
Consequently, much of the research effort in fynbos has focused
on the effects of fire and on how fire drives succession (Cowling
et al., 1997). Studies include those on rates of recovery of
vegetation cover, temporary importance of fire ephemerals,
differing sensitivity of resprouter and obligate reseeder plants to
fire return intervals, and trends in plant diversity (Kruger, 1979;
Kruger and Bigalke, 1984; Bond and Van Wilgen, 1996;
Cowling et al., 1997).
Post-fire studies in fynbos have generally found a decline in
plant diversity in the long-term, typically over periods of longer
⁎ Corresponding author at: Applied Biodiversity Research Division, Kirstenbosch Research Centre, South African National Biodiversity Institute, Private Bag X7,
Claremont 7735, South Africa. Tel.: +27 217998702; fax: +27 217976903.
E-mail address: [email protected] (M.C. Rutherford).
0254-6299/$ - see front matter © 2011 SAAB. Published by Elsevier B.V. All rights reserved.
doi:10.1016/j.sajb.2011.02.002
Please cite this article as: Rutherford, M.C., et al., Early post-fire plant succession in Peninsula Sandstone Fynbos: The first three years after disturbance, S. Afr.
J. Bot. (2011), doi:10.1016/j.sajb.2011.02.002
2
M.C. Rutherford et al. / South African Journal of Botany xx (2011) xxx–xxx
than a decade (Cowling and Pierce, 1988; Cowling and Gxaba,
1990; Privett et al., 2001). This long-term decline in plant
diversity after fire appears to also apply in some Mediterraneantype shrublands on other continents (Kruger, 1983; Bond and
Van Wilgen, 1996). However, changes in plant diversity in
fynbos in the early post-fire years are uncertain. In particular,
studies that determine changes in species richness over the first
three post-fire years in fynbos on sandstone-derived or sand
substrates are few and contradictory. Hoffman et al. (1987)
found a significant increase in species richness in lowland sand
fynbos from the first to the third year after fire. Similarly,
Adamson (1935) reported almost a doubling of the number of
species in fynbos vegetation of the lower western Table
Mountain over the same post-fire period. By contrast, Cowling
and Pierce (1988) indicated a decline in species richness,
effectively from the second to third year after fire in coastal dune
fynbos in the Eastern Cape Province. Despite these uncertainties
regarding the early post-fire years in fynbos, generalizations
persist that there is a monotonic decline in the number of plant
species with increasing stand age after fire (Bond and Van
Wilgen, 1996), or after an undefined ‘immediate post-fire phase’
(Cowling et al., 1992).
This study investigated initial, short-term succession dynamics
of plants for a period of three years after fire in sandstone fynbos.
Apart from characterizing short-term responses of species and
species groups to fire, we also tested the claim that there is a
monotonic decline in species richness after fire in fynbos.
2. Study area
The study was conducted after a wildfire at Olifantsbos in the
Cape of Good Hope section of the Table Mountain National
Park (Fig. 1). The site is part of a sample area of the BIOTA
monitoring programme (Haarmeyer et al., 2010). The area falls
within Peninsula Sandstone Fynbos (Rebelo et al., 2006) on the
exceptionally flora-rich Cape Peninsula (Simmons and Cowling,
1996). The landscape of the study site takes the form of a broken
undulating plain with a number of large boulders. Altitude is
between 60 and 80 m a.m.s.l. Mean annual precipitation falls
mainly in winter and amounts to 573 mm, based on the rain station
of Klaasjagersberg that is 3 km from the site. Winter and summer
temperatures are ameliorated by the nearby Atlantic Ocean. The
area is extremely windy with a high frequency of gale force winds.
The very infertile soils are derived from sandstone of the Table
Mountain Group. Soil depth varies, but shallow lithosols and
surface rock are common. In the final year of study, vegetation
structure comprised a dominant lower stratum of 0.2 to 0.4 m tall
Restionaceae (e.g. Hypodiscus aristatus) and low shrubs (e.g.
Diastella divaricata). Scattered 2 to 3 m tall shrubs (Mimetes
fimbriifolius and Leucospermum conocarpodendron) were conspicuous but not common. A few seep areas, which have standing
water in winter but dry surface soil in summer, also occurred in the
study area. These seeps were dominated by Restionaceae plants,
mainly Elegia cuspidata, that were 0.8 to 0.9 m tall.
The vegetation of the area studied was burned by the wildfire
of April 2007, and had been burned 17 years previously, in
January 1991. Before this, fire histories differed. About half the
Fig. 1. Location of (1) study area in relation to other sandstone or sand fynbos
study sites in the vicinity that have been monitored annually for at least three
growth seasons since a fire. These are (2) western Table Mountain (Adamson,
1935), (3) Jakkalsrivier, Grabouw (Kruger and Bigalke, 1984), (4) Swartboskloof,
Stellenbosch (Van Wilgen and Forsyth, 1992), and (5) Pella Nature Reserve,
Atlantis (Hoffman et al., 1987).
area had burned five years earlier in February 1986 whereas the
other half had not been burned for at least the 16 years for which
reliable fire records had been kept for the area (G.G. Forsyth,
personal communication). The average age of vegetation in the
Table Mountain National Park is 13.5 years (Forsyth and Van
Wilgen, 2008) and 10 to 13 years for most of the larger areas
elsewhere in fynbos (Van Wilgen et al., 2010). The 2007 burn
was a headfire with flame lengths between two and three metres.
South-east winds were in excess of 40 km per hour (C. Dilgee,
personal communication). The fire was considered to be a clean
burn that killed all above ground plant parts except for most
of the large individuals of thick-barked M. fimbriifolius and
L. conocarpodendron.
3. Methods
Data were collected from sixteen 10 × 10 m sample plots laid
out randomly within the BIOTA sampling grid of 1 km2
(Haarmeyer et al., 2010). Thirteen plots fell in dryland areas and
three plots in seeps. The 13 dryland plots could be clearly
divided according to prior fire history: seven plots with a fiveyear prior fire return period (short fire interval) and six plots
with a period of more than 16 years (long fire interval) (G.G.
Forsyth, personal communication). Data recordings for the first
and second post-fire years were carried out in January 2008 and
Please cite this article as: Rutherford, M.C., et al., Early post-fire plant succession in Peninsula Sandstone Fynbos: The first three years after disturbance, S. Afr.
J. Bot. (2011), doi:10.1016/j.sajb.2011.02.002
M.C. Rutherford et al. / South African Journal of Botany xx (2011) xxx–xxx
January 2009 respectively. Recordings for the third post-fire
year were done in December 2009 and January 2010. This
approach of monitoring the same plots differs from that
commonly used to analyse post-fire vegetation changes in
fynbos, namely, comparing independent samples of differently
aged stands (e.g. Privett et al., 2001). Although the latter
approach is probably effective in longer term studies, this may
not apply in the shorter term where spatial heterogeneity may
confound detection of changes.
Canopy cover of each plant species was estimated. In addition,
a count of individual plants of each species was made, except for
graminoid forms (hereafter applied to include Restionaceae,
Poaceae and Cyperaceae) and a few other species in which plant
individuals were difficult to define. Most geophytic species
were not counted. Counts of plants in dense populations, with
sometimes over 1000 plants per 100 m2, were approximate.
Mature, surviving plants and seedlings of M. fimbriifolius and L.
conocarpodendron were recorded separately. Plant species were
identified at the Compton and Bolus Herbaria. Fynbos species
identification in the field over the first three post-fire years was
challenging, particularly in the first year. For example, in groups
such as Restionaceae, seedlings appear very different from adult
plants in their first year and only in the third year do plants reach
the typical adult morphology, i.e. without finely-branched culms
(Haaksma and Linder, 2000). Also, a number of resprouter
species could not be separately identified in the first post-fire year.
Two observers were used: one for the first post-fire year and the
other for the remaining two years. Since plant cover estimates can
vary according to observer (Milberg et al., 2008; Bergstedt et al.,
2009) due to different approaches or definitions (Fehmi, 2010),
the two observers, independently of each other, were asked to
estimate, in the same way each had done before, the plant canopy
cover of a range of species and growth forms on the study site.
Close correspondence between the two observers was found in
cover estimates for graminoid species. For other species, estimates
in the lower range of cover values diverged; consequently, a linear
regression was applied to calibrate the cover estimates of the one
observer against those of the other observer.
Data were not normally distributed and, therefore, the nonparametric, two-tailed Wilcoxon Signed-Ranks test for repeated
measurements was used (XLSTAT Version 2010.4.01 by
Addinsoft) for each species and species group. Differences
between areas, according to prior fire intervals, were analysed
using the non-parametric, two-tailed Mann–Whitney U-test for
independent samples.
4. Results
4.1. Plant growth forms and groups
The post-fire recovery of mean plant canopy cover was
relatively rapid, yet less than 50% cover was attained by the third
year (Fig. 2a). A non-significant decline in cover during the second
year was followed by a significant increase in the third year
(p b 0.01). Cover regeneration patterns differed according to growth
form. Changes in cover of shrubs and graminoids closely paralleled
that of total plant cover, also with significant increases in the third
3
a
b
Fig. 2. Post-fire changes in mean canopy cover of a) all plants and b) growth
forms. Bars indicate standard error.
year (p b 0.05 for shrubs, p b 0.01 for graminoids) (Fig. 2b). By
contrast, cover of herbs and geophytes peaked in the first year.
Subsequent decline in their cover was significant (p b 0.05 for herbs
in the second year, p b 0.01 for geophytes in second and third
years). Cover of annual plants increased significantly from 0.008%
in the first year to 0.25% in the third (p b 0.01). Annual plants were
rare, and comprised only 3% of the species recorded.
Patterns of post-fire development of canopy cover of main
fynbos families were distinctly different from one another. Change
in cover of Restionaceae closely followed that of graminoids
(Fig. 3) with a significant decline in the third year (p b 0.01).
However, cover of Proteaceae peaked in the first year, declined
rapidly and significantly (p b 0.01) in the second year, and appeared
to stabilise in the third year. By contrast, cover of the species of
Ericaceae showed a slow, but significant increase (p b 0.001 in the
third year), and reached a similar level to that of Proteaceae (Fig. 3).
The responses of reseeder and resprouter species of
Restionaceae to the post-fire environment appeared habitat
dependent. Canopy cover of reseeders increased rapidly in the
seep areas in the first year (Fig. 4b), but their recovery was
delayed in the dryland areas (Fig. 4a). Cover of resprouting
species showed no such delay in either habitat.
4.2. Plant species
A total of 206 plant species were recorded in the study. Mean
species richness was observed to increase steadily after fire
(Fig. 5). Mean richness increased significantly between the
Please cite this article as: Rutherford, M.C., et al., Early post-fire plant succession in Peninsula Sandstone Fynbos: The first three years after disturbance, S. Afr.
J. Bot. (2011), doi:10.1016/j.sajb.2011.02.002
4
M.C. Rutherford et al. / South African Journal of Botany xx (2011) xxx–xxx
Fig. 5. Post-fire changes in mean species richness. Bars indicate standard error.
Fig. 3. Post-fire changes in mean canopy cover for three typical families of
fynbos vegetation. Bars indicate standard error.
second and third years (p b 0.001). It is known that richness for
the first year was an underestimate (see Discussion), which
therefore renders statistical significance of the change in
richness from the first to second years questionable. Erica
species constituted 2% of mean species richness in the first year,
and rose to 10% by the third year. Only in two seep plots did
species richness start to decline in the third year (each by less
than 15%).
The more frequent plant species (those that occurred in more
than two thirds of the sample plots — Table 1) are listed below
according to their patterns of response to the post-fire
conditions. Species that started to establish in the first year
and that increased significantly in the second or third years were
a
b
Fig. 4. Post-fire changes in mean canopy cover of reseeder and resprouter
Restionaceae in a) dryland and b) seep areas. Four species of Restionaceae could
not be assigned to a regeneration type but collectively constituted an average of
only 0.07% canopy cover and were excluded from the analysis. Bars indicate
standard error.
Erica exleeana (p b 0.05 for cover and density in the second
year — Tables 1 and 2), Ischyrolepis capensis (p b 0.01 for
cover in the second year), Lobelia setacea (p b 0.01 for cover
and density in the second year), Microdon dubius (p b 0.05 for
cover in the third year and density in the second year) and
Syncarpha vestita (p b 0.05 for cover in the third year). There
was a significant increase in density of Adenandra villosa and in
cover of Isolepis marginata only from the first to the third years
(p b 0.05 for both species). Species that showed no significant
differences in cover or density (p b 0.05) from the first year
onwards were Indigofera glomerata, Osteospermum polygaloides, Roella triflora, Tetraria bromoides and Ursinia
paleacea.
The typical fire ephemeral response was best shown by
Itasina filifolia which was observed to produce a large number
of seedlings in the first year, with mean plant density decreasing
significantly by two orders of magnitude and cover by 99% in
the second year (p b 0.01 for density and cover). In the third year
both density and cover remained at low levels. Other fire
ephemerals that declined significantly in the second and/or third
years were Corymbium africanum (p b 0.05 for cover in the
second year), Pseudoselago spuria (p b 0.01 for cover in the
third year and for density in the second and third years) and
Tritoniopsis dodii (p b 0.05 for density in the second year).
A species that was also observed to produce seedlings most
prominently in the first year was D. divaricata (density N 1000
per 100 m2 Table 2) but such did not follow the pattern of fire
ephemerals. The significant decline in cover (p b 0.05) and
density (p b 0.01) of D. divaricata in the second year was
followed in the third year by a significant increase in cover
(p b 0.05) while density remained at a low level. The decline in
density of R. decurrens was of intermediate magnitude
(p b 0.05) with no significant decrease in cover. A similar
pattern was found in Leucadendron laureolum which after the
initial flush in the first year density and cover decreased
significantly in the second year (p b 0.01 for density, p b 0.05 for
cover), with density in the third year falling significantly again
(p b 0.01) while cover tended to increase. Despite insignificant
declines in density in Metalasia densa, cover increased
significantly in the third year (p b 0.05).
The large flush of seedlings (density N 2000 per 100 m2) of
R. decurrens in the first year was followed by a significant
decline in the second year (p b 0.05) but with no concomitant
decline in cover. There was no significant change in density and
Please cite this article as: Rutherford, M.C., et al., Early post-fire plant succession in Peninsula Sandstone Fynbos: The first three years after disturbance, S. Afr.
J. Bot. (2011), doi:10.1016/j.sajb.2011.02.002
M.C. Rutherford et al. / South African Journal of Botany xx (2011) xxx–xxx
5
Table 1
Post-fire response patterns in canopy cover of species which occurred in more than two-thirds of the sample plots.
Species
Canopy cover (%)
Year 1
Shrubs
Adenandra villosa
Aspalathus retroflexa
Diastella divaricata subsp. divaricata
Edmondia sesamoides
Erica exleeana
Erica imbricata
Gnidia tomentosa
Indigofera glomerata
Leucadendron laureolum
Lobelia setacea
Metalasia compacta
Metalasia densa
Microdon dubius
Osteospermum polygaloides
Prismatocarpus sessilis
Pseudoselago spuria
Roella decurrens
Roella triflora
Stoebe cyathuloides
Struthiola dodecandra
Syncarpha vestita
Thesium capitellatum
Thesium strictum
Thesium sp.
Ursinia paleacea
Graminoids
Elegia stipularis
Hypodiscus aristatus
Ischyrolepis capensis
Ischyrolepis cincinnata
Isolepis marginata
Merxmuellera decora
Pentaschistis curvifolia
Staberoha cernua
Tetraria bromoides
Tetraria cuspidata
Tetraria microstachys
Thamnochortus cf. fruticosus
Thamnochortus lucens
Herbs
Corymbium africanum
Itasina filifolia
Geophyte
Tritoniopsis dodii
Remaining speciesb
Year 2
Relative mean
canopy covera
(%)
Year 3
Mean
S.E.
Mean
S.E.
Mean
S.E.
0.06
0.09
3.95
0.00
0.02
0.00
0.30
0.14
1.13
0.02
0.00
0.29
0.02
1.00
0.15
1.10
3.31
0.26
0.00
0.00
0.28
0.00
0.00
0.00
0.05
0.03
0.09
1.50
–
0.02
–
0.11
0.07
0.27
0.02
–
0.11
0.02
0.46
0.07
0.57
1.40
0.08
–
–
0.11
–
–
–
0.03
0.03
1.00
0.56
0.05
0.12
0.06
0.05
0.10
0.36
0.22
0.07
0.07
0.06
0.39
2.70
0.28
3.85
0.29
0.14
0.16
0.23
0.04
0.04
0.15
0.13
0.01
0.37
0.14
0.02
0.04
0.03
0.01
0.04
0.07
0.06
0.02
0.03
0.02
0.14
1.87
0.14
1.94
0.09
0.04
0.04
0.08
0.01
0.01
0.05
0.09
0.06
1.14
1.21
0.42
0.43
0.21
0.07
0.05
0.45
0.39
0.15
0.31
0.11
0.50
1.45
0.04
3.12
0.44
0.77
0.52
0.85
0.04
0.19
0.32
0.05
0.02
0.44
0.42
0.18
0.24
0.06
0.03
0.02
0.08
0.19
0.06
0.11
0.03
0.18
0.80
0.01
1.20
0.17
0.44
0.23
0.33
0.01
0.06
0.06
0.01
0.1
2.1
5.4
0.4
0.5
0.3
0.4
0.3
1.8
0.6
0.2
0.6
0.2
1.8
4.0
1.3
9.6
0.9
0.9
0.6
1.3
0.1
0.2
0.4
0.2
0.00
1.59
0.01
2.29
0.01
*
0.61
*
0.29
*
*
*
2.50
–
0.92
0.01
0.63
0.01
–
0.14
–
0.25
–
–
–
0.34
0.33
0.68
0.32
0.85
0.03
0.09
0.01
0.34
0.28
0.23
0.26
0.09
1.27
0.11
0.19
0.09
0.22
0.01
0.01
0.01
0.13
0.13
0.06
0.07
0.08
0.36
0.53
1.13
0.44
1.65
0.22
0.12
0.02
0.32
0.29
0.31
0.26
0.30
1.68
0.20
0.28
0.18
0.38
0.08
0.05
0.02
0.10
0.15
0.07
0.08
0.18
0.39
0.8
3.2
0.7
4.5
0.2
0.2
0.6
0.6
0.8
0.5
0.5
0.4
5.1
1.80
1.06
0.81
0.54
0.11
0.01
0.05
0.00
0.15
0.01
0.08
0.00
1.9
1.0
0.07
12.94
0.04
4.93
0.02
12.89
0.01
5.51
0.01
21.61
0.00
6.33
0.1
44.5
* Usually resprouter, but not detected in year 1. a Mean over three years. b Species occurring in fewer than two-thirds of sample plots.
cover in the third year. Similarly, the density of Gnidia
tomentosa declined in the second and third years (both p b 0.05)
without any significant change in cover.
A group of graminoid species showed similar trends but
cover of only Ischyrolepis cincinnata decreased significantly in
the second year (p b 0.01) then increased significantly in the
third (p b 0.01). Cover of Pentaschistis curvifolia and Thamnochortus lucens decreased significantly in the second year
(p b 0.01) but the increase in the third was not significant. The
decrease in cover of Hypodiscus aristatus in the second year
was not significant but the increase in the third was significant
(p b 0.01). Densities of graminoid species were not recorded.
Two species appeared to peak later than those above, namely
in the second year. Thus significant increases were found in the
second year in Aspalathus retroflexa (p b 0.01 for cover,
p b 0.05 for density) and in Prismatocarpus sessilis (p b 0.01
for cover) before density declined significantly in the third year
(p b 0.05 for both species).
Species that were slow to regenerate and started only in the
second post-fire year and that increased significantly from the
Please cite this article as: Rutherford, M.C., et al., Early post-fire plant succession in Peninsula Sandstone Fynbos: The first three years after disturbance, S. Afr.
J. Bot. (2011), doi:10.1016/j.sajb.2011.02.002
6
M.C. Rutherford et al. / South African Journal of Botany xx (2011) xxx–xxx
Table 2
Post-fire response patterns in plant density of selected species which occurred in more than two-thirds of the sample plots.
Species
Plant density per 100 m2
Year 1
Shrubs
Adenandra villosa
Aspalathus retroflexa
Diastella divaricata subsp. divaricata
Edmondia sesamoides
Erica exleeana
Erica imbricata
Gnidia tomentosa
Indigofera glomerata
Leucadendron laureolum
Lobelia setacea
Metalasia compacta
Metalasia densa
Microdon dubius
Osteospermum polygaloides
Prismatocarpus sessilis
Pseudoselago spuria
Roella decurrens
Roella triflora
Stoebe cyathuloides
Struthiola dodecandra
Syncarpha vestita
Thesium capitellatum
Thesium strictum
Thesium sp.
Ursinia paleacea
Herb
Itasina filifolia
Geophyte
Tritoniopsis dodii
Year 2
Year 3
Mean
S.E.
Mean
S.E.
Mean
S.E.
4.4
3.1
1248.1
0.0
31.3
0.0
318.8
45.6
266.1
3.4
0.0
232.8
1.9
553.4
56.9
126.9
2162.5
196.9
0.0
0.0
107.8
0.0
0.0
0.0
16.9
2.6
3.1
430.5
0.0
31.3
0.0
136.6
25.6
76.6
3.1
0.0
107.8
1.9
243.4
26.9
40.5
1224.8
78.1
0.0
0.0
43.5
0.0
0.0
0.0
8.9
7.5
15.1
102.3
6.5
222.6
225.8
59.8
12.7
55.7
31.7
49.2
27.1
8.1
60.3
66.3
32.0
367.2
83.1
49.2
129.9
43.2
9.0
2.7
13.2
20.4
3.0
4.4
23.2
2.3
93.3
87.2
18.6
4.2
8.5
8.9
16.2
9.4
2.3
16.9
50.2
13.2
132.4
21.7
16.4
37.3
14.8
3.1
0.9
4.2
12.3
7.4
7.4
80.6
18.2
126.8
281.0
27.1
4.3
38.7
26.8
44.3
42.5
7.4
76.7
0.6
3.6
340.1
97.9
52.3
236.1
74.0
5.4
12.8
30.5
12.1
2.7
2.3
26.5
8.1
85.6
94.1
10.6
1.4
6.3
10.0
18.3
19.4
2.1
33.0
0.4
1.2
124.3
38.5
14.4
79.9
39.9
1.6
3.7
12.3
4.5
724.7
376.4
2.6
1.3
5.0
3.0
23.1
12.3
3.3
0.9
1.8
0.6
second to the third year were Edmondia sesamoides (p b 0.01 for
cover, p b 0.05 for density), Erica imbricata (p b 0.01 for cover),
Struthiola dodecandra (p b 0.05 for cover) and Thesium
strictum (p b 0.01 for cover and density). Other species that
were also not found in the first year but did not increase
significantly in the third year were Elegia stipularis, Metalasia
compacta, Stoebe cyathuloides, Thesium capitellatum and
Thesium sp. Known sprouter species not recorded in the first
year are indicated in Table 1. Of these, the significant increase
in cover of Thamnochortus cf. fruticosus was marginal
(p = 0.044) in the third year.
Very few seedlings of M. fimbriifolius and L. conocarpodendron were recorded in the study.
The responses of very few species appeared to be linked to fire
history. Canopy cover of the annual, Helichrysum indicum, was
significantly higher on the short fire interval plots in the second
year (p b 0.01) and in the third year (p b 0.05) (Fig. 6a). In contrast,
cover of the reseeder shrub, D. divaricata, was greatly reduced on
these plots in the first post-fire year (p b 0.05) (Fig. 6b).
5. Discussion
Although the fire was regarded to be a clean burn, the very
low number of recorded seedlings of M. fimbriifolius and L.
conocarpodendron may indicate that the fire was not of a
particularly high intensity, according to observations by Bond et
al. (1990). Nevertheless, occurring in autumn, the burn would
have been expected to maximize seedling production of many
other species (Bond et al., 1984; Le Maitre, 1988b; Van Wilgen
et al., 1992). Moisture conditions were favourable for
germination and establishment of seedlings after the fire, as a
result of subsequent, regular occurrences of rain; also, the rain
in the first winter season was above average (Fig. 7).
Germination is likely to have substantially augmented the
flush of cover of resprouting fire-ephemerals in the first postfire year. Fire stimulated germination has been seen as an
evolutionary response to increased availability of the resources
of light, space and nutrients (Le Maitre and Midgley, 1992).
However, post-fire flushes of nutrients have been shown to be
of relatively short duration with availability of nitrogen
decreasing rapidly after nine months and phosphorous levels
returning to pre-fire levels after only four months (Cowling et
al., 1997).
Those species with a high density of establishing plants in
the first post-fire year experienced a sharp decline in the second
year. These include four fire ephemeral species found in this
study. Our results show that two other known fire ephemeral
species appear to peak later, not before the third year. S. vestita
Please cite this article as: Rutherford, M.C., et al., Early post-fire plant succession in Peninsula Sandstone Fynbos: The first three years after disturbance, S. Afr.
J. Bot. (2011), doi:10.1016/j.sajb.2011.02.002
M.C. Rutherford et al. / South African Journal of Botany xx (2011) xxx–xxx
a
b
Fig. 6. Patterns of response in canopy cover following the 2007 fire for
a) Helichrysum indicum and b) Diastella divaricata according to those areas
burnt previously in 1986 (short fire interval of five years) and areas not burned
previously for at least 16 years i.e. burnt at some time before 1975 (long fire
interval).
is viewed as a fire ephemeral (Brown, 1993) but in this study its
cover increased significantly in the third year. E. sesamoides,
which is prominent ‘in the few years after fire’ (Bean and Johns,
2005), was first recorded in the second year and its cover and
density increased significantly in the third year. The decline in
plant density was not always followed by a decline in canopy
cover. Thus, the mass mortality of D. divaricata seedlings in the
second year was followed by another significant reduction in
density in the third year while cover increased significantly, as a
Fig. 7. Monthly rainfall recorded at Klaasjagersberg (Rainfall Station 0004734 2,
which is 3 km from the study site) for the period of the study. The fire occurred
early in April 2007 before any rain had fallen in that month.
7
result of presumed rapid compensatory growth of the surviving
plants in the third year. The significant decrease in cover of the
graminoid Ischyrolepis cincinnata in the second year followed
by a significant increase in the third is difficult to interpret in the
absence of density data. Le Maitre (1987) suggested that a large
decline in plant density found in the second post-fire year in
granite fynbos was the result of density-dependent thinning
after establishment. By contrast, in a field study of Proteaceae,
drought was apparently found to be responsible for most seedling
deaths, especially in shallow soil over rock (Bond, 1984).
Relatively dry conditions may have contributed to seedling deaths
on the current study site, especially on the commonly found
shallow soil over rock. In the second year, rainfall in the first half
of the normal rainy season (April to July) was about 30% below
the mean rainfall for this period (Fig. 7). Possible effects of
herbivory on seeding mortality were not assessed in this study.
Bond (1984) established that predation was not a significant cause
of post-emergence seedling mortality in the Proteaceae that he
studied. However, in two species of this family, post-fire
herbivory on young seedlings was found to have a significantly
negative effect on seedling survival (Le Maitre, 1988a).
Most of the species of Erica were only recorded from the
second year onwards, which may suggest a delayed regeneration after fire (but see possible early detection limitations
below). All of these apparently later-appearing species of Erica
were obligate reseeders. Germination and establishment of
some Erica species elsewhere in fynbos have been observed to
be delayed until more than a year after fire (Kruger and Bigalke,
1984; Manders and Cunliffe, 1987), with one study reporting a
delay of more than two and a half years after fire (Adamson,
1935). Although Erica seed may be expected to be relatively
short-lived in the soil seed bank (Holmes and Newton, 2004),
this should not significantly affect the viable seed supply in the
first post-fire year. Germination of a wide range of Erica
species, stored in ambient room conditions, starts to decline
after about three years and after 10 years germinate very poorly
or not at all (A. Hitchcock personal communication). Ericaceae
have fine seeds (b 1 mg, le Maitre and Midgley, 1992) or very
fine seeds (b 0.1 mg, Bond et al., 1999; Holmes and Newton,
2004). Seedlings of fine-seeded species cannot emerge from soil
depths ≥ 20 mm and are likely to be killed by heat of a high
intensity fire at shallower depths (Bond et al., 1999). Most
ericoid species store their seed in the upper layers of soil
(Cowling et al., 1997) and are thus vulnerable to fire. Evidently,
storage at greater depths would not ensure survival. It has been
suggested that the delayed regeneration of Erica species after
fire may depend on transport of seed to the burned site, or on
other special germination requirements that are met only at a
later stage (Kruger and Bigalke, 1984; Manders and Cunliffe,
1987). The small size of seeds of most Erica species should
permit their relatively long-distance dispersal, especially in the
high-wind environment of the current study site. In addition, it
appears that some level of spatial pyrodiversity (Parr and
Andersen, 2006) may be necessary to allow transfer from fertile
sources to sterile areas. Spatial configurations of different fires
and size of the burnt areas (Manders and Cunliffe, 1987) can
critically affect such transfers.
Please cite this article as: Rutherford, M.C., et al., Early post-fire plant succession in Peninsula Sandstone Fynbos: The first three years after disturbance, S. Afr.
J. Bot. (2011), doi:10.1016/j.sajb.2011.02.002
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M.C. Rutherford et al. / South African Journal of Botany xx (2011) xxx–xxx
Other species with apparently delayed regeneration after fire
include the reseeder Restionaceae under dryland conditions,
i.e. outside seep areas. This delay may be a consequence of
competition. Sprouting species can have a competitive edge
over reseeders because of the ability of the former to re-establish
rapidly after fire (Vlok and Yeaton, 2000). By contrast, the
delayed regeneration found in the reseeder hemiparasite, T.
strictum, (Moore et al., 2010) may also result from a delay in the
recovery of the host plant(s) after fire (see Kazanis and
Arianoutsou, 2004). Moreover, the presumed limited mobility
of seeds of this species due to dispersal through myrmecochory,
may further delay post fire recovery (Johnson, 1992).
If germination of most fynbos annuals is triggered by smoke
or charred wood (Keeley and Bond, 1997) then it would be
expected that annuals might be largely confined to the first postfire year (see Kruger, 1979). However, the current study shows
increasing canopy cover of annuals from the first to third years
after fire. This may relate to the less inhibiting effect of the
relatively open vegetation canopy that remained below 50%
cover by the third post-fire year. This is in contrast to the
relatively high cover (Kruger and Bigalke, 1984) or near canopy
closure (Adamson, 1935) found on some other sandstone
fynbos sites by the third post-fire year. Indeed, canopy cover in
fynbos can reach 70 to 90% in only two years after fire (Kruger,
1983). Nevertheless, cover of annuals and the number of annual
species in this study remained relatively low, which is expected
in fynbos (Taylor, 1978; Kruger, 1979), even in the very open
areas after fire (Adamson, 1935). It has been suggested that the
rarity of fire annuals in fynbos is linked to limitation by low
levels of soil nutrients (Le Maitre and Midgley, 1992).
The role of smoke in fynbos has largely focussed on its use
as a germination cue for horticultural purposes (Brown and
Botha, 2004), yet its ecological role often appears speculative.
Germination of certain reseeder species, such as E. sesamoides,
is reported to be stimulated by smoke but appear to establish in
only the second year. It is possible that the timing of the possible
stimulatory effect of smoke may be long after the fire for some
species. The smoke cue may be long-lasting and the smokederived residue persists in the soil for a considerable time (Van
Staden et al., 2000) and hence may still be effective as a
germination stimulant some time after fire. However, the active
constituent of smoke is water-soluble and may be leached out
by winter rains (Van Staden et al., 2000). The ecological
significance of smoke remains unclear (Cowling et al., 1997).
The increase in species richness over the first three years
after fire agrees with the findings of Adamson (1935) and
Hoffman et al. (1987) and does not support the notion of a
monotonic decline in number of species over this period. Our
results also appear to bear out Kruger's (1983) suggestion of an
initial increase in fynbos species richness after the first post fire
year. Although the study of Privett et al. (2001) indicated a
steady declining trend in species richness from the first year for
several decades after fire, the data that they presented for the
first three years are too variable to support a decline over this
early post-fire period. Furthermore, Taylor (1984), whose data
were used almost exclusively for the first three post-fire years
by Privett et al. (2001), stated that species richness peaks at
between two and five years after fire. A study in granite fynbos
found that plant species richness remained at least relatively
constant over the first three post-fire years before gently
declining toward the sixth year after fire (Le Maitre, 1987).
We have previously indicated that the challenge in determining
fynbos species richness over the first three post-fire years is
detecting and identifying immature species, especially in the first
post-fire year. Indeed, Kruger (1983) cautioned that the findings of
apparent increasing richness in the first few post-fire years by
Adamson (1935) in fynbos and Posamentier et al. (1981) in
Australian heathland may be spurious owing to the difficulties in
detecting or identifying plants in the early stages of regeneration.
Musil and de Witt (1990) have also pointed to the difficulty in
distinguishing some fynbos species in the juvenile stage in the field.
Even under relatively favourable conditions, species richness is
potentially subject to errors of omission or identification at the
species level (Archaux et al., 2009). In the current study 20 known
resprouter species (out of the total of 206 species) were not detected
in the first year. Their addition would result in a corrected richness
of 46 per 100 m2 in the first year. This is very similar to the mean of
47 species per 100 m2 that was found for 30 replicate plots of one
year old fynbos after a fire on a sandstone site in the Swartberg
(Vlok and Yeaton, 1999). Although it is likely that many shortlived seedling individuals had disappeared by the time of sampling
in January (Cilliers et al., 2004), it is unknown whether any species
(including some annuals) may have been consequently missed on
our site. It would require the addition of more than 10 such species
to have been undetected in the first year for species richness in the
first year to surpass that found in the second year. However, if the
majority of very young establishing Erica species as well as some
other reseeder species had been undetected in the first post-fire
year, a temporary decline in species richness cannot be ruled out.
Nevertheless, from the second to third post-fire years, when species
were more readily identifiable, the increase found in species
richness remained significant. Therefore, a short-term monotonic
decline in species richness after fire in fynbos cannot be supported
by the current study. It remains an open question to what extent the
apparent long-term decline in fynbos species richness after fire
would be offset, were seed banks to be included (see Holmes and
Cowling, 1997).
Vlok and Yeaton (1999) have suggested that fire history prior
to the last fire may be important in determining diversity levels in
fynbos. However, we could not detect any trend in species
richness according to the prior short (five year) fire interval and the
prior long (more than 16 years) fire interval. It might be argued
that the significant positive association of the annual, Helichrysum
indicum, with the prior short fire interval area relates to annuals
benefitting from more open vegetation. It begs the question,
however, how this effect could have been sustained through a
subsequent period of 17 years under presumed closed canopy. The
negative association of the obligate reseeder, Diastella divaricata,
with the short fire interval areas does not seem adequately
explained by increased survival risk through immaturity. Thus, in
most five-year old post-fire populations of D. divaricata, at least
50% of plants would have flowered at least twice (A.G. Rebelo,
personal communication). However, even with juvenile periods of
only three to four years for some Proteaceae, it can take at least
Please cite this article as: Rutherford, M.C., et al., Early post-fire plant succession in Peninsula Sandstone Fynbos: The first three years after disturbance, S. Afr.
J. Bot. (2011), doi:10.1016/j.sajb.2011.02.002
M.C. Rutherford et al. / South African Journal of Botany xx (2011) xxx–xxx
another 10 years for adequate seed banks to build up (Van Wilgen
and Forsyth, 1992). If seed production of D. divaricata had indeed
been adversely affected by the short fire interval, this myrmecochorous species with very short dispersal distances (Cowling and
Gxaba, 1990; Johnson, 1992) could take many years to repopulate
an area. However, since fynbos normally requires four to five years
before a sufficient fuel load has accumulated to carry a fire (Kruger,
1983), the burn of the young five-year old vegetation may not have
been clean, and small pockets of unburned vegetation may have
remained as close sources of propagules. It is important to note that
any inferences from these prior fire interval areas suffer from
unavoidable pseudoreplication in that neither interdispersion of
samples nor independent treatment requirements of sampling
replication are met. This means that despite the very similar
physical conditions on both prior burn areas, other factors could be
responsible for, or contribute to, the differences.
The current study has helped quantify fire responses of
different plant forms and species in fynbos, in contrast to useful
yet purely descriptive accounts (for example, Trinder-Smith,
2006). It has also shown that claims of a short-term monotonic
decline in species richness after fire in fynbos are not supported.
One of the problems in comparing some results of different
post-fire studies in fynbos may be linked to the poor
repeatability of the results, given the great variation in
composition caused by fire. Post-fire research carried out over
30 years in the vicinity of the current study has shown highly
unstable species composition at local scale (Thuiller et al.,
2007) with, for example, L. laureolum becoming locally extinct
on nearly 30% of sites but colonizing a similar proportion of
new sites in the same period (Privett, et al., 2001). Conditions
following a fire are very complex and change rapidly (see Bond
and Van Wilgen, 1996). There is still much to be learned about
how these conditions are reflected in interactions and levels of
temporal segregation of species in the early stages after fire, and
what impact they may have on later successional development.
Acknowledgements
We thank Ruth Cozien, Khuze Mahlabegwane and Gwen Curry
for assistance during the fieldwork. The research comprised a sub
project within BIOTA Southern Africa Phase III incorporated in a
Research and Development Contract between the University of
Hamburg (UHH) and the South African National Biodiversity
Institute (SANBI) sponsored by German Federal Ministry of
Education and Research (Bundesministerium für Bildung und
Forschung (BMBF)) — Promotion number 01LC0624A2. We
thank SANParks for granting a research permit and providing
access to the Cape of Good Hope section of the Table Mountain
National Park.
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Edited by RI Yeaton
Please cite this article as: Rutherford, M.C., et al., Early post-fire plant succession in Peninsula Sandstone Fynbos: The first three years after disturbance, S. Afr.
J. Bot. (2011), doi:10.1016/j.sajb.2011.02.002