Journal of Ecology 2011, 99, 288–299
doi: 10.1111/j.1365-2745.2010.01737.x
Biotic and abiotic controls on tree colonization
in three early successional communities of
Chiloé Island, Chile
Marcela A. Bustamante-Sánchez1,2*, Juan J. Armesto1,2 and Charles B. Halpern3
1
Departamento de Ecologı´a, Center for Advanced Studies in Ecology and Biodiversity (CASEB), Pontificia Universidad
Católica de Chile, Casilla 114-D, Santiago, Chile; 2Institute of Ecology and Biodiversity (IEB), Casilla 653, Santiago,
Chile; and 3School of Forest Resources, University of Washington, Seattle, WA 98195-2100, USA
Summary
1. Most studies of tree regeneration are limited to particular environments and may not capture
variation in the biotic or abiotic factors that regulate recruitment at larger spatial scales. Critical
processes such as competition and facilitation can vary spatially, along gradients in resource availability and environmental stress, and temporally, with plant development.
2. We examined patterns of natural tree recruitment and experimentally followed germination and
seedling survival of five tree species (pioneer to late seral) in three early successional communities of
contrasting bio-physical environments in a rural landscape on Chiloé Island, Chile.
3. We quantified natural recruitment of juveniles and saplings and assessed relationships between
tree density and local environment. We used a removal experiment to test the influence of early successional vegetation on seed germination and early survival of tree species. In each community,
seeds and seedlings were placed in paired experimental plots from which vegetation was removed or
left intact (control). To identify potential correlates of germination and seedling survival, we measured light transmittance and soil properties in each plot.
4. In all communities, established vegetation had either a positive or neutral effect on germination
and ⁄ or survival although responses varied among life stages and species. Germination and survival
were correlated with the lower levels of light in controls, consistent with negative correlations
between natural tree densities and light. Vegetation cover was not dense enough to facilitate survival
of late successional species, but not too dense to inhibit survival of shade-intolerant or mid-tolerant
species. Among communities, natural densities of juveniles were greatest under conditions where
experimental germination rates were highest. Seedling height growth was lowest in the community
characterized by waterlogged soils, consistent with the naturally low transition rate from juveniles
to saplings and a negative correlation between density of shade-intolerant trees and soil moisture.
5. Synthesis. Our experiments indicate strong, mostly positive controls (facilitation) on tree recruitment in early seral shrublands with differing bio-physical environments. Benefits of shading are
manifested at different stages in the life history. However, community context is critical: variation
in seasonal patterns of soil moisture may explain spatial variation in the density and size structure
of natural tree recruitment.
Key-words: community assembly, ecological filters, facilitation, plant–plant interactions
southern temperate rain forest, species’ interactions, tree regeneration, vegetation heterogeneity
Introduction
Plant community assembly following disturbance is regulated
by abiotic and biotic filters that select potential colonists from
*Correspondence author. E-mail: [email protected]
the regional species pool (Weiher & Keddy 1995; Dı́az, Cabido
& Casanoves 1998). These filters can vary in strength in time
and space, selecting individuals from among the species and
life stages upon which they operate (Armesto & Pickett 1986;
Walker & Chapin 1987; De Steven 1991a; b; Gill & Marks
1991). Abiotic filters that affect seedling establishment include
2010 The Authors. Journal of Ecology 2010 British Ecological Society
Controls on tree colonization in shrublands 289
winteri J.R. et G. Foster, Winteraceae) on elevated surfaces,
such as woody detritus (Aravena et al. 2002). Our objectives
are threefold: (i) to quantify patterns of natural tree recruitment and environmental variation among three floristically
and structurally distinct early successional communities; (ii) to
determine whether early seral vegetation inhibits or promotes
seed germination and early survival of trees, and whether these
effects vary among communities and ⁄ or tree species with differing life histories and (iii) to isolate the abiotic factors (e.g.
light, soil moisture and nutrient availability) that could explain
variation in patterns of germination and early survival. For
shade-intolerant (pioneering) tree species, we hypothesized a
shift from neutral effects in open shrublands (Berberis community), to inhibitory effects in denser shrublands with lower light
availability (Baccharis community). In contrast, for shade-tolerant (late-seral) species, we hypothesized facilitative effects of
increasing strength from more open to more closed communities (Berberis to Baccharis), reflecting the beneficial effects of
shading (reduced light and temperature stress). Finally, we
hypothesized that for all species, rates of germination and
seedling survival would be lowest where soils are anoxic and
the potential for fungal infection is high due to seasonal waterlogging of soils (Baccharis community; Piper et al. 2008).
Materials and methods
STUDY AREA
The study was conducted within the rural landscape surrounding the
Senda Darwin Biological Station (SDBS), which lies 15 km northeast
of Ancud, northern Chiloé Island (42 S; Fig. 1). The vegetation in
this region is a mosaic of remnant fragments of Valdivian and NorthPatagonian evergreen forest (Veblen et al. 1996), grazed pastures and
small agricultural fields (Willson & Armesto 1996). The climate is
wet-temperate with a strong oceanic influence (Di Castri & Hajek
1976). Mean annual precipitation averages 2000–2500 mm; 20% falls
between December and March, during the growing season (austral
42°S
Senda Darwin
Biological Station
ic O
cea
n
Ancud
Argentina
Puerto
Montt
Pac
if
Castro
43°S
Chile
well-known resource and non-resource constraints such as
light, soil moisture, temperature and substrate availability.
Biotic filters include competition and facilitation, the processes by which plants interact for resources or modify their
environments in ways that reduce or enhance germination,
establishment and growth (Connell & Slatyer 1977; Callaway
& Walker 1997). The relative importance of these processes
can vary spatially with resource availability and environmental
stress (Pugnaire & Luque 2001; Tewksbury & Lloyd 2001;
Callaway et al. 2002; Kuijper, Nijhoff & Bakker 2004) or
temporally with succession or plant ontogeny (developmental
stage; Holmgren, Scheffer & Huston 1997; Miriti 2006;
Schiffers & Tielborger 2006). How these processes contribute
to natural reassembly of plant communities, or to restoration
of damaged or degraded systems, requires an understanding of
the physical and biotic context in which species interact, and
the nature, strength and timing of these interactions (Clark
et al. 1999; Garcı́a, Obeso & Martı́nez 2005; Gómez-Aparicio,
Gómez & Zamora 2005; Brooker et al. 2008). Despite the
importance of context dependency for many ecological processes (Jones & Callaway 2007), relatively few studies, mainly
in the northern hemisphere and tropics, have considered the
role of community context in regulating the establishment of
woody plants during early succession (Berkowitz, Canham &
Kelly 1995; Burton & Bazzaz 1995; Gómez-Aparicio et al.
2004; Benitez-Malvido 2006; Acacio et al. 2007). Community
context may be critical to this process because variation in
environment and disturbance, combined with the stochastic
nature of dispersal, can lead to significant heterogeneity in
community structure (Halpern 1988; Traveset et al. 2003;
Zavaleta, Hulvey & Fulfrost 2007).
The rural landscape over much of southern Chile, and in
other regions of South America, is a mosaic of vegetation
patches of varying size, human influence and successional stage
(Echeverrı́a et al. 2007). In the Lake District, and on Chiloé
Island (39–42 S), widespread logging and farming during the
20th century (Lara, Donoso & Aravena 1996), resulted in conversion of native evergreen forests to pastures, croplands and
seral shrublands. Recolonization of these shrublands by trees
has been slow and highly variable, spurring interest in the factors that limit establishment and growth (Aravena et al. 2002;
Dı́az & Armesto 2007; Dı́az, Bigelow & Armesto 2007). Here,
we experimentally test how early seral communities of differing
composition, structure and physical environment influence the
germination and ⁄ or survival of native tree species with
differing life histories and successional roles (shade tolerance).
This is the first study in a temperate ecosystem of the southern
hemisphere to compare controls on tree establishment among
seral scrubland communities with differing bio-physical environments. These include: (i) open shrubland (<25% cover)
dominated by Berberis buxifolia Lam. (Berberidaceae); (ii)
denser shrubland (c. 50% cover) with seasonally waterlogged
soils, dominated by Baccharis patagonica (H. et A.)
(Asteraceae), (Dı́az & Armesto 2007; Dı́az, Bigelow &
Armesto 2007) and (iii) moderately dense shrubland (c. 30%
cover) also dominated by B. patagonica, with significant establishment of small trees (mainly 20- to 30-year-old Drimys
0
25
50 km
74°W
73°W
Fig. 1. Study area at SDBS, northern Chiloé Island.
2010 The Authors. Journal of Ecology 2010 British Ecological Society, Journal of Ecology, 99, 288–299
72°W
290 M. A. Bustamante-Sánchez, J. J. Armesto & C. B. Halpern
summer). Mean minimum and maximum monthly temperatures are
3 C in July and 17 C in January, respectively (SDBS, meteorological records from 1999 to 2007). Soils are primarily ñadis (Veit &
Garleff 1996), characterized by an impermeable hardpan at c. 50–
60 cm depth which results in a shallow water table and saturated soils
during winter (June–August), especially in non-forested sites. Such
soils are particularly well developed over fluvioglacial deposits and in
depressions between late Quaternary moraine fields (Aravena 1991).
ground cover of early successional grasses and herbs (sometimes with
remnants of the original forest understorey). Common ferns and low
shrubs included Blechnum chilense (Kaulf.) Mett. and B. penna-marina (Poir.) Kuhn (Blechnaceae), Gaultheria mucronata (L.F.) Hook.
et Arn. (Ericaceae) and Myrteola numularia (Poir.) Berg (Myrtaceae).
In addition, the Drimys community has significant numbers of juveniles and saplings of two tree species, D. winteri and N. nitida
(Table 1).
EARLY SUCCESSIONAL COMMUNITIES
NATURAL TREE RECRUITMENT AND RELATIONSHIPS
In this rural landscape, early successional vegetation varies greatly
among sites reflecting differences in edaphic conditions, disturbance
severity and the persistence and abundance of biological legacies (Papic 2000; Carmona et al. 2002). Large areas disturbed by logging and
fire have become poorly drained shrublands with dense cover of Baccharis patagonica and scattered patches of Sphagnum moss (Dı́az,
Bigelow & Armesto 2007; Dı́az et al. 2008; Carmona et al. 2010).
Under these conditions, tree colonization is rather low or absent (Papic 2000). Where fire severity is lower and woody debris and stumps
persist, colonization by trees (e.g. D. winteri, Nothofagus nitida (Phil.)
Krasser (Nothofagaceae)) may be more common (Aravena et al.
2002). Better-drained sites are typically invaded by shrubs in the
genus Berberis (B. buxifolia and B. darwinii) (Carmona et al. 2010).
We selected three early successional communities (sites) in SDBS that
typically develop after logging and burning of evergreen forests.
These are named for the primary woody species, i.e. Berberis (B .buxifolia), Baccharis (B. patagonica) and Drimys (D. winteri). All have
developed after fire, are of similar age (c. 50–60 years), and occur on
similar parental material in similar landscape contexts (i.e.
surrounded by pasture and mature North-Patagonian forest). Communities differ, however, in species composition, vegetation structure,
and some bio-physical characteristics (Table 1). Each is dominated
by one of two pioneer shrubs (Berberis or Baccharis) above a dense
WITH ENVIRONMENTAL FACTORS
In each community we first quantified the density of natural (native)
tree establishment. We used a systematic sample of 60 circular plots
(3 m radius) regularly distributed along six parallel transects spaced
16 m apart. In each plot we recorded by species the number of juveniles (<1.3 m tall) and saplings (>1.3 m in height, <5 cm dbh). In
each plot we also measured four components of the biotic and abiotic environment that could potentially explain variation in tree density. (i) Total shrub cover was estimated visually by two observers
by cover class (0, 1–25, 26–50, 51–75 and 76–100%). (ii) Total cover
in the herb layer (sum of the cover of herbs, grasses, ferns and small
shrubs). (iii) Light availability (percent transmittance of diffuse photosynthetic photon flux density; PPFD) was measured at three random locations under overcast sky conditions. Measurements were
taken with a Li-Cor LI-190SA quantum sensor (LI-COR Biosciences, Lincoln, NE, USA) mounted on a small, self-levelling platform
10 cm above the ground surface and, as a reference, 2 m above the
canopy (Parent & Messier 1996). (iv) Volumetric water content of
the topsoil (0–12 cm depth) was measured at four random locations
by time domain reflectometry (TDR) using a portable probe (TDR
100 Soil Moisture Meter, Plainfield, IL, USA). All environmental
measurements were made in November 2007, at the beginning of the
dry season.
Table 1. Vegetation and physical characteristics of the three early successional communities at Senda Darwin Biological Station (SDBS), Chiloé
Island. Communities are named for the dominant shrub or tree species (Berberis, Baccharis, and Drimys)
Communities
Tree layer height (m)*
Shrub layer height (m)
Tree cover (2–3 m) (%)
Shrub cover (0.5–1.6 m) (%)
Herb cover (<0.5 m) (%)
Coarse woody debris cover (%)
Light transmittance (%)
Soil drainage
Berberis
Baccharis
Drimys
–
1.09
<1
22
86.8
<1.5
63.4
High
–
1.61
<1
45
103.9
3.3
23.4
Low
2–3
1.28 (±0.04)
8.5 (±1.8)
28 (±2.9)
97.3 (±3.8)
20.2 (±2.5)
28.1 (±2.2)
High
(±0.03)
(±1.9)
(±8.1)
(±3.1)
(±0.03)
(±2.6)
(±3.4)
(±3.6)
(±2.1)
Juvenile trees†
Density (ind. ha)1)
Number of species
226
9
333
9
1661
17
Sapling trees‡
Density (ind. ha)1)
Number of species
125
3
30
5
1012
13
Values are means (±1 SE).
*Tree stems with dbh ‡5 cm.
†Juveniles, dbh <5 cm and <1.3 m tall.
‡Saplings, dbh <5 cm and ‡1.3 m tall.
2010 The Authors. Journal of Ecology 2010 British Ecological Society, Journal of Ecology, 99, 288–299
Controls on tree colonization in shrublands 291
Table 2. Characteristics of tree species used in experimental studies of seed germination (G) and seedling survival (S) in three early successional
communities at SDB
Species
Family
Shade
tolerance
Seed
mass (mg)
Fruit and dispersal
type
Study
Amomyrtus luma
Amomyrtus meli
Embothrium coccineum
Eucryphia cordifolia
Gevuina avellana
Myrtaceae
Myrtaceae
Proteaceae
Cunoniaceae
Proteaceae
T
T
I
M
M
35
40
14
1.7
1040
berry, ornithochory
berry, ornithochory
samara, anemochory
samara, anemochory
dry nut, gravity
G,
G
G,
G,
G,
S
S
S
S
Shade tolerance is coded as I = intolerant, M = mid-tolerant and T = tolerant.
Source: Dı́az, Papic & Armesto (1999), Figueroa & Lusk (2001) and Aravena et al. (2002).
EXPERIMENTAL SPECIES
Seed germination trials
Experimental trials to test the germination and early survival of trees
included five temperate forest species (Table 2) with differing successional roles (Figueroa & Lusk 2001; Aravena et al. 2002). Species differed in shade tolerance, seed size and seed dispersal strategy. All
were present in the canopy of natural stands in the study area.
Seeds of each of the five tree species were collected during the month
of maximum fruit load (Smith-Ramı́rez & Armesto 1994). Seeds were
extracted from fruits and stored for up to 2 months at room temperature. Seeds were sown in May 2007 (beginning of the wet season) in
plastic pots (15 cm depth, 6 cm diameter) filled with soil from the
communities into which they would be placed. Each experimental
unit received two pots for each tree species; each pot contained 10
seeds placed on the soil surface. To prevent predation, the 10 pots
per experimental unit were enclosed in a wire-mesh cage
(50 · 50 · 70 cm) and placed on the forest floor. To account for natural seed rain, we placed an extra pot containing soil in each experimental unit (no seedlings emerged from these controls by the end of
the experiment). Seed germination was monitored monthly for
11 months.
Seeds of each tree species were assessed for viability in three ways:
percent seed germination in the laboratory under controlled photoperiod (12-h light, 12-h dark) and temperature (10–20 C); percent germination under the canopy of a second-growth forest at SDBS and
by a tetrazolium test (detailed descriptions of methods are provided
in Appendix S1 in Supporting Information).
EXPERIMENTAL DESIGN
Experimental treatments
To test the influence of early successional vegetation on germination
and early survival, seeds and seedlings of each tree species were
sown ⁄ planted in replicate plots (experimental units) in each successional shrubland. Plots were subjected to one of two treatments:
removal of all aboveground vegetation or control (no removal). We
used a randomized block design, with treatments randomly assigned
to paired plots (5 · 5 m, separated by a 5-m buffer) in each of five
blocks in each of the three communities (total of 30 experimental
units). Blocks were separated by >30 m. In removal plots, all plants
were clipped at the ground surface. Treatments were initiated in May
2006 and maintained over the course of study (36 months).
Seedling survival and growth trials
Abiotic factors
To assess the effects of vegetation on local environment, we measured
light availability and soil properties in each of the 30 experimental
units. Light availability (% transmittance, as described above) was
measured at the start of the dry season (November 2007). Four measurements were made (one in each cardinal direction) above seeds
and seedlings (see below). Soils were collected twice in the rainy season (August 2006 and July 2007) and once in the dry season (February 2007). At each sampling date we collected one soil core (0–10 cm)
per experimental unit. Each core was divided into four subsamples.
From these we obtained gravimetric soil moisture content (SMC),
)
pH, available nitrogen (NH+
4-N and NO3 -N) and total carbon (C)
and nitrogen (N). SMC was expressed as mass of water ⁄ mass of dry
soil. pH was determined in a 1:2 suspension (5 g dry soil:10 ml deionized water). For extraction of available N we used a 2% KAl(SO4)2
)
solution. Concentrations of NH+
4-N and NO3 -N were determined by
fractionated steam distillation (Pérez, Hedin & Armesto 1998). To
determine total C and N, soil samples were oven-dried at 70 C and C
and N content were estimated by flash combustion in a Carlo Erba
NA 2500 elemental analyzer at the Biogeochemistry Lab, Pontificia
Universidad Católica de Chile. Total C and N were determined for
only two of the three sampling dates (August 2006 and February
2007).
Seedlings of four of the five tree species were planted at two different times (seeds of the fifth species, A. meli, did not germinate in
sufficient number to include in these trials). We first planted 1-yearold A. luma, 1-year-old E. cordifolia, and 2-year-old Embothrium
coccineum J.R. et G. Foster in July 2006 (middle of the wet season).
Subsequently, we planted 1-year-old Gevuina avellana Mol. and 1year-old E. coccineum in May 2007 (beginning of the wet season).
Seedlings were grown in a greenhouse from seeds collected locally.
After germination, seedlings were placed in 20 · 20 cm black plastic bags containing sieved topsoil from nearby forests and retained
in the greenhouse (SDBS). Before planting out, seedlings were
allowed to acclimate for 1 month outside the greenhouse. Finally,
one seedling of each species was transplanted into each experimental unit (total of 150 seedlings). Survival of seedlings planted in July
2006 was monitored monthly for 1 year, then every 2 months
through the second year (total of 660 days). Seedlings planted in
May 2007 were monitored monthly for 1 year (total of 334 days).
A seedling was considered dead at the point that it had no green
leaves and did not resprout during the next following 4 months.
Shoot length of surviving individuals was measured at the end of
each growing season. To account for size variation at the time of
planting (10–40 cm), relative growth rate (RGR; Hunt 1982) was
2010 The Authors. Journal of Ecology 2010 British Ecological Society, Journal of Ecology, 99, 288–299
292 M. A. Bustamante-Sánchez, J. J. Armesto & C. B. Halpern
calculated as (ln (final shoot length) – ln (initial shoot length)) ⁄ t,
where t = time interval (660 or 334 days).
STATISTICAL ANALYSES
Relationship between natural tree recruitment and
environmental variables
We explored relationships between natural tree recruitment and
the biotic and abiotic environment by modelling densities of juveniles and saplings as a function of local environmental variables.
We fitted general linear models (GLM) using Poisson error, a loglink function and a quasi-likelihood approach to overcome potential difficulties with over-dispersion (McCullagh & Nelder 1989).
Because of the small numbers of some species, we modelled densities for two groups of trees based on shade tolerance: intolerant
vs. mid-tolerant ⁄ tolerant. Separate models were developed for
juveniles and saplings of each group. Successional community was
not included in these models to avoid collinearity with local environmental variables (variance inflation factor >1.5; Booth, Niccolucci & Schuster 1994). Candidate environmental predictors
included total cover in the herb layer, total shrub cover, light
availability and SMC (at the start of the dry season). Initial models included all predictors and were simplified using a backwardremoval procedure to eliminate non-significant effects (Crawley
1993). Only variables explaining a significant amount of deviance
were retained in the final models. Model parameters were fitted
using maximum-likelihood, and statistical significance was assessed
by analysis of deviance (McCullagh & Nelder 1989). Analyses
were conducted using R 2.5.1 software (R Development Core
Team 2005).
Responses to experimental treatments
We used distance-based permutational analysis of variance (Anderson 2001; McArdle & Anderson 2001) to assess effects of removal
treatments and community type on physical environment, seed germination, and survival and height growth of seedlings. We used Euclidean distance as the distance measure and 9999 random permutations
to determine significance of the pseudo F-statistic. For each model,
sources of variation included community type (df = 2; fixed), block
(nested within community, df = 12, random), vegetation removal
(df = 1, fixed), community · vegetation removal interaction
(df = 2), and block · vegetation-removal interaction (df = 12). In
addition, for soil variables that were sampled multiple times (two to
three dates), models included time and all relevant two- and threeway interactions of time with community, block and ⁄ or vegetation
removal.
For germination and seedling survival, tree species were treated as
a multivariate response. If species showed significant differences in
response, the multivariate test was followed by univariate analyses
for each species, followed by a posteriori tests of means (using 9999
permutations to determine significance; Anderson 2005). Survival
was expressed as longevity (number of days that a seedling survived
based on the last census date); separate tests were therefore conducted
for survival of species planted in 2006 and 2007.
For RGR, separate univariate tests were conducted for seedlings of
E. coccineum (both 1- and 2-year-old) and E. cordifolia in control
plots, the only species and treatment for which there were sufficient
numbers of survivors to compare height growth among communities.
For E. coccineum, seedling age was considered a factor in addition to
community type.
Results
RELATIONSHIP BETWEEN NATURAL TREE
RECRUITMENT AND ENVIRONMENTAL VARIABLES
We counted a total of 368 juveniles and 192 saplings of 19
species in the three succesional communities (Appendix S2 in
Supporting Information). Among the juveniles, 201 were
shade-intolerant and 167 were mid-tolerant or tolerant species.
Among the saplings, 145 were shade-intolerant and 47 were
mid-tolerant or tolerant species. Densities of juveniles and
saplings were one to two orders of magnitude greater in the
Drimys than in the Berberis or Baccharis communities
(Table 1). Transition probabilities from juvenile to sapling
stage (expressed as the ratio of densities) were higher in Drimys
(0.61) than in Berberis (0.55) and lowest in the Baccharis
community (0.09).
Abiotic and biotic variables explained a total of 19–35%
of the deviance in models of natural densities of juveniles
and saplings (Table 3). Light transmittance was a significant
predictor of density (negative correlation) in all models
(P < 0.001). It explained greater deviance for intolerant
than for mid-tolerant ⁄ tolerant species (19–23% vs. 12–13%).
Table 3. Results of GLMs predicting densities of natural tree establishment as a function of abiotic and biotic variables. Densities of juveniles
(<1.3 m tall) and saplings (‡1.3 m tall, <5 cm dbh) of individual tree species were pooled by degree of shade tolerance (intolerant vs. midtolerant or tolerant). Values are variable coefficients (Coeff) and deviance explained (Dev, %), plus deviance explained by the full model (R2adj)
Light transmittance
Soil moisture
Shrub cover
Herb cover
Model
Coeff
Dev
Coeff
Dev
Coeff
Intolerant
Juveniles
Saplings
)0.04***
)0.04***
19
23
)0.06**
8
)0.03***
4
Mid-tolerant, tolerant
Juveniles
)0.04***
Saplings
)0.02***
12
13
)0.02***
)0.06***
11
19
Dev
Coeff
Dev
R2adj
0.19
0.35
)0.01*
3
0.26
0.32
Significance of coefficients is coded as: ***P £ 0.001, **P £ 0.01, *P £ 0.05. All measurements were made at the start of the dry season
(November 2007).
2010 The Authors. Journal of Ecology 2010 British Ecological Society, Journal of Ecology, 99, 288–299
Controls on tree colonization in shrublands 293
Shrub cover also showed negative correlations with the density of juveniles and saplings in three of four models
(P < 0.001). However, it explained greater deviance for
mid-tolerant ⁄ tolerant than for intolerant species (11–19% vs.
4%, respectively). Soil moisture was a significant predictor
(P < 0.01) of density for shade-intolerant saplings (negative
correlation, 8% of deviance). Herb cover was a significant
predictor (P < 0.01) for mid-tolerant ⁄ tolerant saplings (negative correlation, 3% of deviance).
EXPERIMENTAL REMOVAL TREATMENTS
Variation in physical environment and soils
As expected, vegetation removal resulted in a highly significant
increase in light transmittance (35–59% in controls vs. 72–
94% in removals; Appendix S3 in Supporting Information).
However, vegetation removal did not have an effect on SMC
(Figs 2a–c), C:N (Fig. 2d), or available N (Fig. 2e; Appendix
S3). Effects on pH varied with community type (significant
treatment by community interaction), but differences among
treatments were very small (Fig. 2f, Appendix S3). Among
communities there was significant variation in light availability
(Berberis = Baccharis > Drimys, data not shown), in SMC
(consistently lowest in Berberis in all seasons and highest in
Baccharis in the dry season; Figs 2a–c), and in C:N (Drimys >
Baccharis > Berberis; Fig. 2d; Appendix S3). The only significant variation in available N occurred among seasons (Fig. 2e;
Appendix S3).
Seed germination
Estimates of seed viability varied by method and species
(Table 4). Viability was consistently low for A. meli (<25%),
but ranged from 49% in A. luma to as high as 96–100% in E.
coccineum and G. avellana.
Tree species varied in their responses to vegetation removal
(pseudo-F(1,12) = 10.01, P(perm) = 0.0005). Germination was
significantly lower in removals than controls for A. luma and
E. cordifolia, but comparable between treatments for A. meli,
E. coccineum, and G. avellana (Fig. 3, Appendix S4 in Supporting Information). For all tree species, responses to removals
were consistent among communities (P > 0.05 for all community · vegetation removal interactions; Fig. 3, Appendix S4).
However, species differed in their germination among communities (pseudo-F(2,12) = 5.49, P(perm) = 0.003). Amomyrtus
luma, E. cordifolia and G. avellana showed greater germination
in the Drimys than in the Berberis or Baccharis communities,
but E. coccineum and A. meli showed comparable but low germination among types.
Seedling survival
Species responded differently to vegetation removal (2006:
pseudo-F(1,12) = 21.13, P(perm) = 0.0001; 2007: pseudoF(1,12) = 26.36, P(perm) = 0.0001). Survival was significantly
lower in removals than in controls for E. coccineum (2006–
2007) and E. cordifolia (2006), but only marginally so for
A. luma and G. avellana (Fig. 4, Appendix S4). In addition, for
(a)
(b)
(c)
(d)
(e)
(f)
Fig. 2. Variation in soil properties among vegetation removal treatments, successional communities and seasons. Significance values (P) are from
univariate permanovas (see Appendix S3 for details) testing effects of season (Seas), community (Comm) and vegetation treatment (Veg).
ns = non-significant main effect (P > 0.05). Two-way interactions are noted if significant. Results of a posteriori tests of means (following
significant main effects or interactions) are coded by lower-case letters. C:N was measured in two seasons. pH varied significantly among seasons,
but differences were small, thus means of seasons are presented. Where main effects or interaction terms were not significant (C:N and available
N), means are presented.
2010 The Authors. Journal of Ecology 2010 British Ecological Society, Journal of Ecology, 99, 288–299
294 M. A. Bustamante-Sánchez, J. J. Armesto & C. B. Halpern
Table 4. Mean viability of seeds (%) of the five experimental species
determined by three methods: percent germination in the laboratory,
percent germination in second-growth forest and viability via
tetrazolium (E. coccineum and G. avellana only)
Germination (%)
Viable (%)
Species
Laboratory
Forest
Tetrazolium
Amomyrtus luma
Amomyrtus meli
Embothrium coccineum
Eucryphia cordifolia
Gevuina avellana
1
11
33
62
64
49
24
7
56
100
–
–
96
–
100
most species, responses to vegetation removals were consistent
among communities. Only 1-year-old E. coccineum (2007)
showed differing responses among communities (i.e. no
response to removal in Baccharis; Fig. 4). Otherwise, rates of
survival did not vary among communities (non-significant
main effects of community type; Fig. 4, Appendix S4). Most
mortality occurred during or after the dry season (between
20% and 100% depending on the species or community).
Seedling height growth
Analyses of RGR were limited to E. coccineum (1- and
2-year-old) and E. cordifolia (1-year-old) in control plots (with
E. cordifolia further limited to Drimys and Berberis communi-
ties). These were the only species-by-treatment ⁄ community
combinations for which there was a sufficient number of
surviving seedlings. For E. coccineum, RGR did not differ for
1- and 2-year-old seedlings (F(1,18) = 3.04, P = 0.09), but it
did vary among communities: height growth was greater in
Drimys and Berberis communities than in the Baccharis
community (F(2,11) = 6.3, P = 0.01; Fig. 5). In contrast,
RGR did not vary among communities for E. cordifolia.
Discussion
Variation or heterogeneity is often viewed as a ‘problem’ in
experimental studies. As a consequence, studies of plant–plant
interactions are often limited to particular environments or
community contexts. However, the strength of important
biotic processes such as competition or facilitation is often contingent on the ecological context (i.e. local resource availability
or environmental stress; Pugnaire & Luque 2001; Tewksbury
& Lloyd 2001; Callaway et al. 2002; Kuijper, Nijhoff & Bakker
2004). Similarly, the relevance of these processes may change
with stage of plant development (e.g. seed, seedling, or adult;
Holmgren, Scheffer & Huston 1997; Miriti 2006; Schiffers &
Tielborger 2006). Ours is the first study in temperate forest ecosystems of the southern hemisphere to consider the importance
of context dependency in time and space for the recruitment of
tree species into early successional shrublands. We demonstrate the overriding importance of biotic controls (facilitation)
on seed germination and early survival, variation in the timing
Fig. 3. Germination (proportion of seeds) in
removal and control plots in each of the three
early successional communities. Values are
means +1 SE. Significance values (P) are
from univariate permanovas (see Appendix
S4 for details) testing effects of community
(Comm) and vegetation treatment (Veg),
ns = non-significant main effect (P >
0.05). Comm · Veg interactions are noted
if significant; results of a posteriori tests of
community means are coded by lower-case
letters.
2010 The Authors. Journal of Ecology 2010 British Ecological Society, Journal of Ecology, 99, 288–299
Controls on tree colonization in shrublands 295
Fig. 4. Survival (number of days seedlings
remained alive) in removal and control plots
in each of the three early successional communities. Seedlings were planted in 2006
(maximum survival of 660 days) or 2007
(maximum survival of 334 days). Seedlings
of E. coccineum planted in 2006 were 2 years
old; all others were 1-year-old. Values are
means + 1 SE. Significance values (P) are
from univariate permanovas (see Appendix
S4 for details) testing effects of community
(Comm) and vegetation treatment (Veg).
ns = non-significant main effect (P >
0.05). Comm · Veg interactions are noted if
significant; results of a posteriori tests are
coded by lower-case letters.
of these effects for species with differing successional roles and
the outcomes of these interactions among communities with
differing physical constraints.
BIOTIC INTERACTIONS IN EARLY SUCCESSIONAL
COMMUNITIES AND CONTEXT DEPENDENCE
Fig. 5. Relative height growth rates of surviving seedlings of Embothrium coccineum and Eucryphia cordifolia in control plots in each of
the three early successional communities. Values are means +1 SE.
Results of a posteriori tests of community means are coded by
lower-case letters. Too few seedlings of E. cordifolia survived in the
Baccharis community to estimate RGR (nd).
The results of interactions between established vegetation and
colonizing trees were highly consistent among the communities
studied. We did not observe changes in the direction of community effects, e.g. from competition to facilitation, or the
reverse, and only one species and life-history stage (1-year-old
E. coccineum seedlings) showed variation in the intensity of
interaction (facilitation under Berberis and Drimys but not in
the Baccharis community; Fig. 5). This pattern is consistent
with the theoretical expectation of stronger positive interactions in more stressful environments (Callaway 1995): facilitation by existing plant cover was observed where soil water
availability was lowest during the dry season (Berberis and
Drimys communities).
2010 The Authors. Journal of Ecology 2010 British Ecological Society, Journal of Ecology, 99, 288–299
296 M. A. Bustamante-Sánchez, J. J. Armesto & C. B. Halpern
CONTRASTING EFFECTS OF ESTABLISHED
VEGETATION ON TREE SPECIES AND THEIR LIFEHISTORY STAGES
Tree species showed a diversity of responses to existing plant
cover. Survival of shade-intolerant E. coccineum and mid-tolerant E. cordifolia in experimental plots was enhanced under
intact vegetation, but survival of shade-tolerant A. luma was
not. Thus, shrub cover in these early successional communities
is not too dense to inhibit establishment of pioneering species,
but it is too sparse to enhance survival of late-seral, shade-tolerant species. This interpretation is supported by observations
of the light environments of naturally established seedlings in
nearby second-growth forests. Amomyrtus luma, G. avellana
and E. cordifolia typically occur where light is low (3–8% of
ambient levels) and E. coccineum, where it is much greater
(>25%; Figueroa & Lusk 2001). Moreover, in a previous
study of forest succession on Chiloé Island, light requirement
was an important determinant of seedling establishment (Aravena et al. 2002). Our study illustrates that small differences in
shrub cover and light availability among early successional
communities may determine which tree species can successfully
establish.
The combined results of germination and survival experiments also illustrate that, for some species, responses to
vegetation removal treatments change during the life history.
For shade-tolerant A. luma, presence of understorey vegetation enhanced germination, but did not affect seedling
survival. In contrast, for shade-intolerant E. coccineum,
understorey vegetation did not affect germination but
enhanced survival. In contrast, in species classified as
mid-tolerant, both life stages responded similarly, either
benefiting from (E. cordifolia), or showing no response
(G. avellana) to shading. In accordance with other studies
(Armesto & Pickett 1986; Walker & Chapin 1987; De Steven
1991a,b; Gill & Marks 1991), patches that promote germination and emergence are not necessarily the same as those
that enhance survival and growth. Alternatively, difference
in performance over time may reflect changes in physiological requirements as organisms grow larger (Miriti 2006;
Schiffers & Tielborger 2006). Understanding how the effects
of vegetation context can change in time or space requires
an integrative, ‘linking-stages’ approach – one that considers
how responses in an earlier stage may cascade to the next
(Schupp & Fuentes 1995).
POTENTIAL MECHANISM OF FACILITATION
There may be several explanations for greater seed germination and seedling survival in control than in removal plots.
Although we cannot confirm the mechanism(s) by which
established vegetation facilitates native tree seedlings, measurements of physical conditions may provide insights. First,
among the physical ⁄ edaphic variables measured, only solar
radiation (PPFD) differed consistently between treatments.
Shrub cover in these early successional communities resulted
in a 34–40% reduction of diffuse radiation relative to experi-
mentally cleared plots. Similarly, GLMs predicting the density of natural regeneration indicated consistent negative
relationships with PPFD (Table 3). High levels of solar radiation may cause photoinhibition in plant species adapted to
low-light environments (Valladares 2003). Presence of shrubs
can enhance germination and survival early in succession by
limiting transmission of solar radiation, protecting against
freezing, reducing air and soil temperatures and potentially
increasing soil moisture availability during critical periods of
the year (Vetaas 1992; Belsky 1994). Soil water content did
not differ between experimental removal and control plots,
suggesting that greater plant uptake in the controls may balance greater evaporation in the removals. Recruitment in the
open thus appears to be limited by the direct effects of radiation, not by insufficient soil moisture. The timing of seedling
mortality in removal plots – during or immediately after the
dry austral summer – further underscores the benefits of
shading for tree seedlings in these early successional communities (Walker & Chapin 1987; Callaway 1995; Baumeister &
Callaway 2006). On the other hand, GLMs indicated negative correlations between shrub cover and natural densities
of juveniles and saplings (particularly for shade-tolerant species). This suggests that positive effects of shrubs in reducing
light may be balanced, in part, by competition for belowground resources (nutrients). Shade-tolerant species, which
show greater allocation to light acquiring structures, may be
more sensitive to this competition (Coomes & Grubb 2000;
Lusk 2004).
BIOTIC AND ABIOTIC FILTERS
Abiotic and biotic factors act as filters that operate in parallel
or sequentially during recruitment from seed to adult stages
(Schupp & Fuentes 1995; George & Bazzaz 1999; Fattorini &
Halle 2004). Our removal experiments provide strong evidence
that the effects of established vegetation do not differ among
the communities studied, i.e. that germination and early survival of trees were not affected by inherent differences in community structure or composition. However, other biotic
factors could contribute to the observed differences in tree density and size structure among communities. The large numbers
of juveniles in the taller Drimys community could be explained
by greater seed inputs, facilitated by an abundance of saplings
that serve as perch sites for frugivorous birds (McDonnell &
Stiles 1983; Hernández 1995; Pausas et al. 2006). Alternatively,
seed predation may be greater in the Berberis and Baccharis
communities. Differential predation of seedlings seems unlikely, however. We did not observe evidence of seedling predation by vertebrates, consistent with previous studies in this
type of forest (Figueroa & Castro 2000). Higher juvenile densities and germination rates under Drimys could also reflect
greater potential for mycorrhizal infection where sapling
densities are higher and greater amounts of detritus from the
pre-disturbance community were left (Allen, Allen &
Gómez-Pompa 2005; Urgiles et al. 2009). Finally, the greater
abundance of coarse woody debris in this community
(Table 1) may enhance tree survival via shading, water
2010 The Authors. Journal of Ecology 2010 British Ecological Society, Journal of Ecology, 99, 288–299
Controls on tree colonization in shrublands 297
retention during summer or effects on soil resources (Papic
2000; Carmona et al. 2002).
Several lines of evidence suggest that variation in establishment and density of trees may also reflect inherent differences
in the edaphic characteristics of these communities. Seedling
mortality in most species occurred during or just after summer
(dry season), when SMC was lowest, suggesting potential for
water stress (Figueroa & Castro 2000). Moreover, dry season
soil moisture was lowest in the Berberis community where juvenile densities were also lowest. Although seasonal drought in
this system (Carmona et al. 2010) may not be as extreme as in
Mediterranean-type ecosystems in central Chile, global and
regional climate models predict increasingly lower summer
rainfall for south-central Chile (Walther et al. 2002; IPCC
2007; Lara, Villalba & Urrutia 2008). If soil moisture during
the dry season currently limits tree establishment, this
projected drying trend could impose even stronger constraints
in the future.
Physical limitations in the Baccharis community may be
very different, however. Here, recruitment may be limited by
seasonal waterlogging of soils, which is characteristic of this
shrubland vegetation (Dı́az, Bigelow & Armesto 2007). Several
relationships support this interpretation. First, for natural tree
populations, sapling density was extremely low, as was the
transition probability (density ratio) from juveniles to saplings
(0.09 vs. >0.55 in Berberis and Drimys communities). This
suggests significantly lower rates of height growth and ⁄ or
higher mortality between juvenile and sapling stages. Second,
for shade-intolerant E. coccineum (the only species for which
comparisons were possible), growth rate was distinctly lower
in the Baccharis community. Reduced height growth and premature leaf abscission (which we also observed) are classic
symptoms of waterlogging in susceptible species (Piper et al.
2008). Finally, the results of GLMs for shade-intolerant trees
were consistent with this pattern: densities of saplings were
negatively correlated with soil water content at the start of the
dry season, suggesting lower rates of recruitment in wetter
areas (predominantly in the Baccharis community).
In sum, our experiments indicate strong biotic controls
(facilitation) on tree recruitment among early seral shrublands
with contrasting structures, compositions and edaphic characteristics. For most tree species in these forests, which are relatively shade tolerant, recruitment in the absence of vegetation
may be limited by excessive solar radiation. Our results also
illustrate that the benefits of shading may accrue at different
stages in the life history for species with differing light requirements. For some species, community context is also critical in
its effect on germination. This, together with differences in seed
dispersal, seed predation, mycorrhizal availability and other
site characteristics, may explain the dramatic differences in
natural tree recruitment in Drimys vs. Berberis or Baccharis
communities.
Seasonal patterns of soil moisture may also contribute to
variation in recruitment. Water deficit during summer may
limit seedling survival in some communities; water excess
during winter may reduce growth. In some community
contexts (e.g. Baccharis shrublands) both types of stress can
occur, limiting both establishment and the ability of juveniles
to progress to the sapling stage and deterring successional
change. In combination, different filters of varying strength
contribute to differing rates of tree establishment and
growth, and ultimately, to differing rates of succession from
shrubland to forest.
IMPLICATIONS FOR RESTORATION
There are many situations in which forest succession is arrested
or delayed by a recalcitrant shrub or herbaceous layer (Walker
1994; Mallik 1995). Various factors may contribute to
regeneration failure in these systems (e.g. resource preemption,
allelopathy, soil type or physical obstruction by dense litter),
but all are indicative of negative interactions that suggest the
need for vegetation removal or soil scarification to enhance
tree establishment. In contrast, in the southern temperate
shrublands of Chiloé Island, strong, positive interactions
between shrubs and trees suggest that attempts at restoration
through planting will only be successful if trees are placed
beneath established vegetation.
The differences in performance among the species in
our experiments underscore the notion that the outcome of
restoration may be highly dependent on the selection of species and their physiological traits. Light requirement is an
important attribute to be considered during the selection process. Species that are likely to survive are those that can maximize resource uptake, but limit water loss during the growing
season (Padilla et al. 2009). Of the four species examined,
only E. coccineum (relatively shade intolerant) and E. cordifolia (mid-tolerant) showed consistently high survival and reasonable height growth. Thus, despite the ecological benefits of
planting for diversity, it may not be possible to introduce later
successional species until environmental conditions are conducive to their survival and growth. Thus, restoration strategies whose goal is to achieve functional or taxonomic
diversity may require that planting is staged over time. Species-specific information on the light requirements and spatial
distributions of seedlings in mature forests (Figueroa & Lusk
2001) could be useful in determining when, and into what
contexts, species can be successfully planted. Alternatively, if
seed sources and dispersers are not limiting in the surrounding
landscape, recruitment may occur naturally but slowly as an
outcome of increasing structural diversity, as pioneer species
begin to overtop the shrub layer and offer perch sites for seed
dispersers (McDonnell & Stiles 1983; Pausas et al. 2006).
Long-lived shrublands in south-central Chile are often
drained and planted with exotic tree species by private owners
and foresters who erroneously assume that forest will not
recover (Armesto et al. 2009).
Natural tree densities and experimental results suggest that
potential for restoration of these three early successional
shrublands may vary. Notably, limited soil drainage and excessive waterlogging may impose severe constraints on regeneration in Baccharis-dominated communities (Dı́az & Armesto
2007; Dı́az, Bigelow & Armesto 2007). Restoration of tree
cover may require planting on elevated substrates or improv-
2010 The Authors. Journal of Ecology 2010 British Ecological Society, Journal of Ecology, 99, 288–299
298 M. A. Bustamante-Sánchez, J. J. Armesto & C. B. Halpern
ing drainage prior to planting (Armesto et al. 2009). A diversity of approaches may be adopted from strategies used in
European or North American ecosystems where saturated
soils limit tree regeneration (Sutton 1993).
Acknowledgements
We are grateful to the many individuals who assisted with this study: Juan
Vidal for collecting seeds and helping to establish the experiments; Javier
Simonetti, Andres Charrier, Wara Marcelo and Victor Sagredo for providing
equipment, assistance in the field and lab, and technical support during soil
analyses; and Fabian Jaksic for lab space. We especially thank members of
the Halpern and Franklin labs (School of Forest Resources, University of
Washington) for helpful discussions and comments on earlier drafts of the
manuscript. We appreciate the constructive suggestions of two anonymous
reviewers. M.A.B.S. was funded by a doctoral CONICYT fellowship, Chile,
and a Fulbright-CONICYT fellowship during preparation of the manuscript.
This is a contribution to the Research Programs of SDBS, LTSER-network,
Chile and the Biogeochemistry Laboratory, Ecology Department at Pontificia
Universidad Católica, Chile.
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Received 21 April 2010; accepted 23 August 2010
Handling Editor: Charles Canham
Supporting Information
Additional Supporting Information may be found in the online version of this article:
Appendix S1. Description of methods used to evaluate seed viability
prior to germination trials.
Appendix S2. Family, species, shade tolerance, and juvenile and sapling abundance for tree species found in the Berberis, Baccharis, and
Drimys communities of northern Chiloé Island.
Appendix S3. Results of permutational analysis of variance for soil
properties and light.
Appendix S4. Results of permutational univariate analysis of variance
(by tree species), for germination and survival of seedlings planted in
2006 and 2007.
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2010 The Authors. Journal of Ecology 2010 British Ecological Society, Journal of Ecology, 99, 288–299