UNIVERSITY OF GOTHENBURG
Department of Economy and Society, Human Geography &
Department of Earth Sciences
Geovetarcentrum/Earth Science Centre
Contrasting responses
of shrubs to reindeer
grazing in alpine heath
and meadow grasslands
Two case studies of the Scandes mountain range
Cajsa Lovehav
ISSN 1400-3821
Mailing address
Geovetarcentrum
S 405 30 Göteborg
Address
Geovetarcentrum
Guldhedsgatan 5A
B930
Bachelor of Science thesis
Göteborg 2016
Telephone
031-786 19 56
Telefax
031-786 19 86
Geovetarcentrum
Göteborg University
S-405 30 Göteborg
SWEDEN
Abstract
The purpose of this study was to examine the responses in plant community composition and
shrub growth to absence of reindeer grazing in the Swedish alpine tundra. I hypothesized that
absence of reindeer grazing leads to increased shrub encroachment and a lowering of plant
diversity. Plant community composition was analysed using cover and height data from two
experimental sites in the Scandes; a low herb meadow at Ritsem in the north, and a grass
heath at Långfjället at the southern margin of the mountain range. Plant cover responses to
grazed and ungrazed conditions were analysed through computations of relative cover change
between plant functional types. Plant cover changes over time and grazing pressure
(grazed/ungrazed) were tested for using repeated measures ANOVA. Plant diversity at both
sites were quantified by the use of species richness estimates through the Chao 2 method, as
well as Simpson’s index of diversity, and the Berger-Parker index of dominance. In
agreement with previous syntheses on shrub expansion across the tundra biome, this study
shows that both the meadow at Ritsem and the grass heath at Långfjället were subjected to
rapid shrubification over the last 16 years, a process that is presumably brought about by
climate warming. Reindeer grazing was found to significantly limit the advancement of dwarf
birch (Betula nana) and willow species (Salix spp.) in the low herb meadow at Ritsem, and
diversity was correspondingly ranked higher under grazed conditions at this site. The low
evergreen shrubs, crowberry (Empetrum nigrum) in particular, that are advancing at
Långfjället however, are relatively unpalatable and grazing does not seem to impede their
progression. The differentiated shrub responses to reindeer grazing emphasises the
importance of evaluating the influences of herbivory on distinctive alpine tundra habitats.
Nonetheless, this study deems reindeer husbandry acting as a limiting factor to deciduous
shrub expansion an encouraging prospect.
Keywords
Alpine tundra, reindeer, plant-herbivore interactions, meadow with low herbs, grass heath
2
Table of Contents
Abstract ...................................................................................................................................... 2
Keywords ................................................................................................................................... 2
1. Introduction ............................................................................................................................ 5
2. Background ............................................................................................................................ 7
2.1 The biodiversity concept .................................................................................................. 7
2.2 Tundra vegetation and the alpine zone ............................................................................. 7
2.3 Climate warming and shrub expansion in the tundra biome ............................................ 8
2.4 Plant – herbivore interactions in the tundra ..................................................................... 9
2.5 Reindeer feeding strategies .............................................................................................. 9
2.6 Reindeer husbandry in Sweden ...................................................................................... 10
3. Study areas ........................................................................................................................... 11
3.1 Långfjället grass heath ................................................................................................... 11
3.1.1 Location ................................................................................................................... 11
3.1.2 Air temperature ....................................................................................................... 11
3.1.3 Vegetation composition........................................................................................... 12
3.1.4 Reindeer densities ................................................................................................... 12
3.2 Ritsem meadow with low herbs ..................................................................................... 13
3.2.1 Location ................................................................................................................... 13
3.2.2. Air temperature ...................................................................................................... 13
3.2.3 Vegetation composition........................................................................................... 14
3.2.4 Reindeer densities ................................................................................................... 15
4. Materials and methods ......................................................................................................... 16
4.1 Data and original experimental design ........................................................................... 16
4.2 Data processing .............................................................................................................. 17
4.2.1 Plant functional types .............................................................................................. 17
4.2.2. ANOVA ................................................................................................................. 17
4.2.3 Biological diversity indices ..................................................................................... 17
4.2.3.1 Observed and estimated species richness ......................................................... 18
4.2.3.2 Simpson´s Diversity index ............................................................................... 19
4.2.3.3 Berger-Parker index ......................................................................................... 19
5. Results .................................................................................................................................. 20
5.1 Grass heath vegetation at Långfjället ............................................................................. 20
3
5.1.1 Relative abundance of plant functional types ......................................................... 20
5.1.2 Cover changes over time and distinctions between treatments ............................... 21
5.1.3 Height of shrubs ...................................................................................................... 23
5.1.4 Species diversity ...................................................................................................... 23
5.2 Meadow with low herbs vegetation at Ritsem ............................................................... 25
5.2.1 Relative abundance of plant functional types ......................................................... 25
5.2.2 Cover changes over time and distinctions between treatments ............................... 26
5.2.3 Height of shrubs ...................................................................................................... 28
5.2.4 Species diversity ...................................................................................................... 30
6. Discussion ............................................................................................................................ 31
6.1 Results Discussion.......................................................................................................... 31
6.2 Method discussion .......................................................................................................... 33
6.2.1 Inventory data and cover estimates ......................................................................... 33
6.2.2. Quantifications of diversity .................................................................................... 33
7. Conclusion ............................................................................................................................ 35
Acknowledgements .................................................................................................................. 36
References ................................................................................................................................ 37
4
1. Introduction
The current anthropogenically induced warming of the northern latitudes is occurring at a rate
two times faster than the global average (Callaghan et al., 2013), and northern Fennoscandia
experienced rapid warming already two decades ago (Olofsson et al., 2009). The length of the
growing season has been prolonged (Wilson & Nilsson, 2009) and changes in Arctic and
alpine vegetation have followed (ACIA, 2005; Olofsson, te Beest, & Ericson, 2013). One of
the most evident responses to warming is shrub expansion, which has been documented
across the tundra biome (Callaghan et al., 2013; Chapin et al., 2005; Christie et al., 2015).
Being the largest plant life form on the tundra, an increased shrub cover may have the
potential to fundamentally alter tundra ecosystems through adjoined effects on other living
organisms, as well as the nitrogen and carbon cycles (Vowles et al., 2015 submitted). A
comprehension of the extent to which reindeer husbandry can limit shrubification of different
alpine tundra habitats in Sweden, is needed when decisions are made on the upholding of
national commitments such as the 2013/14:141 bill on a Swedish strategy for biodiversity and
ecosystem services, as well as Sweden´s environmental quality objective on the magnificent
mountain landscape. Studies on reindeer- vegetation interactions have previously been carried
out within Fennoscandia, yet these studies have typically been concerned with northern
Norway and Finland, and the Abisko area in northern Sweden (Vowles et al., 2015
submitted). The immense size of the Scandes mountain range reflect that there are yet many
alpine tundra areas and vegetation types to be accounted for in terms of studies concerned
with reindeer grazing influences on plant communities.
In 1995, the World Wide Fund for Nature (WWF) initiated a reindeer exclosure experiment,
with the intent of studying the effects of reindeer husbandry practices on mountain vegetation
in Sweden. Five study areas were chosen in the Scandes; Poullanvare in the northernmost
part, Ritsem, Sånfjället, and finally Långfjället and Fulufjället at the southern margin of the
mountains. The sites include five different vegetation types; grass heath, dry heath, meadow
with low herbs, mountain birch forest heath with lichens, as well as mountain birch forest
heath with mosses. All study sites are situated within areas important for reindeer husbandry.
This WWF project originated in the wake of a conference entitled ´The Reindeer Grazing
Problem´ that was held in the context of an ongoing debate about the notion that the Swedish
mountains were overutilised through reindeer husbandry practices (Eriksson et al., 2007).
5
Data collected at the permanently marked grazed and ungrazed low herb meadow plots at
Ritsem, and the grass heath plots at Långfjället have been attained, and form the basis of this
study. The data has been analysed to reveal variations in plant cover, plant diversity and
height of shrubs over time. Accordingly, this study aims to examine how plant community
compositions in grass heath and meadow with low herbs vegetation have responded to a longterm exclusion of reindeer between 1995 and 2011/2012, with the main objective to
investigate shrub cover responses to release from herbivory. Reindeer (Rangifer tarandus)
have previously been found to inhibit the extent of shrub expansion in tundra areas (Cohen et
al., 2013; Olofsson et al., 2013), and the hypothesis of this thesis is that reindeer grazing has a
positive influence on plant diversity and impedes shrub expansion on the tundra.
Problem statements:

How does absence of reindeer grazing influence shrub cover, overall plant cover, and
plant diversity within alpine grass heath and meadow with low herbs vegetation?

What changes over time can be identified at the Långfjället grass heath and Ritsem
meadow with low herbs site?
6
2. Background
2.1 The biodiversity concept
Biodiversity refers to the variety among all life at all levels of biological organization, i.e.
within and between species and of ecosystems (Gaston & Spicer, 2004 p.3). It is a
multifaceted concept that can be quantified in a number of ways through diversity indices. A
diversity index is a mathematical expression that combines the two components of diversity;
richness (the number of different species per unit area) and evenness (relative abundance of
all the species within the area) to quantify diversity (Colwell, 2009).
2.2 Tundra vegetation and the alpine zone
Tundra (the treeless ecosystems found in the Arctic and in alpine areas) vegetation typically
forms a mosaic that is intimately linked to local geomorphology and microclimate. Species
richness can be low within each mosaic, but in a regional context the number of species can
be considerable. The vegetation growth rate is generally slow, and it may take decades
before a patch of bare ground becomes colonized. Because of the low productivity of the
vegetation, vast pasture areas are required to support migratory herds of animals such as
reindeer (Rangifer tarandus) (Holden, 2012).
Alpine tundra areas are located in mountain regions around the world, at altitudes where trees
cannot grow (Sundseth, 2006). The alpine zone is traditionally divided into three belts: low
alpine, mid alpine, and high alpine (European Environment Agency, 2002). The low alpine
zone extends vertically from approximately 300 m above the forest line (the uppermost
altitude where trees form a forest with closed canopy) to the point where Vaccinium myrtillus
(blueberry) is no longer a predominant part of the vegetation. The mid-alpine zone ranges
from the upper edge of the low alpine, up to the point where vascular plant vegetation ceases
to be continuous (Sundqvist, Björk, & Molau, 2008). Within the high alpine zone, the
vegetation is scattered and mainly consists of bryophytes (mosses) and lichens (European
Environment Agency, 2002).
7
2.3 Climate warming and shrub expansion in the tundra biome
Climate warming is an important agent in moderating plant compositions in the tundra biome
(Elmendorf et al., 2012), and palaeorecords show that shrub expansion occurred during warm
periods throughout the Holocene epoch (Chapin et al., 2005). In turn, it is recognized that
changes in vegetation at high latitudes can alter the radiative forcing of the tundra through
lowering of surface albedo, thus increasing solar absorption (Swann, Fung, Levis, Bonan, &
Doney, 2010). Furthermore, shrubs are expected to be able to spread quickly throughout
tundra areas in the future, because they trigger feedback loops that primarily exert a positive
influence on their own expansion rate (Sturm et al., 2005) (Figure 1).
Figure 1. The snow-shrub-soil-microbe feedback loop proposed by Sturm et al. Climate warming increases shrub
abundance. Shrubs trap more snow leading to higher soil temperatures, increased mineralization rates
(decomposition of organic substances into forms accessible to plants) and increased soil nitrogen availability that
further encourages shrub growth.
Sturm et al., BioScience 2005 Vol.55 No.1, p. 24, by permission of Oxford University Press
Shrub expansion has already been documented in several studies conducted across the Arctic
and subarctic (Callaghan et al., 2013; Chapin et al., 2005; Christie et al., 2015; Olofsson et
al., 2009; Olofsson et al., 2013). Hence, shrub vegetation could become a major influencing
factor on summer warming in the future (Chapin et al., 2005).
Earlier studies have found tendencies of plant species richness decrease as a response to
shrub increase (Wilson & Nilsson, 2009). Shrub advancement above the tree line (the edge of
the habitat where trees are capable of growing) could also affect ecosystem functions,
biodiversity, and carbon stocks (Brandt, Haynes, Kuemmerle, Waller, & Radeloff, 2013).
8
2.4 Plant – herbivore interactions in the tundra
Grazing is an important driver in ecosystem responses to warming of the tundra (Wilson &
Nilsson, 2009). Herbivores exercise a top-down effect on plant community structures and
disrupt the relationship between summer air temperature and shrub primary production
(Olofsson et al., 2013). Earlier studies have shown that productive tundra vegetation can
respond quickly to release from herbivory in terms of changes in biomass composition
(Ravolainen et al., 2011). The degree to which herbivores can counterbalance the climate
warming effect on the tundra is variable depending on plant and herbivore species pools
however (Christie et al., 2015; Ravolainen et al., 2011).
2.5 Reindeer feeding strategies
Reindeer foraging strategies are not randomized (Danell, Utsi, Palo, & Eriksson, 1994;
Skogland, 1980), but reindeer dietary preferences are rather interchanging throughout the
seasons, depending on plant nutritional value and forage availability (Ophof, Oldeboer, &
Kumpula, 2013). In the alpine zone, the highest level of selectivity has been shown to occur
for productive meadow vegetation with high biodiversity. A lower level of selectivity has
been attributed to various heath vegetation types within the alpine tundra (Skogland, 1980).
Lichens are generally preferred fodder during winter, while vascular plants such as
graminoids (grasses, sedges and rushes) and shrubs become a more considerable part of the
diet as spring progresses (Ophof et al., 2013; Skogland, 1980). Through studying rumen
samples from reindeer from the Hardangervidda in Norway, Skogland (1980) concluded that
forbs (herbaceous flowering plants that are not graminoids) are most preferred by alpine
reindeer during summer. Favored feeding sites include meadow and heath vegetation (Moen,
Boogerd, & Skarin, 2009).
Reindeer have been listed as one of the herbivore species having the greatest influence on
vegetation in subarctic Abisko in the north of Sweden (Callaghan et al., 2013) and browsing
by reindeer has previously been accorded a substantial inhibiting effect on shrub expansion,
especially for deciduous shrubs like Betula nana (mountain birch) (Olofsson et al., 2009;
Post & Pedersen, 2008), a species that has been observed to spread quickly in tundra areas
(Olofsson et al., 2009). In addition to grazing, reindeer affect vegetation through trampling
and deposition of faeces (Vowles et al., 2015 submitted).
9
2.6 Reindeer husbandry in Sweden
All reindeer in Sweden are semi-domesticated (Moen et al., 2009), and the reindeer herding
right applies exclusively to the indigenous Sami. The reindeer herding area encompasses
around 50 % of the land base in Sweden (Figure 2) and is divided into 51 Sami villages
(Sametinget, 2016). The herding right is regulated under the Reindeer Husbandry Act, stating
that members of a Sami Village have the right to herd and graze reindeer unreservedly within
the reindeer herding area, regardless of land ownership ("Rennäringslag 1971:437,"). The
animals are thus free- range and move over vast areas. According to the Sami Parliament, the
total number of reindeer varies between 225 000 and 280 000 animals in winter herd
(Sametinget, 2016). Most of the Scandes mountain range is used for reindeer pasture during
the snow free season, whereas boreal forests located to the east of the mountain range are
used as pasture during winter (Moen et al., 2009)
Figure 2. Map showing the reindeer herding area in Sweden, the alpine region as defined by the European
Commission Habitats Directive, and the two study sites Långfjället and Ritsem (six plots at each location),
located in the Scandes mountain range.
Rennäringens markanvändningsdatabas, IRENMARK, 2016-05-10 © Sametinget
Tärrängkartan and Biogeographical regions © Lantmäteriet
10
3. Study areas
3.1 Långfjället grass heath
3.1.1 Location
The grass heat site is located in Dalarna County at the southern part of the Scandes (62°05'N,
12°25'E) (Figure 2) at Långfjället. The site is situated in the low alpine zone, at an elevation
of 1010 m.a.s.l. (the forest line is at approximately 840 m.a.s.l.). The bedrock consists of
severely weathered, chemically acidic Dala granite and there is poor potential for formation of
a substrate able to support rich vegetation (Eriksson et al., 2007).
3.1.2 Air temperature
Särna is the closest SMHI (Swedish Meteorological and Hydrological Institute)
meteorological station to the Långfjället experimental site, and is located 65 km away from
the study area. In the 24 years between 1991 and 2014, all but two years were warmer
compared to the meteorological standard normal period 1961-1990. The recorded mean air
temperature at Särna station was on average 1.2 °C warmer between 1991-2014 compared to
the reference period (Vowles et al., 2015 submitted).
11
3.1.3 Vegetation composition
The field layer of the grass heath at Långfjället is dominated by the low shrubs Empetrum
nigrum, Phyllodoce caerulea, Vaccinium myrtillus, and Vaccinium vitis-idaea, alongside
graminoids such as Carex bigelowii and Deschampsia flexuosa (Figure 3). The bottom layer
mainly consists of Cladonia and Cetraria lichen species, while the shrub layer is almost
entirely made up of the tall shrub Juniperus communis (common juniper).
Figure 3. Dominant field layer species at Långfjället experimental site.
a) Empetrum nigrum (crowberry)
b) Phyllodoce caerulea (blue heath)
c) Vaccinium myrtillus (blueberry)
d) Vaccinium vitis-idaea (lingonberry)
e) Carex bigelowii (stiff sedge)
f) Deschampsia flexuosa (wavy hair-grass)
3.1.4 Reindeer densities
The Idre Nya Sami village area spans 5477 km2 and encompasses Långfjället experimental
site. The maximum number of reindeer allowed within the area is 2700 animals in winter
herd, which equals a reindeer density of 0.49 animals / km2 (Sametinget, 2015). In order of
reindeer numbers, Idre Nya Sami village is ranked 32nd out of the 51 Sami villages in Sweden
(Vowles et al., 2015 submitted).
12
3.2 Ritsem meadow with low herbs
3.2.1 Location
The low herb meadow experimental site is situated in the northern part of the Scandes
(67°45'N, 17°40'E) (Figure 2) ca. 50 km north of Sarek national park in Norrbotten County.
This site is also located in the low alpine zone, at an elevation of 820 m.a.s.l. (the forest line is
at approximately 600 m.a.s.l.) The underlying bedrock is locally calciferous and consists of
relatively easily weathered, soft micha schist which influences the vegetation (Eriksson et al.,
2007).
3.2.2. Air temperature
Ritsem is the closest SMHI meteorological station to the Ritsem experimental site, and is
located 15 km away from the study area. In the period 1991- 2014, all but one year were
warmer at the Ritsem station compared to the standard normal period 1961-1990. The mean
air temperature was 1.1 °C warmer between 1991-2014 compared to the base period (Vowles
et al., 2015 submitted).
13
3.2.3 Vegetation composition
The vegetation at the low herb meadow at Ritsem mainly consists of graminoids such as
Deschampsia cespitosa, Carex aquatilis and D. flexuosa, that dominate the field layer along
with forb species including e.g. Viola biflora, Thalictrum alpinum, and Saussurea alpina. B.
nana and Salix lapponum make up the main part of the tall shrubs (Figure 4). The bottom
layer is mainly made up of mosses such as Hylocomium splendens (mountain fern moss) and
Pleurozium schreberi (red-stemmed feathermoss).
Figure 4. Dominant field layer species at Ritsem experimental site.
a) Thalictrum alpinum (alpine meadow rue)
b) Saussurea alpina (Alpine saw-wort).
c) Betula nana (dwarf birch)
d) Salix lapponum (Downy willow)
e) Deschampsia cespitosa (tufted hair-grass)
f) Carex aquatilis (water sedge)
g) Viola biflora (alpine yellow-violet)
14
3.2.4 Reindeer densities
According to the Sami Parliament, The Sami village Unna Tjerusj, encompassing 4763 km2,
has the reindeer herding right in the area where the experimental site is located, and the
maximum number of reindeer allowed within the sami village is 7000 animals in winter herd
(Sametinget, 2014). Divided by the size of pasture area, the estimated reindeer density is 1.68
reindeer / km2. Ranked in the order of reindeer populations, Unna Tjerusj is 22nd out of the 51
Sami villages in Sweden (Vowles et al., 2015 submitted).
15
4. Materials and methods
4.1 Data and original experimental design
WWF initiated an exclosure experiment in 1995, with the intention of studying the effects of
reindeer husbandry practices on mountain vegetation in Sweden (see Eriksson et al., 2007 for
full background). Six adjacent 25x25m plots considered to be as similar as possible were
permanently marked in grass heath vegetation at Långfjället, and in low herb meadow
vegetation at Ritsem in 1995. A 1.7 m high fence was erected around three of the plots
(hereafter referred to as fenced) at each site in order to deny large herbivores grazing access
(while small herbivores such as hares and rodents still had full access to the plots). The
remaining three plots at each site were left open to serve as control plots (henceforth termed
ambient), representing natural grazed conditions. At each of the plots, twenty 1x1m subplots
were chosen by random selection and vegetation inventories were performed within these
subplots (Figure 5). After four years when the plots were revisited, marginal or no treatment
effects (difference between fenced and ambient plots) were observed (Eriksson et al., 2007).
Figure 5. Conceptual diagram of the sampling methodology at Långfjället and Ritsem. Three fenced and three
ambient 25x25m plots were permanently marked at each location in 1995. To avoid edge effects, randomized
samples of 1x1m (20 x 1m2 in 1995 and 12 x 1m2 in 2011/2012) were taken within a net area of 22x22m within
the 25x25m plots.
Following the original inventory methodology to enable comparisons through time as well as
between treatments, the plots were inventoried anew in July 2011 (Långfjället) and July 2012
(Ritsem). Within every plot, twelve subplots of 1x1m were chosen by random selection, and
16
the percentage cover of each individual species (within all subplots) was visually estimated,
using folding rulers to mark the edges of the subplots to assist estimations. To avoid edge
effects, a 1.5 m wide strip around the edges of the plots was left out, and all subplots were
selected within a net area of 22x22m (Figure 5). Because plants can stretch over each other,
the cover estimates of each 1x1m subplot (including bare ground) in some instances amount
to more than 100%.
In order to detect changes in the shrub layer on a larger scale, a further inventory of all shrubs
over 30 cm in the plots was undertaken. Species, height and two perpendicular measurements
of canopy diameter (for a rough approximation of shrub area) were recorded for each
individual shrub.
4.2 Data processing
4.2.1 Plant functional types
To study the vegetation responses to release from, as well as exposure to reindeer grazing, all
plants were divided into seven plant functional types (growth forms); dwarf shrubs, low
shrubs, tall shrubs, graminoids (grasses, sedges and rushes), forbs (herbaceous flowering
plants that are not graminoids), lichens, and mosses.
4.2.2. ANOVA
Analysis of variance (ANOVA) was used to test for significant cover difference between the
years and between treatments at each site. Year, site, and treatment were used as fixed factors
and plot as random factor. The results are given as statistically significant if P < 0.05. The
statistical analyses were executed using the R statistical package.
4.2.3 Biological diversity indices
In order to study changes in plant community structure over time, the reciprocal forms of the
species diversity indices Simpson’s D and the Berger- Parker index, together with species
richness and estimated true species richness (Chao 2) were calculated based on cover data for
the years 1995 and 2011 for Långfjället and for 1995 and 2012 for Ritsem. The fenced plots at
each site were analysed as one, and the ambient plots at each site were analysed as one. It was
not possible to locate one of the three original ambient plots at Ritsem in 2012 however, thus
a new ambient plot was marked at the location that same year. For this reason, only two
ambient plots were used to quantify diversity in 2012 at the Ritsem ambient plots.
17
The Simpsons 1/D index and the Berger- Parker 1/d index were calculated using mean values
for each treatment and year. Species richness is a count of all species within the sample for
each treatment and year. Due to difficulties in identifying lichen and moss species in the field,
together with the associated risk of identification disparities between the first and the second
inventory, calculations were performed using data on the field layer (vascular plants) only.
4.2.3.1 Observed and estimated species richness
The number of different species within the samples (observed species richness) was noted for
fenced and ambient plots respectively, to establish if the number of species differed between
treatments and between the years. Additionally, in order to determine whether the sampling
from all inventories had been extensive enough to include all species present at both sites, an
estimation of the absolute species richness (the total number of species expected to be present
within the plots) for each treatment was calculated. This was done by first converting the
cover data to presence/absence data, and then using the Chao 2 method according to
Magurran (2004). Chao 2 is based on the assumption that as long as rare species are still being
discovered, it is likely that there are more species in the locality that have yet not been
accounted for. When all species have been observed in at least two samples, it is likely that all
species within the study area have been found. Chao 2 is summarized in the equation:
𝑄 2
𝑆𝐶ℎ𝑎𝑜2 = 𝑆𝑜𝑏𝑠 + 2𝑄1
(1)
2
where Sobs is the total number of species observed within the sample assemblage. Q1 is the
number of species occurring in only one sample (singletons), and Q2 is the total number of
species occurring in exactly two samples (doubletons).
18
4.2.3.2 Simpson´s Diversity index
The reciprocal form of Simpson’s Diversity index (1/D) gives the probability that two
individuals drawn at random from a finite community will belong to the same species. The
index is not sensitive to richness, but heavily weighted towards the most abundant species.
The index does not require a large sample size to be used as an indicator of diversity
(Magurran, 2004), meaning that it is applicable to estimating diversity from the inventories of
Långfjället and Ritsem, despite the sample sizes being fairly small. The reciprocal form of the
index is summarized in the equation:
𝑛𝑖(𝑛𝑖 − 1)
𝐷 = 1/ ∑ (
)
𝑁(𝑁 − 1)
(2)
where ni is the total number of organisms belonging to the ith species, and N is the total
number of organisms of all species within the sample. With this formula, diversity (D) will
increase as the assemblage of species becomes more evenly distributed i.e when there is a
reduction in dominance.
4.2.3.3 Berger-Parker index
The Berger-Parker index (d) is a pure dominance measure and expresses the proportional
abundance of the most abundant species within a sample. A reduction in dominance and
higher species diversity accompany higher values of the index when using the reciprocal
form:
𝑑 = 1/(
𝑁𝑚𝑎𝑥
𝑁
)
(3)
where Nmax represents the number of individuals within the most abundant species in the
sample, and N is the total number of species within the sample.
19
5. Results
5.1 Grass heath vegetation at Långfjället
5.1.1 Relative abundance of plant functional types
The lichens and low shrubs together had the highest relative abundance in 1995 as well as in
2011 (Figure 6). Out of the seven functional types, these two also exhibited the largest
changes in relative abundance from 1995 compared to 2011. Lichens were the most abundant
of all the functional types at the baseline inventory in 1995, but had declined within both
fenced and ambient plots by 2011. Low shrubs on the other hand doubled in relative
abundance from 1995 to 2011 for both treatments. Thus, there was a shift from lichens being
most abundant in 1995 to low shrubs being the most abundant in 2011. Changes in relative
abundance of the other plant functional types were overall of minor importance.
Figure 6. Change in relative abundance of each plant functional type within the grass heath experimental site at
Långfjället (Note that forbs were still present in 2011, but had a relative cover of 0.5% within the ambient plots
and 0.1% within the fenced plots and are thus not shown).
20
5.1.2 Cover changes over time and distinctions between treatments
The general cover changes between the years were negligible for all functional types apart
from the low shrubs that had a significant cover increase within both fenced and ambient plots
between 1995 and 2011 (Figure 7). The mean cover increase of low shrubs across treatments
was 138 % (P < 0.001).
***
Figure 7. Low shrub cover change over time (mean per cent cover ± standard error) for both treatments at
Långfjället grass heath. Black circles represent fenced plots and red circles represent ambient plots. Stars denote
significant differences over time between 1995 and 2011 (*** P<0.001).
21
The increase in low shrub cover was mainly attributed to a progressed distribution of
evergreen species, which increased their cover by 125% from 1995 to 2011 (Figure 8). The
evergreens E. nigrum, P. caerulea, and V. vitis-idaea all increased in abundance from 1995 to
2011. The greatest advancement was in E. nigrum, which more than doubled its cover from
15% in 1995 to 33% in 2011. V. myrtillus was the only deciduous low shrub at Långfjället. It
had a small total cover extent, but remarkably quintupled its cover from 1 to 5% between the
years. The V. myrtillus cover increase was nearly the same within both fenced and ambient
plots.
Figure 8. Cover change over time (mean per cent cover ± standard error) per species of low shrubs at Långfjället.
Gray bars represent values in 1995 and white bars represent values in 2011. E. nigrum, P. caerulea, and V. vitisidea are evergreens and V. myrtillus is deciduous. C. vulgaris declined from 0.1% to 0.0% from 1995 to 2011
and is not shown.
22
Lichens were more abundant within the ambient plots than the fenced ones in 1995 (Figure 9).
By 2011, this distinction between treatments had been smoothed out and the lichen abundance
was nearly at equal level between treatments. This was due to a minor cover increase within
the exclosures, whereas the lichen cover was practically unchanged within the ambient plots
over the years. There were no significant treatment effects on any of the functional types in
2011.
Figure 9. Lichen cover change over time (mean per cent cover ± standard error) for both treatments at
Långfjället grass heath. Black circles represent fenced plots and red circles represent ambient plots. Stars denote
significant differences between treatments (** P=0.001-0.01).
5.1.3 Height of shrubs
The tall shrub layer at Långfjället consisted of only J. communis in 2011. Out of a total of
nine specimens within all plots (fenced and ambient), only one was taller than 30 cm (not
shown). There were more J. communis within the ambient plots, but the mean height differed
minimally between treatments. Grazing was not of importance since J. communis is an
unpalatable species.
5.1.4 Species diversity
Most species that are estimated to occur within the plots were accounted for at both the
sampling intensities of 20x1m2 (1995) and 12x1m2 (2011) (Table 1). This signals that it is
23
possible to draw conclusions on diversity changes over time, despite that sampling was
performed with unequal effort.
The overall species richness was modest both in 1995 and 2011, and the highest number of
vascular plant species found was 15, encountered within the ambient plots in 1995. Within
these plots, two plant species; the dwarf shrub Cassiope hypnoides and the tall shrub B. nana
were scarce in 1995, but their cover seem to have become even more meagre by 2011, and the
species were not accounted for in the data set from that year, resulting in 13 species found
within the ambient plots in 2011. The species richness was the same within the fenced plots
through the years (14 species).
At the outset of the baseline inventory, the plant diversity in terms of community evenness
was lower within the plots that were to be enclosed, compared to the plots to remain open.
That is to say that there were natural diversity differences between the plots already before the
fences had been erected. By 2011 however, these differences had been reduced resulting in a
more uniform plant community despite differentiated treatments (fenced vs. ambient).
Table 1. Diversity indices Simpson’s 1/D, Berger-Parker 1/d, observed species richness, and estimated true
species richness (Chao 2). All fenced plots are analysed as one, and all ambient plots are analysed as one. The
indices 1/D and 1/d are derived from mean values for each treatment and year with ± standard error. Species
richness is a count of all species within the sample per treatment and year. Only vascular plants are included.
Year Plots Samples (m2)
Treatment Simpsons 1/D
1995
3
60
Fenced
2.68 ± 0.36
1995
3
60
Ambient
2011
3
36
2011
3
36
1.85 ± 0.18
Observed
richness
14
Estimated
richness
15
2.85 ± 0.32
1.94 ± 0.20
15
15
Fenced
2.54 ± 0.45
1.89 ± 0.27
14
15
Ambient
2.56 ± 0.30
1.82 ± 0.14
13
13
24
Berger- Parker 1/d
5.2 Meadow with low herbs vegetation at Ritsem
5.2.1 Relative abundance of plant functional types
The functional types that had the greatest relative abundance at Ritsem in both years were
graminoids and forbs (Figure 10). The bottom layer was dominated by moss. Graminoids had
the highest relative abundance in both years, but exhibited a minor cover decline within both
fenced and ambient plots from 1995 to 2012. Forb relative abundance had dropped by 50 per
cent in ambient plots by 2012, while relative abundance remained essentially unchanged
within the fenced plots. Moss cover had increased markedly from 21% in 1995 to over 30%
over both treatments by 2012.
Figure 10. Change in relative abundance of each plant functional type within the low herb meadow experimental
site at Ritsem (Note that tall shrubs were present in 1995, but had a relative cover of 0.2% and are not shown)
25
5.2.2 Cover changes over time and distinctions between treatments
In general terms, the cover changes over time were minor at Ritsem. But there was a trend (P
= 0.09) towards increase in moss cover between 1995 and 2012 (Figure 11). Mosses increased
with 105% on average (from 20% to 41%) over time. All the other plant functional types
demonstrated minor changes in cover distribution over time.
Figure 11. Moss cover change over time (mean per cent cover ± standard error) for both treatments at Ritsem
low herb meadow. Green triangles represent fenced plots and yellow triangles represent ambient plots.
26
There was a significant difference in forb cover between treatments in 2012 (Figure 12). The
average per cent increase in forb cover over time was 38% within the exclosures, as opposed
to a forb cover decrease of 41% within the ambient plots. Thus in 2012, the difference
between treatments was significant with a mean forb cover of 33% (P < 0.01) within the
exclosures, and 16% (P < 0.01) within the ambient plots. This variation was largely attributed
to a high abundance of S. alpina (see Figure 4b) within the fenced plots, whereas its
abundance was much lower within the ambient plots, indicative of it being a grazed species.
Apart from the forbs, no functional types indicated a significant treatment effect in neither
1995 nor 2012.
Figure 12. Forb cover change over time (mean per cent cover ± standard error) for both treatments at Ritsem low
herb meadow. Green triangles represent fenced plots and yellow triangles represent ambient plots.Stars denote
significant differences between treatments (*** P<0.001).
27
5.2.3 Height of shrubs
Even though it is not shown in the cover data, the difference between the heights of the tall
shrubs varied substantially between fenced and ambient plots when they were inventoried in
2012 (Figure 13). The average height of the tall shrubs, as well as the mean maximum height
of the tallest individual shrub within each fenced plot was markedly greater than the height of
shrubs within the ambient plots. On average, the shrubs within the three fenced plots
combined were twice as tall (35 cm) as the mean height (17 cm) of all shrubs within the
ambient plots (Figure 13a). The mean maximum height of shrub was three times higher
within fenced plots compared to ambient plots (Figure 13b) The tallest individual within the
fenced plots (144 cm) were more than four times the height of the tallest individual within the
ambient plots (32 cm) (not shown). The area of the shrubs within the fenced plots was also
considerably larger than the area of the shrubs within the ambient plots (not shown).
Figure 13. Treatment effects on height of tall shrubs at Ritsem in 2012. Green bars represent fenced plots and
yellow bars represent ambient plots. Stars denote significant differences between treatments (* P=0.01-0.05).
a) Mean height of all individuals of shrub (± standard error) within fenced and ambient plots.
b) Mean maximum height of the tallest individual shrub (± standard error) within fenced and ambient plots.
28
The tall shrub cover was mainly made up of B. nana and Salix spp., and both were
particularly favored by exclusion of reindeer. Sorted by species, Salix spp. was the most
favored from exclusion and exhibited a significant treatment effect for both average height
(P<0.01) and the maximum height (P<0.05) (Figure 14) of shrub.
Figure 14. Treatment effects on height of Salix spp. at Ritsem in 2012. Green bars represent fenced plots and
yellow bars represent ambient plots. Stars denote significant differences between treatments
(* P=0.01-0.05, ** P=0.001-0.01).
a) Mean height of all individuals of Salix spp. (± standard error) within fenced and ambient plots.
b) Mean maximum height of the tallest individual Salix spp. (± standard error) within fenced and ambient plots.
The mean heights of B. nana also differed significantly (P<0.05) between fenced and ambient
plots. Although not significant, there was also a difference between the maximum heights of
the B. nana shrubs between treatments (Figure 15).
Figure 15. Treatment effects on height of Betula nana at Ritsem in 2012. Green bars represent fenced plots and
yellow bars represent ambient plots. Stars denote significant differences between treatments (* P=0.01-0.05).
a) Mean height of all individuals of B. nana (± standard error) within fenced and ambient plots.
b) Mean maximum height of the tallest individual B. nana (± standard error) within fenced and ambient plots.
29
5.2.4 Species diversity
Within the fenced plots, 75 species were recorded in 2012, which was nine species fewer than
what was found in 1995 (Table 2). It also seems that there might have been more species to
account for within the exclosures in 2012 compared to 1995. The number of species within
the ambient plots was estimated to be 105 both in 1995 and 2012, despite that the sampling
effort was not uniform between the years.
A diversity difference can be recognized between treatments for both inventory years. Both
the Simpson´s and the Berger-Parker indices showed higher values for ambient than for
fenced plots in 1995 and 2012. Hence, the dominance of the most abundant species in the
samples was higher at the fenced plots at both inventories, i.e. ambient plots displayed a
greater evenness in the plant community than the fenced plots in both years. As previously
mentioned, the difference between treatments was expressed already at the outset of the
baseline inventory. Nevertheless, the difference between treatments signals a greater disparity
between fenced and ambient plots in 2012 than in 1995 (note that the variance between the
plots in 2012 is bigger however). This indicates that plant community composition had
become more distinguishable between treatments up until the time of the most recent
inventory.
Table 2. Diversity indices Simpson’s 1/D, Berger-Parker 1/d, observed species richness, and estimated true
species richness (Chao 2). All fenced plots are analysed as one, and all ambient plots are analysed as one. The
indices 1/D and 1/d are derived from mean values for each treatment and year with ± standard error. Species
richness is a count of all species within the sample per treatment and year. Only vascular plants are included.
Year Plots Samples (m2) Treatment Simpsons 1/D Berger- Parker 1/d
1995
3
60
Fenced
7.61 ± 0.10
3.69 ± 0.38
Observed
richness
84
1995
3
59
Ambient
7.79 ± 0.32
3.74 ± 0.28
89
105
2012
3
36
Fenced
5.68 ± 0.45
3.13 ± 0.29
82
100
2012
2
24
Ambient
6.18 ± 1.77
3.33 ± 0.59
75
105
30
Estimated
richness
90
6. Discussion
6.1 Results Discussion
A contrasting development on shrub cover over the years was found at the two grasslands
investigated. At Långfjället, the low shrub cover increased extensively between 1995 and
2011, whereas at Ritsem there were no significant changes in shrub cover over the years, but
in stead, the height of the tall shrubs had increased. The shrub expansion at both sites
presumably occurred in response to the air temperature increase by 1°C that have occurred in
both areas over the last two decades (Vowles et al., 2015 submitted). At Långfjället, the low
shrubs, E. nigrum in particular, advanced within both fenced and ambient plots, but there
were no significant treatment effects i.e. there was no indication of reindeer grazing limiting
this progression. The shrub layer at Ritsem however, is mainly made up of the deciduous tall
shrubs B. nana and Salix spp., both of which were highly favored from reindeer exclusion.
Within the exclosures, the tall shrubs were at least twice as tall as within the ambient plots.
The tallest individual was four times taller than its counterpart within the ambient plots. This
shrub expansion is in line with earlier studies reporting on increasing shrubification across the
Arctic tundra in response to climate warming (see e.g. Callaghan et al., 2013; Chapin et al.,
2005). The increase in evergreen shrub species at Långfjället seems to be at odds with
observations of deciduous shrubs leading the expansion throughout the tundra biome (Christie
et al., 2015). The results are concurrent however, with earlier findings from an inventory of
dry heath and mountain birch forest plots established along the Scandes (Poullanvare, Ritsem,
Långfjället and Fulufjället) (Vowles et al., 2015 submitted). Further, the limited influence of
reindeer grazing on the grass heath is corresponding with another study conducted in the
central part of the Scandes concluding that reindeer grazing had a minor effect on grass heath
vegetation (Moen et al., 2009). At this location however, the dominance of grazing-tolerant
graminoid species was deemed a plausible reason for the inconsiderable effect of herbivory
on species composition. At Långfjället, the conceivable reason behind the minor grazing
effects is that relatively unpalatable evergreen shrubs unfavored by reindeer (Danell et al.,
1994), are dominating the field layer. The advancement of deciduous shrubs at Ritsem, in
particular in response to absence from grazing, is in line with previous findings of reindeer
grazing limiting deciduous shrub expansion (Olofsson et al., 2009; Post & Pedersen, 2008).
31
At Ritsem, forbs were the only functional type that responded to differentiated treatments in
terms of cover extent. The forb cover within the exclosures remained practically the same
from 1995 to 2012, as opposed to the ambient plots where forb cover declined, resulting in a
significant difference between treatments in 2012. This finding signifies that forb cover did
not increase in response to release from reindeer grazing, further implying that grazing
pressure conceivably increased from 1995 to 2012 in this area. The cover difference between
treatments was markedly pronounced for S. alpina, highly abundant within the exclosures,
while its cover was clearly smaller at the ambient plots. At Långfjället, lichens were more
abundant within the ambient plots than the fenced ones at the baseline measurement in 1995,
but by the time of the 2011 inventory, there were no longer any noteworthy differences
between treatments. The lichen cover at Långfjället seems to have been favored by exclusion
of reindeer, indicated by a slight cover increase within the fenced plots over the years. This
development would be expected since lichens form a big part of reindeer diet (Ophof et al.,
2013; Skogland, 1980). Similarly, forbs have formerly been recognized as favored forage by
reindeer (Skogland, 1980) and reindeer grazing has been found to limit forb growth
(Ravolainen et al., 2011). Thus, it is likely that the significant reduction in forb cover within
the ambient plots at Ritsem by 2012 is attributed to grazing. It cannot be ruled out however,
that an abiotic factor such as the air temperature increase may have changed the competitive
interactions between functional types, so that some other species have increased at the
expense of the forbs over the years. The low shrubs at the ambient plots for instance, could
have contributed to the forb decline.
There was a decrease in diversity at the ambient plots at Långfjället between 1995 and 2011.
Plant community richness remained the same within the exclosures, but there was an
inconsistency in the biodiversity measures for these plots, making it hard to draw conclusions
on changes over time in the community evenness. At Ritsem on the other hand, the field layer
diversity in terms of evenness was higher at the ambient plots than the exclosures both in
1995 and 2012, but the differences between treatments had been enhanced over the years. The
estimated richness remained the same at the ambient plots at Ritsem, meaning that this change
did not originate from a change in species pools, but rather a reduction in dominance of the
most abundant species. On the contrary it seems that the species assemblage within the fenced
plots had become less diverse in terms of evenness by 2012.
32
The fact that diversity (in terms of evenness) was higher at the ambient plots at Ritsem might
be occasioned by that the height of shrubs was limited by grazing. For the same reason, the
decrease in species richness as well as diversity at the ambient plots at Långfjället over the
years could possibly be a response to the shrub cover increase. Such a relationship has
previously been established in earlier studies (Wilson & Nilsson, 2009). The fact that the two
grassland types responded differently to herbivory, despite that reindeer densities are
comparable between the two areas, accentuates the importance of considering local plant and
herbivore species pools to gain an overview of grazing effects on the mosaic vegetation
landscape of the alpine.
6.2 Method discussion
6.2.1 Inventory data and cover estimates
Species cover estimates are a useful way to quantify plant species abundance because it can
be measured directly in the field. One challenge in using this abundance measure however is
that organisms can overlap, making it hard to distinguish species that are hidden by other
species. This type of cover estimate is typically expressed as individual species percentage
cover of the area of study (Magurran, 2004), and was the chosen inventory methodology that
this study is based on. An alternative approach to data collection of plant assemblages is to
measure biomass. This method is universally applicable, but is invasive and demands that
plants are harvested to enable measurements. The best sampling approach is usually to
standardize the sample size and apply this to every assemblage included in the study
(Magurran, 2004). This was the approach taken at all vegetation inventories. A source of error
is that different people performed sampling over the years. Even though the same sampling
methodology was used, the cover estimate values were consistently somewhat higher in
2011/2012 compared to 1995.
6.2.2. Quantifications of diversity
The number of species recorded within a locality is almost always an underestimate of the
true species richness within that locality, because rare species are often subjected to
undersampling. When sampling is arranged in such a way that smaller areas that lie within
larger areas are used to account for the characteristics of the habitat, the smaller areas may be
too small (or inadequately sampled) to account for the species richness within the habitat they
represent. The use of the Chao 2 richness estimator helps correct this undersampling bias
33
(Colwell, 2009). It also has the advantages that estimations can be based on presence/ absence
data, and it is intuitive and easy to calculate. By calculating Chao 2 it was confirmed that the
sample sizes from both inventories at Långfjället were adequate, while it established that the
chosen sample sizes were too small to be able to account for the species richness at the more
species- rich low herb meadow site at Ritsem in 1995 and 2012 respectively. Therefore the
resulting biodiversity values derived from the Ritsem data must be viewed in light of existing
uncertainties as to how comprehensive the data set is.
When comparing different diversity indices it is possible that species assemblages are ranked
in different ways, meaning that the conclusions drawn on which assemblage is most diverse,
may depend on what index that was used (Magurran, 2004). Some biodiversity estimates are
however more sensitive to sample size than others. Naturally, species richness for instance is
vulnerable to variations in sampling effort. Simpson’s D index on the other hand performs
well in this respect because it is weighted towards evenness and ranks assemblages
consistently, even if sample sizes vary (Magurran, 2004).
There are a number of ways of estimating species diversity and the most frequently used
diversity indices combine the two aspects of diversity: evenness and richness. In alpine areas
where many of the plant species are specialized to thrive in harsh environments, plant
community evenness is a point of interest. When the focus is on the dominance of species in a
community rather than on rare and sensitive species, the Simpson’s index that is weighted
toward evenness would be a preferred way to quantify diversity (Nagendra, 2002). The
Berger-Parker index being solely dependent on dominance is also useful in this sense. The
Simpson’s index and the Berger-Parker indices are both stated to be widely used and are
considered two of the most robust biodiversity indices available (Magurran, 2004).
Dividing plants into functional types has previously been suggested as an advantageous way
of studying differentiated vegetation responses to environmental changes (Chapin FS III,
1996; Christie et al., 2015). However, the bottom layer (lichens and mosses) was excluded
from biodiversity calculations, because of difficulties with species identification in the field
and the associated risk of differentiated identifications between the years. Therefore,
statements on plant diversity at any of the sites could only be made about the field layer.
34
7. Conclusion
Shrubification was found to be occurring rapidly at both the northern and southern study site,
which is in line with earlier studies from across the tundra biome. Tall deciduous shrubs such
as B. nana and Salix spp., have advanced in the low herb meadow at Ritsem, and low
evergreen shrubs, E. nigrum in particular have advanced at Långfjället. The hypothesis was
that absence of reindeer grazing would support shrub encroachment and lead to a lowering of
diversity, and at the more productive low herb meadow at Ritsem, tall deciduous shrubs did
advance significantly in the absence of grazing. Herbivory also had an influence on plant
cover and diversity in the meadow, where ambient plots exhibited higher diversity than the
exclosures. In the nutrient-poor grass heath at Långfjället however, the greatest shrub increase
was seen in low evergreen species irrespective of grazing, and no clear tendency of lower
diversity within the exclosures was found. Since alpine tundra areas consist of a mosaic of
vegetation types and reindeer migrate over vast areas, further studies are advocated to be
complimented by satellite navigation data from GPS- (Global Positioning System) collared
reindeer, in order to assist grazing pressure estimates, and extend the understanding of
reindeer husbandry effects on differentiated tundra habitats. Nevertheless, this study, along
with former reports, shed light on the prospect of reindeer husbandry providing an important
regulatory ecosystem service through mitigation of the ongoing deciduous shrub expansion on
the tundra.
35
Acknowledgements
I would like to express my great appreciation to my supervisors Assoc. Prof. Björk and PhD
Student Vowles, who with seemingly infinite patience guided me through the field of
ecology, as well as helping me with a whole range of practicalities throughout the course of
this project. I could have asked for nothing more from their dedicated involvement. I would
like to acknowledge Prof. Thorsson for improving this paper through her constructive
feedback on my sample texts. I would like to thank Mr. A. Anderberg, Ms. Anderberg, and
Mr. R. Anderberg for permitting me to publish their photos in this study. My thanks are also
extended to Ms. Føllesdal and Mr. Ellefsen for their reviews of this paper. Finally, I wish to
express my gratitude to Mr. Tveitane for his unconditional and unwavering support- it made
all the difference.
36
References
ACIA. (2005). Arctic climate impact assessment- scientific report Arctic Tundra and Polar
Desert Ecosystems. New York: Cambridge University Press.
Brandt, J. S., Haynes, M. A., Kuemmerle, T., Waller, D. M., & Radeloff, V. C. (2013). Regime
shift on the roof of the world: Alpine meadows converting to shrublands in the
southern Himalayas. Biological Conservation, 158, 116-127.
doi:10.1016/j.biocon.2012.07.026
Callaghan, T. V., Jonasson, C., Thierfelder, T., Yang, Z., Hedenas, H., Johansson, M., . . .
Sloan, V. L. (2013). Ecosystem change and stability over multiple decades in the
Swedish subarctic: complex processes and multiple drivers. Philos Trans R Soc
Lond B Biol Sci, 368(1624), 20120488. doi:10.1098/rstb.2012.0488
Chapin, F. S., 3rd, Sturm, M., Serreze, M. C., McFadden, J. P., Key, J. R., Lloyd, A. H., . . .
Welker, J. M. (2005). Role of land-surface changes in arctic summer warming.
Science, 310(5748), 657-660. doi:10.1126/science.1117368
Chapin FS III, B.-H. S., Hobbie SE. (1996). Plant functional types as predictors of transient
responses of Arctic vegetation to global change. Journal of Vegetation Science, 7,
347–358.
Christie, K. S., Bryant, J. P., Gough, L., Ravolainen, V. T., Ruess, R. W., & Tape, K. D. (2015).
The Role of Vertebrate Herbivores in Regulating Shrub Expansion in the Arctic: A
Synthesis. BioScience, biv137. doi:10.1093/biosci/biv137
Cohen, J., Pulliainen, J., Ménard, C. B., Johansen, B., Oksanen, L., Luojus, K., & Ikonen, J.
(2013). Effect of reindeer grazing on snowmelt, albedo and energy balance based
on satellite data analyses. Remote Sensing of Environment, 135, 107-117.
doi:10.1016/j.rse.2013.03.029
Colwell, R. K. (2009). Biodiversity: Concepts, Patterns, and Measurement. In S. R. C.
Simon A. Levin, H. Charles J. Godfray, Ann P. Kinzig, Michel Loreau, Jonathan B.
Losos, Brian Walker, David S. Wilcove, associate editors (Ed.), The Princeton
Guide to Ecology (pp. 257-263). Retrieved from
http://press.princeton.edu/chapters/s3_8879.pdf.
Danell, K., Utsi, P. M., Palo, R. T., & Eriksson, O. (1994). Food Plant Selection by Reindeer
during Winter in Relation to Plant Quality. Ecography, 17(2), 153-158.
Elmendorf, S. C., Henry, G. H. R., Hollister, R. D., Björk, R. G., Boulanger-Lapointe, N.,
Cooper, E. J., . . . Wipf, S. (2012). Plot-scale evidence of tundra vegetation change
and links to recent summer warming. Nature Climate Change, 2(6), 453-457.
doi:10.1038/nclimate1465
Eriksson, O., Niva, M., Caruso, A., Institutionen för biologisk, g., Uppsala, u., Biologiska, s.,
. . . Teknisk-naturvetenskapliga, v. (2007). Use and abuse of reindeer range.
European Environment Agency. (2002). Europe’s biodiversity – biogeographical regions
and seas Biogeographical regions in Europe The Alpine region – mountains of
Europe Retrieved from
http://www.eea.europa.eu/publications/report_2002_0524_154909/biogeograp
hical-regions-in-europe/alpine.pdf/view.
Gaston, K. J., & Spicer, J. (2004). Biodiversity an Introduction. United Kingdom
Cornwall: Blackwell Publishing.
Holden, J. (2012). An introduction to Physical Geography and the Environment (Vol. 3rd).
England
Harlow: Pearson.
37
Magurran, A. E., -. (2004). Measuring Biological Diversity. United Kingdom Oxford ::
Blackwell.
Moen, J., Boogerd, C., & Skarin, A. (2009). Variations in mountain vegetation use by
reindeer (Rangifer tarandus)
affects dry heath but not grass heath. Journal of Vegetation Science, 20, 805–813.
Nagendra, H. (2002). Opposite trends in response for the Shannon and Simpson indices
of landscape diversity. Applied Geography, 22, 175–186.
Olofsson, J., Oksanen, L., Callaghan, T., Hulme, P. E., Oksanen, T., & Suominen, O. (2009).
Herbivores inhibit climate-driven shrub expansion on the tundra. Global Change
Biology, 15(11), 2681-2693. doi:10.1111/j.1365-2486.2009.01935.x
Olofsson, J., te Beest, M., & Ericson, L. (2013). Complex biotic interactions drive longterm vegetation dynamics in a subarctic ecosystem. Philos Trans R Soc Lond B Biol
Sci, 368(1624), 20120486. doi:10.1098/rstb.2012.0486
Ophof, A. A., Oldeboer, K. W., & Kumpula, J. (2013). Intake and chemical composition of
winter and spring forage plants consumed by semi-domesticated reindeer
(Rangifer tarandus tarandus) in Northern Finland. Animal Feed Science and
Technology, 185(3-4), 190-195. doi:10.1016/j.anifeedsci.2013.08.005
Post, E., & Pedersen, C. (2008). Opposing plant community responses to warming with
and without herbivores. Proc Natl Acad Sci U S A, 105(34), 12353-12358.
doi:10.1073/pnas.0802421105
Ravolainen, V. T., Bråthen, K. A., Ims, R. A., Yoccoz, N. G., Henden, J.-A., & Killengreen, S. T.
(2011). Rapid, landscape scale responses in riparian tundra vegetation to
exclusion of small and large mammalian herbivores. Basic and Applied Ecology,
12(8), 643-653. doi:10.1016/j.baae.2011.09.009
Rennäringslag 1971:437.
Sametinget. (2014, 2014-04-28). Unna tjerusj. Retrieved from
https://www.sametinget.se/unnatjerusj
Sametinget. (2015, 2015-05-22). Idre Nya sameby. Retrieved from
https://www.sametinget.se/8836
Sametinget. (2016, 2016-01-18). Rennäringen i Sverige. Retrieved from
https://www.sametinget.se/rennaring_sverige
Skogland, T. (1980). Comparative Summer Feeding Strategies of Arctic and Alpine
Rangifer. Journal of Animal Ecology, 49(1), 81-98. doi:10.2307/4278
Sturm, M., Schimel, J., Michaelson, G., Welker, J. M., Oberbauer, S. F., Liston, G. E., . . .
Romanovsky, V. E. (2005). Winter Biological Processes Could Help Convert Arctic
Tundra to Shrubland. BioScience, 55(1), 17-26. doi:10.1641/00063568(2005)055[0017:WBPCHC]2.0.CO;2
Sundqvist, M. K., Björk, R. G., & Molau, U. (2008). Establishment of boreal forest species
in alpine dwarf-shrub heath in subarctic Sweden. Plant Ecology & Diversity, 1(1),
67-75. doi:10.1080/17550870802273395
Sundseth, K. (2006). European Commission
Natura 2000 in the Alpine region Retrieved from
http://ec.europa.eu/environment/nature/info/pubs/docs/brochures/nat2000_
alpine.pdf
Swann, A. L., Fung, I. Y., Levis, S., Bonan, G. B., & Doney, S. C. (2010). Changes in Arctic
vegetation amplify high-latitude warming through the greenhouse effect. Proc
Natl Acad Sci U S A, 107(4), 1295-1300. doi:10.1073/pnas.0913846107
38
Vowles, T., Molau, U., Gunnarsson, B., Hickler, T., Klemedtsson, L., & Björk, R. (2015).
Shrub expansion in the Scandes over 16 years but reindeer grazing of minor
importance. Manuscript submitted for publication.
Wilson, S. D., & Nilsson, C. (2009). Arctic alpine vegetation change over 20 years. Global
Change Biology, 15(7), 1676-1684. doi:10.1111/j.1365-2486.2009.01896.x
39