Linköping Studies in Science and Technology
Dissertation thesis No. 1471
Studies on spatial and temporal
distributions of epiphytic
lichens
Håkan Lättman
School of Life Science
Södertörn University
SE-141 89 Huddinge, Sweden
Department of Physics, Chemistry and Biology
Division of Ecology
Linköping University
SE-581 83 Linköping, Sweden
Linköping, October 2012
© Håkan Lättman 2012
Linköping Studies in Science and Technology
ISBN 978–91–7519–810–1
ISSN 0345–7524
Printed by LiU-Tryck
Linköping, Sweden, 2012
II
Table of contents
LIST OF PAPERS
IV
MY CONTRIBUTIONS TO THE PAPERS
IV
ABSTRACT
V
INTRODUCTION
1
WHAT IS A LICHEN?
DISPERSAL STRATEGIES
ENDURING HARSH ENVIRONMENTS, YET ALSO SENSITIVE
SUNLIGHT
THE AIR
CLIMATE CHANGE: TEMPERATURE AND MOISTURE
SUBSTRATE
2
2
4
4
5
6
7
AIMS OF THE THESIS
8
FURTHER BACKGROUND AND THE INCLUDED PAPERS
9
GROWTH
GENERATION TIME OF LICHENS
Paper I
IS SUBSTRATE OR DISPERSAL LIMITING?
Phorophyte and stand
Spore dispersal
Paper II
LARGE-SCALE DYNAMIC OF LICHENS
Dynamics of lichen thalli
Lichens on the move
Paper III
LICHENS IN THE URBAN ENVIRONMENT
Trees: an important urban element
Urban effects on lichens
Paper IV
9
10
11
12
12
12
13
16
16
17
17
19
20
21
21
CONCLUDING REMARKS
25
POPULÄRVETENSKAPLIG SAMMANFATTNING
26
VAD ÄR EN LAV?
RESULTAT FRÅN MIN FORSKNING
26
26
ACKNOWLEDGEMENT
28
REFERENCES
29
III
List of papers
The following papers are included in the thesis and are referred in the text by their
Roman numerals:
Paper I
Lättman H, Brand A, Hedlund J, Krikorev M, Olsson N, Robeck A,
Rönnmark F & Mattsson J-E. (2009) Generation time estimated to be
25–30 years in Cliostomum corrugatum (Ach.) Fr. The Lichenologist 41:
557–559.
Paper II
Lättman H, Lindblom L, Mattsson J-E, Milberg P, Skage M & Ekman S
(2009) Estimating the dispersal capacity of the rare lichen Cliostomum
corrugatum. Biological Conservation 142: 1870–1878.
Paper III
Lättman H, Milberg P, Palmer MW & Mattsson J-E (2009) Changes in
the distribution of epiphytic lichens in southern Sweden using a new
statistical method. Nordic Journal of Botany 27: 413–418.
Paper IV
Lättman H, Bergman K-O, Rapp M, Tälle M, Westerberg L & Milberg
P. Biodiversity in the wake of urban sprawl: loss among epiphytic
lichens on large oaks. Submitted manuscript.
Published papers are reproduced with kind permission from the publishers.
My contributions to the papers
I have, together with the co-authors, designed all field work for Paper I–II and IV. I
performed all field work by myself for Paper II, about half in III–IV and in
collaboration with the other authors for Paper I. I also made DNA extractions, PCR
amplification and sequencing for Paper II as well as editing and alignment. I made all
the statistical analyses in Paper I as well as parts of the analyses in Paper II and IV. I
have been writing most of Paper III–IV, and contributed to Paper I–II.
IV
Abstract
Lättman, H. 2012. Studies on spatial and temporal distributions of epiphytic lichens
Doctoral dissertation
Lichens are an important group of organisms in terms of environmental issues,
conservation biology and biodiversity, principally due to their sensitivity to changes in
their environment. Therefore it is important that we develop our understanding of the
factors that affect lichen distribution. In this thesis, both spatial and temporal
distributions of epiphytic lichens at different scales have been studied in southern
Sweden.
Generation time of the red-listed lichen Cliostomum corrugatum was examined
using Bjärka-Säby as the study site. The results showed that the average age of an
individual of C. corrugatum is 25–30 years at the onset of spore production.
The rarity of C. corrugatum was also examined. DNA analysis of an intron from 85
samples, collected at five sites in Östergötland, yielded 11 haplotypes. Results from
coalescent analysis, mantel test and AMOVA indicated that C. corrugatum have a high
ability to disperse. The study concluded that its rarity is most likely connected with the
low amount of available habitat, old Quercus robur.
The changes in the distribution of epiphytic lichens in southern Sweden, between
1986 and 2003, were also compared. For each year a centroid was calculated on all
combinations of tree and lichen species. The three significant cases showed that the
centroid movement pointed toward a north-east or north-north-east direction.
Finally differences in species richness and cover of lichens on large Q. robur were
examined between urban and rural environment. The results demonstrated that species
number and percent cover was significantly higher on oaks standing rural compared to
oaks standing urban. Effects of urban sprawl showed a decline in species richness and
cover with increasing age of the surrounding buildings.
Keywords: centroid, Cliostomum corrugatum, direction, dispersal, generation time,
global change, habitat availability, lichen, movement, Quercus robur, range shift, urban
Authors address: Håkan Lättman, School of Life Sciences, Södertörn University,
SE-141 89 HUDDINGE, Sweden; IFM Division of Biology, Linköping University,
SE-581 83 LINKÖPING, Sweden.
E-mail: [email protected]; [email protected]
ISSN 1652–7399
ISBN 978–91–7519–810–1
ISSN 0345–7524
V
Introduction
Lichens are amazing, fascinating and slightly peculiar organisms. They can be described
as small ecosystems in their own right, in which several groups of organisms live
together in the same body (Bates et al. 2011). This symbiosis is sensitive to changes in
the external environment and therefore is an important model in providing answers to
many of our questions concerning the environment. Throughout the history of the Earth
the environment has constantly been changing in response to various causes.
Undoubtedly, today humans have had the greatest impact on the environment (Vitousek
et al. 1997, Foley et al. 2005) and species diversity (Jenkins 2003). The development of
human civilisation has resulted in a widespread exploitation of nature with significant
degradation effects, including recent global climatic changes (Vitousek et al. 1997).
Due to mankind’s large and unprecedented impact on our surrounding, it has been
suggested that the current geological epoch Holocene has come to an end and that we
are now entering Antropocene (Zalasiewicz et al. 2008). Due to our actions, more and
more of the Earth’s surface is exploited, resulting in an increasing habitat loss and
fragmentation of the remaining habitats. This has led to the decline in abundance and
distribution of many species, and also their extinction in several cases (Gonzalez et al.
1998).
In order to understand and predict how species will respond to human activities in
natural communities, basic knowledge about species behaviour is vital. It is also
important to study their different requirements i.e. sunlight, chemical composition of the
atmosphere, temperature, humidity and the choice of substrate in order to conserve
biodiversity. How environmental changes will affect individual species are difficult to
predict. Many species’ environmental requirements are not fully understood and
therefore it is important to be able to draw general conclusions. The environmental
impact on a species can in turn affect other species to form a chain reaction where more
and more species will be affected either positively or negatively. In Europe especially,
the broad-leaved forests have been affected by human disturbance (Hannah et al. 1995).
Many lichens, insects, and fungi are dependent on these forests and are unable to extend
their range to other habitats. Globally, lichens are a group of organisms that have been
less studied than other comparable multi-cellular organisms. Thus, there is a gap in the
scientific knowledge concerning lichen species’ dispersal capacity and establishment on
different substrates, their habitat requirements, and population structure. Our lack of
knowledge of lichens is probably explained by their inconspicuousness and their small
thalli, which may make them difficult to identify. This might also explain why Carl von
Linné (1707–1778) effectively ignored lichens. Fortunately, his protégé, Erik Acharius
(1757–1819), made great progress by identifying many lichen species, estimating them
to comprise of more than 300 taxa (Krempelhuber 1867). In Sweden today there are
more than 2400 known taxa (Feuerer 2009). Hale (1974) reported the worldwide
number of lichen species to be approximately 17000. Ten years later Hawksworth and
1
Hill (1984) reported the number to be 13500. At present there are 18803 described
lichen species (Feuerer 2009). The reported number of lichens occurring on Earth is
probably underestimated. Swedish lichens and flora are well-studied in comparison with
other countries, and contains a large proportion of the worlds lichens (Table 1). Shown
in Table 1 is the total number of species of some groups of organisms in Sweden
(Gärdenfors 2010) and worldwide (Chapman 2009) and the proportion in Sweden. It is
almost certainly an exaggeration that 13% of the earth lichen species exist in Sweden,
and is an artefact of this extensive local analysis. It is worth noticing the large
proportion of moss and mushroom species that are also present in Sweden which are
also probably due to a large number of undescribed species worldwide.
Table 1. Total number of species in Sweden and worldwide for ten major groups of
organisms.
Groups of organisms
Sweden
Worldwide
Proportion of species (%)
Lichens
2419
18803¤
12.86
Mosses
1049
16236
6.46
Mushrooms
~5000
98998
5.05
Birds
253
9990
2.53
Insects
23900
~1000000
2.39
Arachnids
1821
102248
1.78
Mammals
63
5487
1.15
Vascular plants
1556
281621
0.55
Fishes
142
31153
0.45
Amphibians & reptilians
19
15249
0.12
¤
Number taken from Feuerer (2009).
What is a lichen?
Lichen symbiosis always consists of a mycobiont and photobiont. The mycobiont is a
fungus, mostly an ascomycete, but in some lichens it is a basidiomycete. A photobiont
that is capable of photosynthesis is an algae or a cyanobacterium. In some lichens, both
an algae and a cyanobacterium are present together with the fungus. There is one lichen
described in which the algae belong to phaeophyceae (Sanders et al. 2004). The
different organisms belong to different kingdoms and include two domains. The nature
of the symbiotic relationship is not trivial. It appears that the fungus sometime acts like
a parasite on the photobiont (Brodo et al. 2001). In other lichens, however, the
relationship should be considered mutualistic. Both the mycobiont and photobiont can
be obligately or facultatively associated with the symbiosis. When the association is
obligate, the mycobiont and photobiont can only occur in the lichenized stage which is
the stage that describes the symbiosis while in the case of facultative association the
organisms may either be free living or a part of the symbiosis (Nash III 1996).
Dispersal strategies
The fungal partner in lichens reproduces sexually with spores like other fungi. After
fertilization between two different ascomycete mating types, ascospore production is
established. When talking of sexual reproduction of lichens, it is only the fungal
ascospores that act and function as dispersal units, this is referred to as mycobiont
dispersal. Lichens furthermore have the ability to disperse asexually and in those cases
fungi and photobiont disperse together in a process called vegetative dispersal. Soredia
2
and isidia are two examples of asexually produced dispersal units. Soredia are
microscopic globule-shaped units that usually originate from the algae layer and consist
of singular algal cells surrounded by some fungal hyphae. Isidia are small outgrowths
on the cortex of the lichen that have a similar internal structure as the thallus. Moreover,
the thallus in some lichens gets easily fragmented and small pieces of the thallus may
disperse from one location to another. Dispersal is when sexually and asexually
produced units move from one place to another often away from their place of origin.
Successful dispersal is when vegetative dispersal units or spores from the mycobiont
succeed in finding a suitable algae at the new site and establish a relichenization. Thus,
dispersal is never successful unless spores or vegetative diaspores (e.g. isidia, fragments
of the thallus or soredia) spread and establish functional thalli on uncolonized patches.
The spores and vegetative diaspores of lichens disperse from one site to another by
means of both biotic and abiotic vectors. Ants and oribatide mite are examples of biotic
vectors (Bailey 1970, Stubbs 1995), while the wind is an abiotic vector (Hansson et al.
1992). The total numbers of dispersed ascospores increase with decreased ascospore
size, i.e., small spores often disperse over a greater distance than large spores. Usually,
most lichen ascospores are small, approximately 1–30 µm. The ascospores of lichens
are often assumed to have unlimited dispersal over great distances (Hansson et al.
1992). Large ascospores and most vegetative diaspores are supposed to disperse over a
shorter distance. Thus, they mainly contribute to population turnover at a site rather than
to dispersal of the species over a longer distance (Hansson et al. 1992). Long distance
dispersal (LDD) of ascospores and eventually vegetative diaspores of lichens is
necessary in order to expand a species distribution; Nathan (2006) discussed LDD for
plants and claimed the importance of extreme weather events to ensure LDD; this may
also apply to lichens.
Dispersal of lichens is the activity when the spores or vegetative diaspores of an
individual move from one place to another. To establish offspring in new habitats, fast
dispersal over long distances may be promoted by easily spread diaspores. This
dispersal may be passive or active. The passive dispersal refers to a situation when a
vector, e.g., wind, water or animals carries the diaspore. Active dispersal does not occur
in lichens but it describes the process when the individual itself actively moves to
another site. Conditions such as a high density and high competition are known to
influence individuals of certain species to engage in acts of active dispersal (Johst &
Brandl 1997, Bowler & Benton 2005). The success of the dispersal is related to the
abundance of acceptable habitats and substrates.
The results of studies examining the distance that different lichen species are able
to disperse differ. Tapper (1976), Armstrong (1987, 1990), Heinken (1999) and
Lorentsson and Mattsson (1999) have shown that dispersal can range up to a few
hundred meters. Dispersal limitation has also been reported for lichens on trees, e.g.,
within tree stands, between compact tree stands, or between trees up to a few kilometres
apart (Dettki et al. 2000, Sillett et al. 2000, Hilmo & Såstad 2001, Johansson & Ehrlén
2003, Walser 2004, Öckinger et al. 2005). However, genetic studies of Xanthoria
parietina and Lobaria pulmonaria suggest that they indeed have efficient dispersal
within a few kilometres radius and furthermore, there was no sign of any of them being
dispersal-restricted (Lindblom & Ekman 2006, 2007, Wagner et al. 2006, Werth et al.
2006a, 2006b). At a large spatial scale, for populations separated by hundreds of
kilometres or more, genetic studies of lichen populations have revealed considerable
gene flow (Printzen et al. 2003, Palice & Printzen 2004, Walser et al. 2005), whereas
3
studies relying on biogeographic patterns (Munoz et al. 2004), trapping of lichen
fragments in the atmosphere (Harmata & Olech 1991) and observations of lichen
fragments on bird feet (Coppins & James 1979) concluded that effective dispersal was
frequent. Finally, the small size and weight of the ascospores has been taken as indirect
evidence that lichens are able to disperse widely (Nordén & Appelqvist 2001).
Enduring harsh environments, yet also sensitive
Lichens can give us answers to many questions regarding changes in the environment
caused by humans since they are sensitive and respond more or less instantly. One
explanation for this sensitivity to human activities is the complex structure of the
symbiosis. However, somewhat paradoxically, this is also the key to their survival in
very harsh environments. The association between the mycobiont and the photobiont in
lichens has been a successful relationship lasting at least 4 × 108 years (Taylor 1995). In
partnerships, the two organisms are able to withstand harsh environments that they
could not withstand individually. Together they are tough and can survive inhospitable
environments where many other organisms have difficulties to survive. For example
lichens occur at high altitude and latitude, where other groups of organisms have
difficulties to survive. Furthermore, some species live inside stones. With their hyphae
they can penetrate even granite, with the only externally visible structure the fruiting
body (Brodo et al. 2001).
Another clear example of their ability to withstand harsh environments has been
demonstrated by the lichens Rhizocarpon geographicum and Xanthoria elegans. The
two species have survived in outer space (Sancho et al. 2007a). A prerequisite to cope
with these harsh environments is that the lichens have had the time to dehydrate. In this
dry state, called cryptobiosis, the metabolic rate drops dramatically and the lifesustaining activities in the cells are very low. Once they have entered this state, lichens,
as just mentioned, can survive in extreme environments. On the other hand, lichens are
sensitive to other changes in abiotic and biotic factors, for instance, polluted air in urban
areas. One reason for this is, compared with spermatophytes, the lack of a protective
cuticle. A cuticle is a waxy outer layer on, for instance, the leaves that prevents harmful
airborne compounds from entering and damaging the internal tissue.
Lichens do not have roots for their uptake of water like spermatophytes; instead
they satisfy their need of water from precipitation and moisture in the air. In
spermatophytes, the casparian strip in the roots is also a barrier that helps the plants
avoid harmful substances. Thus, it is not the difference in absorption capacity that
makes lichens sensitive compare to spermatophytes. It is when the precipitation and
surrounding air contains harmful and toxic substances that uptake of water becomes a
problem as they are unable to control what enters the thallus. The lack of a cuticle
allows the substances to enter the tissues and cells and injure or kill the lichen. Thus,
lichens can withstand extremely harsh environments, yet they are sensitive to certain
types of changes in the environment, especially those related to an increase in harmful
substances in the atmosphere and to changes in air humidity. Hence, in addition to the
general threats to biodiversity – habitat loss, fragmentation and global warming –
lichens are also (locally) threatened by air pollution (Giordani 2007, Affum 2011).
Sunlight
Lichens are dependent on sunlight, since they are photosynthetic organisms. The
amount of light or electromagnetic radiation needed for photosynthesis and how much
4
each lichen species can withstand varies among species (Demmig-Adams et al. 1990,
Green & Lange 1991, Gauslaa & Solhaug 1996, Kappen et al. 1998). The lichen
Verrucaria rubrocincata, which is endolithic and lives in desert environments, has been
shown to tolerate a high degree of electromagnetic radiation; even at 2600 μmol/m2 s–1
during a hot summer day the lichen still continues to photosynthesise (Garvie et al.
2008). For the epiphytic forest lichen Lobaria pulmonaria, Gauslaa and Solhaug (2000)
showed that the electromagnetic radiation never exceeded 610 μmol/m2 s–1 on the
Quercus trunk where the lichen occurred. These authors also transplanted L. pulmonaria
to a neighbouring tree trunk where the lichen did not occurred and appeared to have
similar light conditions. The results showed that the neighbouring Quercus trunk had 6
hours more radiation above 1000 μmol/m2 s–1 during early spring and peaked with 2000
μmol/m2 s–1. The transplanted specimen showed extensive bleaching and damage on the
chlorophyll. Forestry practices in Sweden have changed during recent centuries
including clear-cut. Lichens that have adapted to live in a forest, relatively shaded
compared with open habitat, may be permanently damaged due to high electromagnetic
radiation (Gauslaa & Solhaug 1999, Gauslaa et al. 2006). Out of all land areas in
Sweden (41.3 × 106 ha), 23 × 106 ha consists of production forest (NBF 2007). The
human impact on the production forest is significant and the forests are managed,
primarily to promote economic interests. Prevailing forest management is unfortunately
often incompatible with the necessary life conditions for lichens, dependent as they are
on old tree trunks and standing and prostate dead wood. Such habitats have become
scarce. Thus, the change in the demography of forests towards a greater proportion of
younger trees, and the practice of clear-cutting, are negative for many lichen species.
The air
Soon after the industrial revolution in Europe in the late 18th century, air quality started
to change. Independent observers in Manchester, England (Grindon 1859), Munich,
Germany (cf. Gries 1997) and Paris, France (Nylander 1866) documented in the mid
19th century that lichens were disappearing from the cities. Fifty years later, in the
beginning of the 20th century, the same pattern was recognised across the whole of
Europe (Erisman & Draaijers 1995, Mylona 1996). The cause for the decline of lichens
in cities was first suggested to be the dust from coal, but later it was realised that
sulphur dioxide (SO2) was the main toxic agent. This is one of the most lethal
substances for lichens and responsible for most of the modern decrease of lichens
(Gilbert 1968). Hawksworth and Rose (1970) showed that lichens could be used to
monitor the SO2 content in the air. They used an ordinal scale with teen categories of
lichens with increasing sensitivity to SO2.
International co-operation with the goal to reduce the effects of air pollutants on the
environment has been successful, and a dramatic decrease in SO2 has been observed in
recent decades (UNECE 1999). Schopp et al. (2003) predicted a continued decrease of
SO2 in Europe up until the end of 2030. Recolonization of lichens in London, United
Kingdom, has been shown by Hawksworth and McManus (1989) and has been
attributed to the decrease in sulphur dioxide. Another example of successful urban
recolonisation is the rare Parmelina tiliacea that was found in central Malmö, southern
Sweden, ten years ago (Kärnefelt & Lättman 2001). At that time, this species was redlisted in both Denmark and Sweden (critically endangered and vulnerable, respectively).
Most remarkable was that the specimen was healthy-looking and well-developed, and
apparently unaffected by the urban environment. This was in stark contrast to the other
5
lichens present on the tree that were not fully healthy-looking, and thereby of typical
appearance of lichens in a larger city. Another indication of the improved SO2conditions for this lichen is that subsequently, P. tiliacea has been removed from the red
list.
Climate change: Temperature and moisture
Anthropogenic greenhouse gases are the main cause of climate change that occurs today
(Rosenzweig et al. 2008). Climate is a description of weather for an extended period of
time that includes variables such as temperature, precipitation, humidity, atmospheric
pressure and wind. During the last decades, many studies of climate change have
revealed its enormous impact on numerous different organisms including lichens
(Vitousek 1994, Hughes 2000, Saxe et al. 2001, Walther et al. 2002, Parmesan & Yohe
2003, Root et al. 2003, Sanz-Elorza et al. 2003, Perry et al. 2005, Thuiller et al. 2005,
Parmesan 2006, Maclean et al. 2008, Allen et al. 2010, Dawson et al. 2011, Gosling et
al. 2011). It is difficult to separate and determine which of the various abiotic factors are
most important for species’ distribution and abundance. For example, Warren et al.
(2001) expected a positive response to climate warming in their study on butterflies
since they benefit from warm conditions. Unexpectedly three-quarters of the surveyed
species decreased. The authors concluded that the positive response to climate warming
had been overshadowed by the negative response to habitat loss.
Climate warming is probably the major contributor to changing the range
boundaries of terrestrial and freshwater habitats (Thomas 2010). Different groups of
organisms will have variable successes in meeting the challenges of necessary dispersal.
Malcolm et al. (2002) used several vegetation models to determine whether species are
able to move as fast as climate zones change. The results showed that changes of
climate zones will, in many cases, exceed species capability to migrate. In all vegetation
models, high migration rates ≥ 1000 m per year were relatively common. Species that
do not have the ability to move and establish on new sites fast enough will face a
rapidly changing environment and extinction of some of these species will be inevitable.
Thomas et al. (2004) modelled extinction risk for some species in 20% of Earths
terrestrial environments and estimated that by the year 2050, 15–37% of the species will
be committed to extinction.
Discussions about global warming and whether humans affect the process has been
and is still under debate. However, according to a survey among climatologists, 96%
believe that global average temperatures the past 100 years have increased and 97% that
it is induced by man (Doran & Zimmerman 2009).
Among the weather variables, it is temperature and water in different forms, e.g.,
precipitation, fog and dew that has most impact on lichens. For instance photosynthesis
is only activated when the lichen thallus is moist, and it stops when the lichen thallus is
dehydrated. In the dehydrated state, the lichen is highly resistant to extreme external
environment as discussed earlier, but this does not apply for the moist condition where
the thallus is sensitive. Lichens can withstand a wide range of temperature when
dehydrated, with temperatures ranging from 90°C (or even higher) to –196°C for
several days and –60°C for several years (Nash III 1996). However the exposure of a
moist thallus of Evernia prunastri to 80°C showed to have a strong negative effect
(Pisani et al. 2007). Furthermore some species of the genus Cladonia have proved to be
even more sensitive when moist. Kappen and Smith (1980) examined how close
Cladonia oceanica could grow a hot steam area of Hawaii. The results of maximum
6
temperature were 27.2°C in stunted form and 23°C in ramified growth form. Grüninger
(1988) sampled ten specimens of Hypogymnia physodes in Reutlingen, West Germany
and enveloped the thallus in paper. Two days later he transplanted the lichen thallus on
tree trunks on the campus of the University, San José, Costa Rica. Three thalli had died
after one week and after ten months they were all dead. The reason for this could be due
to the large differences in temperature between Germany and Costa Rica. When a
thallus is in a moist state and photosynthesis is activated lichens are most heat sensitive
(Rogers 1971, Kappen & Smith 1980). Photosynthesis requires water and the thallus
can absorb large amounts. The thallus of green algae lichens can absorb and maintain a
water content of 250–400% whilst in cyanolichens this figure can be as high as 600–
2000% and some even higher on a weight basis. When water is available, the thallus
reservoir is filled very quickly. In a few seconds 60–70% of the thallus becomes
saturated and full saturation is reached within minutes. Photosynthesis has its maximum
speed, for green algae lichens, when the thallus has a water content of 70–150% and
corresponding value for cyanolichens is 300–600% (Nash III 1996).
Substrate
Corticolous (growing on bark), lignicolous (growing on wood), saxicolous (growing on
rock) and terricolous (growing on soil) are convenient terms to describe the substrate
preferences of lichens. Some species have the ability to grow on several kinds of
surfaces while others are limited to just one or a few. Some epiphytic lichens are able to
grow on a variety of tree species, while others only grow on one or just a few.
Furthermore, some are demanding in terms of size or age of the tree and prefer a large,
old tree. Wedin et al. (2004) made a remarkable discovery regarding the genera Stictis
and Conotrema. They showed that a fungal species can adopt two different lifestyles
depending on the circumstances. If the fungal spore ends up on the substrate wood, the
fungus adopts a non-lichenized, saprophytic lifestyle and has been called Stictis. On the
other hand, if the fungal spore ends up on bark – and the needed photobiont is available
– the fungus adopts a lichenized lifestyle and has been called Conotrema. Hence, two
different genera proved to be the same fungal species but with different lifestyles:
lichenized or non-lichenized. The fungus’ plasticity in terms of lifestyle and the
frequency with which it occurs is unknown but may be common (Hawksworth 2005).
Thus, requirements on the substrate quality for different lichen species vary from
generalists to highly specialised lichen species.
As our human population increases, we use more and more land area: we claim
more space. As a consequence, the amount of available habitat reduces for most other
organisms. This unceasing demand to exploit new areas and their resources for our
benefit has created severe situations for many other non-human organisms, particularly
those that require large areas for their survival. Furthermore, remaining habitats become
fragmented and thus divided from each other into patches often surrounded by areas
affected by different kinds of intense human activity (e.g. agriculture, urban areas, roads
and railway tracks). Organisms with a limited ability to disperse face a severe situation
in fragmented areas. Fragmentation and its edge effects have been shown to have a
negative impact on lichen biodiversity (Turner 1996) and abundance (Esseen &
Renhorn 1998).
7
Aims of the thesis
The overall goal of this thesis is to increase our knowledge regarding the changes in
time and space in the occurrence of epiphytic lichen species and their communities with
the use of existing or new methods. The complexity of these factors and their
synergistic effects make it necessary to undertake studies on different spatial and
organisational levels. These range from genetic, individual and population levels
through to community and biotope levels. Thus, different studies with different
objectives and methods were designed in order to target the overall aims of the thesis.
Estimating the generation time of the red-listed crustose epiphytic lichen
Cliostomum corrugatum was the objective of Paper I. A method for assessing
generation times, from meiospore to meiospore, is often necessary in order to
understand population dynamics as well as to describe evolutionary history both in the
short as well as in a long run perspective. The results were used in the following study
(Paper II).
Paper II makes an attempt to determine whether the rarity of C. corrugatum is due
to difficulties with dispersal or if it is its habitat – old and often large Quercus robur –
that is limiting. Using the analysis of a nuclear ribosomal RNA gene, three different
methods to analyze the pattern, the lichens ability to disperse were tested.
Paper III examines the problems of describing changes in the lichen flora on a
regional scale. From field surveys of common epiphytic lichens in southern Sweden,
conducted in 1986 and 2003, the change in position of the centroids of these species
over time was assessed. A centroid of a species is the mean position of its sites in an
area, calculated from the coordinates of sampling sites.
Finally, in Paper IV the focus is to investigate differences in species richness and
cover of some common and rare epiphytic lichens on Q. robur standing in urban and
rural environments. The effect of urban sprawl was also examined on species richness
and cover on common and rare epiphytic lichens on Q. robur. Two methods were used
to measure the degree of urbanization, one of which took into account the average age
of five adjacent buildings and the second the area of nearby buildings at six radii
centred on a visited tree.
In the rest of this thesis, each of the papers is presented with a general background
introducing the specific study.
8
Further background and the included papers
Growth
Lichens have been studied for a long time and knowledge of their biology has
accumulated. However in the light of for example vascular plants we have just begun to
understand how lichens function, with this field much less studied. Crustose lichens
growth and growth rate has been studied and especially so in the lichen Rhizocarpon
geographicum (Armstrong 1983, Proctor 1983, Haworth 1986, Armstrong 2002, 2006,
Hansen 2010). In general, lichens are slow-growing and have a reputation for not only
this but also for being long-lived (Hale 1973, Matthews & Trenbirth 2011). However,
there are both fast- and slow growing lichens. Benedict (2008) reported an annual radial
growth rate for R. superficiale to be as little as 0.006 mm per year. An example of a fast
growing lichen is Usnea longissima, which showed a maximum growth of 18.4 cm in a
single year (Keon & Muir 2002). The variation in growth within a single species may
also be large. Hill (1981) reported annual radial growth rates (RGR) of Lecanora
muralis to be 0.03–0.55 mm per year while Seaward (1976) showed the RGR to be
2.84–6.05 mm per year. Several methods for study growth rate have been developed
(Platt & Amsler 1955, Farrar 1974, Honegger et al. 1996). Lichenometry is a frequently
used method where measurement of the lichen thallus RGR is central. The method is
used for example to date moraines and the retreat of glaciers (Karlén & Black 2002).
A lichen’s growth curve varies depending on growth form. The growth curve of the
crustose lichen R. geographicum can be divided into three parts: 1) where RGR
gradually increases to a maximum; 2) maximum speed is kept for a short time period;
and 3) the speed of RGR decreases (Armstrong 2005, Armstrong & Bradwell 2010).
There is no evidence that foliose lichens would have a phase where there is a reduction
of RGR as there is in crustose forms. The growth of lichens and their population
dynamics is, of course, also influenced by abiotic and biotic factors, but these affects
lichens differently depending on the species. The abiotic weather factors that mainly
affect lichens are temperature, humidity and sun exposure. Regardless of season there is
a growth all year round in R. geographicum but predominantly during the summer
months (Armstrong 2006). Other abiotic factors that affect lichens are nutrients.
Gauslaa et al. (2006) performed growth experiments with Lobaria pulmonaria. Some of
the thallus was sprayed with clean water and others with nutrients added to the water.
The results showed that water with added nutrients only slightly increased the growth of
L. pulmonaria. Results with stronger support for the importance of nutrients were
presented by McCune and Caldwell (2009), where L. pulmonaria thalli were immersed
in a bath of phosphorus for twenty minutes. After one year the biomass was doubled
compared to the control group. Thus, growth and growth rate is relatively well studied
at least for the lichens R. geographicum and L. pulmonaria.
9
Another population dynamic aspect which is also, at least in some areas, relatively
well studied is mortality. Most studies on mortality have dealt with pollution such as
emission from industries and urban areas to determine the highest concentrations of
harmful substances lichens can withstand before they disappear. The knowledge gap in
mortality is in knowing the dynamics under normal conditions. For instance, at what age
(or size) does a lichen thallus die a natural death or what is the lifespan of a lichen
thallus? There is some knowledge about the well studied lichen R. geographicum which
can reach an age of about 1000 year (Matthews & Trenbirth 2011) but the maximum
age is likely to be very much shorter in epiphytic lichens. In fact, Hypogymnia physodes
never reach such an age. Studies have shown that H. physodes has a growth rate of 3–4
mm per year (Gorbach & Kobzar 1981). Since the thallus rarely exceeds 5 cm in
diameter this means that the individuals of H. physodes reach approximately 6 to 8
years of age before they die. Furthermore Mattsson et al. (2006) have shown that H.
physodes has a rapid turnover in the sense of appearance/disappearance at sites.
By using the growth rate and lichen thallus diameter, it is possible to estimate the
age on an individual thallus. However there are many unanswered questions: what is the
population dynamics in terms of mortality, at what age do they die and begin to
reproduce sexually and asexually? Hence, to better understand the population dynamics
of lichens, we certainly need much more basic field research.
Generation time of lichens
Generation time can be explained as the time span from a given point in the parent life
cycle to the same given point in the offspring. Two different types of generation time
can be distinguished i.e. fundamental and realized generation time. The fundamental
generation time is based on the shortest possible time (age) of reproduction for an
individual of that particular species while realized time is the average parental age at
reproduction under natural conditions. Generation length is sometimes used
synonymously with the word generation time. Generation time varies for different
eukaryotic organisms, for instance, the oriental latrine fly, Chrysomya megacephala
have a short generation time in only 20.7 days (Gabre et al. 2005), in the same way
human generation time is approximately 25 years and Japanese timber bamboo,
Phyllostachys bambusoides, have a generation time on about 120 years (Janzen 1976).
The consequences of different generation time are that the genetic material will evolve
at different rates depending on the species.
The time span between meiosis events is important to estimate as these events have
a potential for genetic recombination, while the vegetative phase of organisms is more
inert at the genetic level. Thus, knowledge of species’ generation time is essential for
calculations of the speed of evolutionary changes.
The relative importance of sexual versus asexual reproduction depends on the
species. The spores in mycobiont-dispersed species have undergone genetic
recombinations that may increase genetic variation and the spores are often small and
may be dispersed far away. The downside is that the mycobiont must find a suitable
photobiont before being able to become lichenized. Species that disperse with
vegetative diaspores has the advantage that both partners are spread along together but
there is no recombination as they are clonal (Nash III TH & Gries 2002).
10
Paper I
In Paper I, the generation time of Cliostomum corrugatum was studied in Bjärka-Säby,
Östergötland, Sweden. The largest thallus area and largest diameter on apothecia were
recorded on Quercus robur. Only large trees were included in the study since the
occurrence of C. corrugatum is low on trees with a small circumference at breast height
(CBH) (e.g. Ranius et al. 2008, Johansson et al. 2009). By plotting thallus area or
apothecia diameter against oak diameter and extrapolate the regression line it was
possible to identify the age of Q. robur when C. corrugatum colonised it and at what
age it becomes fertile. Quercus robur CBH were translated into oak age and the
estimated time it takes for C. corrugatum to become fertile (fundamental generation
time) for C. corrugatum were found to be 25–30 years (Figure 1). The fertile thallus
may then continue to produce spores for many years ahead.
Figure 1. The epiphytic lichen Cliostomum corrugatum becomes fertile at an age of 25–
30 years.
It is surprising that sexual maturity takes such a long time to form in this lichen,
especially considering that by being an epiphyte, its substrate has a limited life span. To
my knowledge, this is the first time that anyone has determined the generation time of
lichens. It would be interesting to estimate this in other suitable lichen species. The
question of whether the generation time is longer among rare and red-listed species than
among common lichens appears to be particularly pertinent. Clearly more research is
required for a comprehensive picture of lichen generation times to emerge.
11
Is substrate or dispersal limiting?
Phorophyte and stand
Epiphytes are plants that grow on other plants, principally trees and shrubs but also
dwarf shrubs (e.g. Calluna and Vaccinium sp.) and even leaves (in the tropics) can be
used as a substrate. The common name for the various substrates is called phorophyte.
The utility of the bark substrate for various lichen species may vary considerably. The
common lichen Hypogymnia physodes does not have high demands on the substrate but
can grow on a wide range i.e. corticolous, lignicolous or sometimes saxicolous but also
man-made substances such as rubber and steel (personal field observations) (Figure 2).
Figure 2. The lichen Hypogymnia physodes is common in Sweden and grows on many
different substrates. It inhabits mostly bark, wood or stones but sometimes also manmade substrates such as rubber and steel.
What may be important for the epiphyte is to what species the phorophyte belongs, and
its age (or dimension of the trunk). Several phorophytes together form a stand that is
scattered in the landscape in different ways making it more or less suitable (such as sun
exposure, pH and structure of the bark), influencing accessibility for the establishment
and also how beneficial these stands are for long term survival of the lichen. In general,
areas designated for forestry have a lower value than protected stands since production
areas are more homogeneous and often lack old and large trees and also are low in dead
wood.
Spore dispersal
Present lichen distribution is in part a result of their ability to spread from one place to
another. Lichen ascospores are relatively small ranging in size from 2–3 μm for
Chaenotheca furfuracea to 150 μm for Phlyctis argena with a few other species having
even larger spores (Foucard 2001). The production of spores is large and one fruit body
may contain more than 1 × 106 spores, which can be equated with 12–18 × 106 spores in
one square centimetre (Tibell 1994). The main vector for their dispersal is wind which
probably is important for long distance dispersal (LDD). The air at ground level
contains a large amount of spores. Gregory (1978) measured the levels of spores, from
various cryptogams, during five summer months in Rothamsted, England, and showed
average concentrations of 12000 spores m-3 but for short periods as many as 1 × 106
may be present. The concentration decreases with increasing altitude with 10000 spores
m-3 one kilometre above the ground and hundreds of spores three kilometres up (Hirst et
al. 1967). In contrast to vascular plants, many lichens have very large geographic
distributions and if their distribution is not cosmopolitan, it may be pantemperate,
12
pantropical (Lücking 2003) or amphitropical (Søchting & Olech 1995, Myllys et al.
2003). The lichen Porpidia flavicunda has a circumpolar distribution and Buschbom
(2007) showed that the gene flow was high among the four surveyed sites. Furthermore,
the gene flow occurred in almost all possible directions and the lichen has had several
repeated LDD of vegetative diaspores between the sites. Högberg et al. (2002) made an
exciting discovery on Letharia vulpina. They found that in North America, this lichen
dispersed sexually by spores but in Europe the populations spread clonally by soredia
and/or isidioid soredia. Long distance dispersal with spores of species from the genus
Umbilicaria has on several independent occasions traveled to Antarctica from
surrounding temperate areas. This does not only apply to the spores but also to the algal
cells. Thus despite the hostile environment prevailing in Antarctica, successful
relichenizations have been established, on multiple occasions, between fungi and algal
cells (Romeike et al. 2002).
Several papers have argued that some lichen have limited dispersal abilities.
However, knowledge is often lacking about whether dispersal or substrate availability is
the limiting factor for a specific lichen population to survive in the long run, both
locally and regionally. To fully understand the importance of these different factors in
isolation or in combination, it is necessary to study both dispersal efficiency and the
impact of substrate abundance and microhabitat characteristics.
Paper II
In Paper II, I investigated whether the limited occurrence of the lichen C. corrugatum is
due to limitation by dispersal or limitation by habitat availability. The investigation was
conducted in Östergötland, south-eastern Sweden, at five sites. The laboratory methods
involved DNA extraction, PCR amplification and sequencing of a group 1 intron at the
end of the small subunit (SSU) nuclear ribosomal RNA gene. Attempts were made
using other genomic regions (ITS and IGS) but the variability at these regions were too
low for our purposes. Out of the 96 collected samples of C. corrugatum, 85 were
successfully extracted and shown to represent 11 haplotypes (Figure 3).
Figure 3. An unrooted haplotype network of the epiphytic lichen Cliostomum
corrugatum. The most common haplotypes 1 (N = 30) and 2 (N = 46) are in the centre
of the network and are most likely the oldest. The terminal haplotypes are rare (N = 1)
and have most likely derived from haplotypes 1 or 2.
13
Several statistical methods were used to analyse the genetic variation and to make
inference about the lichen’s ability to disperse. Firstly, a coalescent simulation showed
that the gene flow was considerable between the five investigated sites. Secondly, a
mantel test showed that there were no significant correlation between the genetic
distance and the geographic distance matrices. Thirdly, an AMOVA test showed that
0.4% of the variation was between the populations and 99.6% of the variation was
within the populations. All three tests indicate that C. corrugatum does not seem to have
any difficulties dispersing from one place to another. In addition our results indicate that
the five sites behave more or less as a single, sexual interbreeding population, i.e. a
panmictic population. Consequently, C. corrugatum rarity is likely to be connected with
the limited amounts of the suitable habitat, old oaks. The distribution of Q. robur is
more or less the whole of Europe, but unfortunately large, old oaks are relatively scarce.
During several hundred years an oak trunk serves as a suitable substrate for this lichen,
a time during which it can, with exception of the first 25 years, disperse spores (Paper
I). However, genetic methods have unfortunately its limitations. The result of the
genetic information reflects, to some extent, the population of C. corrugatum historical
status and not the lichens situation today.
Another context in which to view these results regards the peculiar Quaternary
history of the study area that involves land uplift. Since the end of the last glaciation, all
sites have gone from being submerged to terrestrial (Figure 4).
14
a)
b)
Figure 4. The two maps show a snapshot over the five studied sites during and after
the glacial retreat. a) Twelve thousand years ago the glacier edge (solid blue line)
stretched in a WSW/ENE direction and all five sites were submerged. b) Ten thousand
years ago the glacier had retreated and the three sites in the west have become
terrestrial.
15
At 10000 BP the three western sites had become terrestrial (Figure 4b). The remaining
two sites in the east (Stegeborg and Bråborg) had to wait, approximately, another 5000
years until they emerged from the sea. In southern Scandinavia Quercus invaded early
after the last glaciations and they have been present in the study area for at least 6000
years (Bradshaw 2000, Rasmussen 2005). It is important to note that a site becoming
terrestrial is just one prerequisite for oak colonisation. Another is the edafic conditions
and a third is whether oaks were present in the area at the time. Notably, the three sites
in the west, i.e. those with the longest history as terrestrial, also had greater haplotype
diversity (Figure 5). What this means for our understanding of C. corrugatum ability to
disperse and the evolution is difficult to evaluate. Have the rare haplotypes evolved at
these sites, or have they accumulated here through migration over time due to their
relatively long history? Should the lack of rare haplotypes in Bråborg and Stegeborg be
interpreted as an indication that the species is limited by dispersal?
6
Number of haplotypes
Solberga
y = 0.0271x + 2.2536
R2 = 0.8259
5
Bjärka-Säby
4
Orräng
Bråborg
3
2
Stegeborg
1
0
0
20
40
60
80
100
Height above see level (m)
Figure 5. The three sites in the west are higher above the sea level and have more
haplotypes than the two sites in the east.
Large-scale dynamic of lichens
Dynamics of lichen thalli
Lichens can be divided into crustose (resembling a crust), foliose (resembling a leaf)
and fruticose (resembling a shrub). In the group crustose lichens, many are slowgrowing (see above) and some have been used in lichenometry, which is a method for
dating exposed rock. I believe this is the major reason why the general picture of lichens
is that they grow slowly and have a slow population turnover. In this context, slow
population turnover means that the rate of mortality and nativity are low, i.e. individual
thallus remain attached to the same site year after year. In comparison with other sporeproducing organisms such as bryophytes, most crustose lichens have a slow turnover
16
(Pharo & Beattie 1997, Pharo et al. 1999), but this slow turnover does not apply to
foliose and fruticose lichens, at least not all of them. Gustafsson and Milberg (2008)
demonstrated that the foliose lichen Lobaria pulmonaria had a high turnover in southeastern Sweden (permanently marked thalii). Mattsson et al. (2006) showed that several
common species, e.g. the fruticose lichen Hypogymnia physodes in southern Sweden,
has a high turnover (presence/absence at sites). In California, USA, Boucher and Nash
III (1990) estimated the annual turnover of biomass of the common fruticose epiphytic
lichen Ramalina menziesii to be 29% where the annual turnover of biomass is the sum
of all fallen thallus. This implies that R. menziesii contribute substantially to the nutrient
turnover in the ecosystem. It also means that lichen thalli are constantly replaced by
new ones. Therefore, even though it might look like it is the same lichen thallus that sits
on the tree trunk every year, it may be a new one.
Lichens on the move
Being able to move from one place to another is generally fundamental for most
organisms, and this also holds true also for lichens. This has become especially
important in recent decades with rapid climate change. Since lichens have been shown
to be sensitive to changes in their environment, they are a useful group of organisms to
help detect the biological response to global warming (Pisani et al. 2007, Sancho et al.
2007b). The change in the amount of lichen biomass will ultimately affect the
distribution pattern of individual species.
In a climate that is changing it is expected that lichen species will alter their
distribution pattern. Long-term empirical studies in alpine and arctic environment have
shown conflicting results in terms of an increase or decrease in the amount of lichen
biomass. A decrease has been shown by Kari (2008) and Hudson and Henry (2010) that
contradicts findings by Hollister et al. (2005) who found an increase of lichen cover
across time. Unchanged amount of lichen biomass have also been reported (Hudson &
Henry 2009). Hauck (2009) proposes an alternative cause for the decline in lichens in
alpine and arctic environment. He argues that changes in the use of land and to high
atmospheric SO2 levels in the mid-20th century better correlated with the decline of
lichens than an annual increase in average temperature. Models incorporating different
types of scenarios have been used to predict future distribution and abundance of
different lichen species (Ellis et al. 2007a, Ellis et al. 2007b). To date, research has not
yet demonstrated that lichens have been seriously threatened because of climate change
but rather that a decrease or increase of species distribution will occur. Aptroot and van
Herk (2007) have shown that lichens with green algae Trentepohlia sp. as a photobiont
are increasing their distribution. There are areas though which could ultimately be
threatened by climate change. Aptroot (2009) highlights areas and habitats likely to
experience problems e.g. low-level islands with endemic lichens, arctic and tundra
regions and high ground in the tropics. It is important that we have methods to quantify
changes in the distribution of different lichens so we are able to take appropriate actions
and develop new and improve existing types of lands use, e.g. in forestry.
Paper III
This brings us to Paper III, where studies on lichens movements were investigated. In
this paper, movement of the, centre of distribution (centroid) within southern Sweden,
of some common epiphytic lichens were studied, based on a small repeated sampling.
The study was conducted at 64 sites and the inventory in the field was carried out in
17
1986 and 2003. Fifty-six epiphytic lichens and 22 tree species (phorophytes) were
included in the study. Thirty cases were possible to analyze, out of which three showed
a significant shift in the centroid. The centroid movements of the lichens Hypogymnia
physodes and Vulpicida pinastri on the tree species Juniperus communis were 50 km
and 151 km (p-value 0.0258, 0.0002) with the direction 27° and 48°, respectively. The
movement of the centroids of H. physodes on Pinus sylvestris was 41 km (p-value
0.0066) with the direction 30° (Figure 6). All three significant cases had moved in a
north-east or a north north-east direction.
Figure 6. The three arrows indicate direction and distance that the centroid had moved
between the years 1986 and 2003 of Hypogymnia physodes and Vulpicida pinastri on
the tree species Juniperus communis but also H. physodes on the tree species Pinus
sylvestris. Squares represent the study sites.
The data set was fairly small with only presence information of epiphytic macrolichens
on different substrates recorded. The sites were only roughly described, without
information on tree size and tree species abundance and trees without lichens were not
recorded. Hence, the statistical power of the analyses was low. Of the two species that
turned out significant at least H. physodes has a large ecological amplitude, which
18
should provide a strong resistance against small changes in environmental conditions.
Nevertheless, some significant changes were recorded on the movement of the centroid.
This indicates a greater impact of global warming on the epiphytic lichen flora than
previously presumed.
The temperature of the planet is increasing and has done so at least during the past
40 years (Rosenzweig et al. 2008). In our study area, the current trend appears to be the
same. On 29 meteorological stations in the study area the average temperature has
increased by 0.056°C year –1 during the period 1986 to 2003 (Figure 7) (SMHI 1986–
2003). Another explanation why some lichens have moved their centroid may be due to
large scale forest structure changes in recent years. Hedwall et al. (2012) used data from
the Swedish National Forest Inventory to compare the field- and tree layer in boreal and
temperate Sweden between 1994 and 2010. They found that the canopy has become
denser and the distribution of species abundance on the forest floor has changed as it
has become darker. The lichens could respond in a similar way to a darker environment.
9.0
The average temperature
8.0
7.0
6.0
5.0
y = 0.056x + 6.804
R² = 0.112
4.0
3.0
2.0
1.0
0.0
86 87 88 89 90 91 92 93 94 95 96 97 98 99 00 01 02 03
Year
Figure 7. The average temperature over 29 meteorological stations in southern
Sweden during the period 1986 to 2003.
Lichens in the urban environment
Urbanization transforms a natural, semi-natural or agricultural landscape into an
environment with buildings. In an ecological perspective, the organisms and the
environment they live in are, to a varying degree, interconnected and dependent on each
other. As an example, trees are of paramount importance for epiphytic lichens and many
tree species are, more or less, dependent on mycorrhizal fungi. There are many reasons
to maintain a high biodiversity in urban areas especially in highly urbanized areas. In a
study in Flanders in Belgium, an area where the proportion of forest is only 10%,
Cornelis and Hermy (2004) showed that their 15 surveyed city parks contained about
30%, 50%, 40% and 60% of all wild plant species, nesting birds, butterflies and
19
amphibians, respectively, of the total national number. This highlights the magnitude of
biodiversity even in cities. Urbanization is a major cause of a homogenization of biotic
factors (McKinney 2006), but the goal of biodiversity conservation should be towards
diversity.
In Sheffield, United Kingdom, Gaston et al. (2005) tested various methods to
increase biodiversity. They added nests for various insects, ponds for birds, dead wood
for fungi and patches of Urtica dioica for butterflies. Since urban areas are largely
composed of private and residential gardens, the investigation was focused on these
spaces. Some of the tested methods were found to boost biodiversity, indicating the
potential for retaining biodiversity within city borders.
Ranta and Viljanen (2011) list the causes of the relatively high biodiversity of
vascular plants in Finnish cities as spacious urban structure, small human populations,
late urbanisation, and abundant remnant natural vegetation (forest). The investigations
demonstrated that much can be done to increase biodiversity in urban environments and
there is certainly room for improvement. Hahs et al. (2009) examined the factors
affecting the rate of plant extinction from urban environment. They concluded that it
largely has to do with the city’s history in combination with current proportion of native
vegetation. They also claim that the transformation of the landscape to an urban area is
likely to involve an extinction debt, i.e. that there is a time delay in the loss of
biodiversity from our cities when they grow.
Trees: an important urban element
Trees in cities are in a stressful environment because usually they stand in a soil with
poor quality and their roots often have a limited ability to spread. In addition, roots are
frequently damaged during ground works i.e. when various types of pipes and cables
should be buried (Jim 1998). Increased runoff from buildings, hard surfaces and
drainage (Leopold 1968) which reduces the impact of a positive long-term effect of
moister from the precipitation in cities and the smog in some cities can cause tissue
damage on the trees (Middleton et al. 1958, Tripathi & Gautam 2007, Honour et al.
2009). Nevertheless, trees are often an important feature of cities. Hence, trees have
been planted, or retained from native or rural setting during urban sprawl, for a number
of reasons: wellbeing of people, economic value, aesthetics, the production of shade
and, in some situation, even for fire prevention purposes. Lohr et al. (2004) surveyed
residents in the United States concerning advantages and disadvantages associated with
trees in a city. Results from their study showed that the public evaluated trees ability to
shade and cool the surroundings highest and secondly it helped people feel calmer.
Highest ranked disadvantages were that they can cause allergies and block store signs.
Lohr et al. (2004) concluded that most people clearly appreciated the value of urban
trees in their lives.
Several studies have shown that trees in cities are highly economically valued. For
instance Donovan and Butry (2011) showed that an increased number of trees in a
residential garden increased the monthly rent by USD5.62 and trees in the public right
of way increasing the rent even more. Not surprisingly, the sale prices of buildings are
affected by trees. An increase in the number of trees also added an extra USD8870 to a
house in Portland, Oregon and reduced the time on the market by 1.7 days (Donovan &
Butry 2010). Nowak et al. (2002) calculated the value of urban forests using tree
valuation methods and field data. The evaluated total compensatory value was 101 ×
106 and 5.2 × 109 USD in Jersey City and New York, respectively. Donovan et al.
20
(2011) showed a reduced risk of poor birth outcomes by 1.42 per 1000 births if canopy
cover were increased by 10% within 50 m of residential buildings. Clearly is that those
who can afford, like to have trees around.
The studies above give us an indication that even in the future; trees will continue
to be highly valued in cities, although increasing land prices adds a threat towards
retained trees. On balance, the prospect for continued survival of epiphytic organisms in
cities seems brighter than for many other types of organisms.
Urban effects on lichens
Lichens have different tolerance or preferences concerning substrates. This means that
some lichens grow on different substrates, even some man-made. Hence, there are
lichens that grow on either coniferous or deciduous trees or others confined to a
particular tree species (Washburn & Culley 2006, Spier et al. 2010). Other lichens can
be restricted to a particular size of a tree e.g. large tree (Washburn & Culley 2006) that
usually has a coarse bark structure. Such specialized lichens have declined since the
supply of substrate is less in urban environments (Shukla & Upreti 2011).
Large trees of Quercus have proved to be very rich in species that also include
lichens (Rose 1974, Hultengren 1995), but few studies of lichens have been conducted
in an urban setting, an exception being Larsen et al. (2007) who investigated lichen
distribution on Quercus robur and Q. petraea in London, England in relation to air
pollution. The authors were able to distinguish three zones where lichen species
increase in number from the inner to the outer zones.
Lichens have been used to estimate air quality in urban areas. Wielgolaski (1975)
inventoried several groups of organisms such as vascular plants, bryophytes, lichens,
microorganisms, invertebrates, fish and plankton on their biological value to use as a
tool for evaluating air, freshwater and marine quality. The author argued that most
foliose and fruticose lichen in a wider sense can be used for evaluating the air quality.
Hawksworth and Rose (1970) well-known work was more fine-tuned and they also used
crustose lichens and algae. They were positioning a selection of lichens (and algae), on
a ten-point scale reflecting various sensitivities, and then used it to estimate the air
quality with particular attention to SO2. During the 1970s a number of studies showed
that lichens are adversely affected by road traffic emissions of SO2 (e.g. Brawn &
Ogden III 1977). Even the air quality and quantity of heavy metals has been
investigated by means of lichens (Pandey et al. 2002, Montero Alvarez et al. 2006).
Even though there are positive signs that the air quality is beginning to improve in urban
environments (Lisowska 2011), SO2 is still, 30 years later, considered the main limiting
factor for lichens in urban settings (Giordani 2007). Trees surrounding an expanding
city will be enclosed with buildings. Lichens on these trees will be isolated and
activities such as dispersal and establishment may be negatively affected. Mobile
organisms like animals can escape while trees will inevitably be incorporated and
trapped into the city or cut down during urban sprawl. The fate of epiphytic lichens will
parallel that of the trees.
Paper IV
In Paper IV, I examined species number and cover of epiphytic lichens on remnant oaks
in urban and rural environment. In Linköping, County of Östergötland, Sweden, 105
urban and 109 rural oak trees were surveyed for 17 selected lichen species. Trees with a
CBH > 250 cm were selected from a database developed by the County Administrative
21
Board in Östergötland. The majority of the available urban trees were selected since
there were only a limited number. I then selected a population of rural trees, aiming for
(i) a population of similar circumferences and (ii) similar densities of oaks in the
surrounding. I calculated densities with a radius of 302 m, as a previous study had
identified this as the appropriate scale for lichen richness in the study area (Muhammadi
2011). Nine of the lichens were common and eight were rare. During the field-work,
CBH, depth of bark crevices and sun exposure were documented per tree. The density
of oaks in the vicinity of target trees was calculated within radii of 150, 250, 350, 500,
700 and 1000 m. Two variables were also constructed as a measure of degree and age of
urbanisation. Firstly, the density of buildings around each Q. robur within a radius of
150, 250, 350, 500, 700 and 1000 m was calculated. Secondly, the average age of the
five closest buildings were also used as a measure of the age since urbanization.
Lichens richness and cover was higher in the rural environment than in the urban
environment there were, however, one exception: Lepraria incana (Figure 8).
22
Figure 8. a) to c) Fourteen investigated lichen species and their occurrence and cover
between urban and rural environment on tree trunks of Quercus robur are compared.
a) Shows the proportion of trees with lichen occurrence. b) Shows the average
percentage cover on trees with occurrence. c) Shows the average percentage cover of
all trees (also those lacking the species in question).
23
Our results recorded a clear effect of urban sprawl. Both species richness and cover of
lichens were significantly higher in younger compared to older parts. When including
all the external factors, it was possible to analyse eight out of the 14 lichens. Input
factors include: age of urbanization, tree circumference (held constant), bark crevices,
sun exposure, density of oaks (held constant) and the cover of buildings. Several species
were negatively affected by the age of the urbanization process but one was positively
affected. Age of urbanization affected five species negatively while cover of buildings
affected three species negatively, for the radii 500, 700 and 1000 m and one positive for
radius 350 m. The tree circumference and the bark crevices were positively affected
except for one species where the whole model showed to be insignificant.
Cities are artificial environments but, despite this, they can be compared to natural
ones. If we think of the buildings as small hills and the tiles and asphalt on the ground
as rocks, there are certain similarities to natural habitats. But what makes them
different? In the city there are several factors that make them different from rural areas,
i.e. pollution (more harmful particles in the air), temperature (higher), moisture,
disturbance and habitat configuration (McDonnell et al., 1997), but also habitat loss and
fragmentation (Niemelä, 1999, McKinney, 2006). Thus, the reason for the decrease in
species number and amount of lichens in the urban area is probably a mix of several
factors, for example the lichens dispersal mode, a higher temperature and lower
humidity.
24
Concluding remarks
All studies in this thesis have dealt with epiphytic lichens and their movement through
space and time. Noteworthy findings that contribute to the lichenology are, in my view,
the following. First, an epiphytic lichen must be relatively old before they start
producing spores, 25–30 years for the red-listed species C. corrugatum. Whether
generation time is a distinguishing feature for rare lichens and how it relates to common
epiphytes is an open question. Second, this rare species does not seem to be limited by
dispersal, at least as judged by the examined gene haplotype distribution. This finding
contributes to the ongoing debate of dispersal vs. habitat availability as the cause for
rarity, and to our view of epiphytic lichens as more mobile then generally thought.
Third, I was able to detect shifts in the distribution of some epiphytic lichens within just
a few decades using crude field data. This point to the benefits of using lichens for
monitoring large-scale changes in biodiversity. Fourth, trees that are integrated in a
growing city do not provide the same epiphytic lichen abundance or diversity as their
rural counterparts. Furthermore, age of urbanization affected species occurrences
negatively, which means that we can expect further extinctions during urban sprawl if
habitats for epiphytic lichens are surrounded by buildings.
Overall, I hope this thesis will contribute to our understanding of epiphytic lichens
as highly mobile organisms and their great value for monitoring changes, as well as
fascinating and unpredictable in their own right.
25
Populärvetenskaplig sammanfattning
Vad är en lav?
En lav består av en svamp och en alg som lever i symbios vilket betyder ungefär; att
leva tillsammans. Ibland finns även en cyanobakterie i symbiosen. Svampen kallas för
mykobiont och är för det mesta en sporsäckssvamp, men hos vissa lavar är det en
basidsvamp. Algen är för det mesta en grönalg eller cyanobakterie och kallas för
fotobiont. Dessutom finns ett fall beskrivet där algen är en brunalg. I Sverige finns det
drygt 2100 lavarter som är vetenskapligt beskrivna och i hela världen nästan 19000.
Lavar som växer på träd kallas för epifyter. Det kommer från grekiskan där epi betyder
”på” och phyton (fyt) betyder ”växt”. Lavar är en viktig organismgrupp att studera
eftersom de kan ge oss svar på många frågor när det gäller förändringar i vår miljö. De
är nämligen känsliga för yttre förändringar samtidigt som de är tuffa och kan leva i
ogästvänliga miljöer där många andra organismer har svårt att överleva.
Kartlav har visat på deras förmåga att överleva i extrema miljöer. Under 14 dagar
vistades laven i yttre rymden. En förutsättning för att de ska klara av de här extrema
miljöerna är att vävnaderna har haft en chans att torka upp och få ett lågt fuktinnehåll. Å
andra sidan så är de känsliga för t.ex. luftens föroreningar i urbana miljöer. Orsaken är
att de saknar en skyddande kutikula. Växter däremot har en kutikula, det består av ett
vaxartat yttre lager som bl.a. täcker bladen och hindrar skadliga luftburna föreningar
från att tränga in och skada cellerna.
Lavar har inte några rötter utan tar upp vatten från det regn som faller och från den
fukt som luften innehåller. Vad som är problematiskt för laven är när nederbörden och
luften innehåller skadliga och giftiga ämnen. En avsaknad av en kutikula gör att de
skadliga och giftiga ämnena tränger rakt in i vävnader och celler och skadar och dödar.
Så de kan alltså uthärda hårda miljöer men samtidigt är de känsliga för föroreningar och
förändringar i miljön. I den här doktorsavhandlingen har jag studerat utbredning och
spridning av epifytiska lavar i tid och rum i södra Sverige.
Resultat från min forskning
I Bjärka-Säby, Östergötland, sökte jag på flera ekstammar efter den sällsynta laven gul
dropplav för att mäta diametern på det apothecium som var det största men även mäta
och registrera arean på den största bålen (Papper I). Apothecier kallas lavens
fruktkroppar, de bildas vid sexuell reproduktion. Gul dropplav är epifyt på framförallt
stora gamla ekar men kan även förekomma på alm och asp. Den är trefärgad; ljusgrå
bål, ljusgula apothecier och svarta pyknid och relativt lätt att känna igen i fält (pyknid är
asexuellt bildade fruktkroppar som många gånger har formen av en liten urna).
Undersökningens syfte var att ta reda på hur många år det tar för en spor av gul
dropplav att växa upp och själv bilda egna sporer. En spor är lavar, mossor och
26
svampars motsvarighet till de blommande växternas frön. Resultaten visade att gul
dropplav bildar sporer först i 25–30 årsåldern.
Med hjälp av genetiska metoder försökte jag ta reda på varför gul dropplav är
sällsynt (Papper II). Jag hade som hypotes att det antingen kunde bero på att den har
svårt för att sprida sig från en plats till en annan. Eller så kunde det bero på att stora
gamla ekar är sällsynta. Är stora gamla ekar sällsynta så kommer också laven att bli
sällsynt. Fem lokaler i Östergötland med ekar inventerades på gul dropplav och
undersöktes genetiskt. Samtliga tester antydde att gul dropplav inte har svårt för att
sprida sig från en plats till en annan, utan det verkade vara den låga tillgången på stora
och gamla ekar som gör gul dropplav sällsynt. Genetiska metoder har tyvärr sina
begränsningar. Resultaten från den genetiska informationen avspeglar, till viss grad, gul
dropplavs status bakåt i tiden och inte lavens situation idag.
I en annan studie undersökte jag skillnader i förekomsten hos några vanliga
epifytiska lavar åren 1986 och 2003 hos (Papper III). Studien genomförds i södra
Sverige på 64 lokaler. För vart och ett av åren beräknades en medelpunkt, en så kallad
centroid, för varje kombination av lokal, träd- och lavart. Utav alla tänkbara
kombinationer var det 30 fall som kunde analyseras, varav 3 uppvisade en signifikant
förflyttning av centroiden, samtliga i en nordöstlig eller nord-nordöstlig riktning. Vår
tolkning är att detta speglar den globala uppvärmningen. De två lavarna har en nordlig
utbredning så riktningen av centroid-förflyttningen mot norr på grund av ett varmare
klimat är inte osannolikt.
I en fjärde studie (Papper IV) studerade jag epifytiska lavars antal och
täckningsgrad på ek där hälften stod i stadsmiljö i Linköping och andra hälften i
närliggande landsbygd. Jag valde träd så att de två grupperna skulle få så lika omkrets
som möjligt (> 250 cm) och så lika täthet av ekar i omgivningen som möjligt. Det fanns
betydligt fler lavarter och större täckning av lavar på landsbygden än i staden. Samtliga
ekar är stora, och fanns på platsen innan urbaniseringen började. Jag beräknade
tidpunkten för urbaniseringen som medelvärdet på byggnadsåret för de fem närmaste
byggnaderna runt varje ek. Även här var resultaten tydliga: antal lavarter och täckning
av lavarna minskar med tiden en ek stått omgiven av byggnader. Anledningen till att
artantalet och mängden lavar är lägre i staden beror troligtvis på flera faktorer och säkert
är de samverkande. I staden är det högre halter föroreningar i luften, en högre
temperatur, ändrade fuktighetsförhållanden i luften, förlust av habitat och
fragmentering.
Samtliga studier har behandlat lavar och deras förflyttning i tid och rum.
Undersökningarna har visat att lavar måste vara relativt gamla innan de själva bildar
sporer. De har visat sig bra på att sprida sig från en plats till en annan. Att de troligtvis
svarar på förändringar i klimatet. Och, trots förbättrade luftkvalitet, är de mindre
förkommande i urbana miljöer än på landsbygden.
27
Acknowledgement
I am indebted to so many people and I am sure that I am neither the first nor the last
with this debt of gratitude. There are so many people I would like to thank who meant a
lot to me during my time as a PhD. In order not to exclude anyone by mistake I:
collectively want to thank friends and colleagues who have contributed and
supported me to a doctoral thesis.
However, I cannot help but mention a few who has been particularly important to
me.
Great thanks to my supervisors Jan-Eric Mattsson and Per Milberg. Thanks to
unofficial stand in co-supervisors and co-authors Patrik Dinnétz and Karl-Olof
Bergman.
Thanks to my loving mother Ingegerd and my sister Eva for all your support and
encouragement. Last but not least, I want to thank my family Åsa, Elin, Gustav, Linn,
Kim and David for your presence and support.
28
References
Affum HA, Oduro-Afriyie K, Nartey VK, Addo MA, Nyarko BJB, Mumuni II, Adu PS,
Appiah GK & Coleman A (2007) Lichens as a monitor for atmospheric manganese
pollution. Research Journal of Environmental and Earth Sciences 3: 534–540.
Allen CD, Macalady AK, Chenchouni H, Bachelet D, McDowell N, Vennetier M,
Kitzberger T, Rigling A, Breshears DD, Hogg (Ted) EH, Gonzalez P, Fensham R,
Zhang Z, Castro J, Demidova N, Lim J-H, Allard G, Running SW, Semerci A &
Cobb N (2010) A global overview of drought and heat-induced tree mortality
reveals emerging climate change risks for forests. Forest Ecology and Management
259: 660–684.
Aptroot A (2009) Lichens as an indicator of climate and global change. In: Letcher TM
(Ed.) Climate change: Observed impacts on planet Earth. Elsevier, Oxford.
Aptroot A & van Herk CM (2007) Further evidence of the effects of global warming on
lichens, particularly those with Trentepohlia phycobionts. Environmental Pollution
146: 293–298.
Armstrong RA (1983) Growth curve of the lichen Rhizocarpon geographicum. New
Phytologist 94: 619–622.
Armstrong RA (1987) Dispersal in a population of the lichen Hypogymnia physodes.
Environmental and Experimental Botany 27: 357–363.
Armstrong RA (1990) Dispersal, establishment and survival of soredia and fragments of
the lichen, Hypogymnia physodes (L) Nyl. New Phytologist 114: 239–245.
Armstrong RA (2002) The effect of rock surface aspect on growth, size structure and
competition in the lichen Rhizocarpon geographicum. Environmental and
Experimental Botany 48: 187–194.
Armstrong RA (2005) Growth curves of four crustose lichens. Symbiosis 38: 45–57.
Armstrong RA (2006) Seasonal growth of the crustose lichen Rhizocarpon
geographicum (L.) DC. in south Gwynedd, Wales. Symbiosis 41: 97–102.
Armstrong RA & Bradwell T (2010) Growth of crustose lichens: A review. Geografiska
Annaler 92 A: 3–17.
Bailey RH (1970) Field and study notes: Animals and the dispersal of soredia from
Lecanora conizaeoides Nyl. ex Cromb. The Lichenologist 4: 256.
Bates ST, Cropsey GWG, Gregory Caporaso J, Knight R & Fierer N (2011) Bacterial
communities associated with the lichen symbiosis. Applied and Environmental
Microbiology 77: 1309–1314.
Benedict JB (2008) Experiments on lichen growth III. The shape of the age-size curve.
Arctic, Antarctic and Alpine Research 40: 15–26.
Boucher VL & Nash III TH (1990) The role of the fruticose lichen Ramalina menziesii
in the annual turnover of biomass and macronutrients in a blue oak woodland.
Botanical Gazette 151: 114–118.
29
Bowler DE & Benton TG (2005) Causes and consequences of animal dispersal
strategies: relating individual behaviour to spatial dynamics. Biological Reviews
80: 205–225.
Bradshaw RHW, Holmqvist BH, Cowling SA & Sykes MT (2000) The effects of
climate change on the distribution and management of Picea abies in southern
Scandinavia. Canadian Journal of Forest Research 30: 1992–1998.
Brawn K & Ogden III JG (1977) Lichen diversity and abundance as affected by traffic
volume in an urban environment. Urban Ecology 2: 235–244.
Brodo IM, Sharnoff SD & Sharnoff S (2001) Lichens of North America. Yale
University Press, New Haven.
Buschbom J (2007) Migration between continents: Geographical structure and longdistance gene flow in Porpidia flavicunda (lichen-forming Ascomycota). Molecular
Ecology 16: 1835–1846.
Chapman AD (2009) Numbers of living species in Australia and the world. Report for
the Australian Biological Resources Study, Australian Biological Resources Study,
Canberra.
Coppins BJ & James PW (1979) Birds and the dispersal of lichen propagules. The
Lichenologist 11: 105–106.
Cornelis J & Hermy M (2004) Biodiversity relationships in urban and suburban parks in
Flanders. Landscape and Urban Planning 69: 385–401.
Dawson TP, Jackson ST, House JI, Prentice IC & Mace GM (2011) Beyond
predictions: Biodiversity conservation in a changing climate. Science 332: 53–58.
Dettki H, Klintberg P & Esseen PA (2000) Are epiphytic lichens in young forests
limited by local dispersal? Ecoscience 7: 317–325.
Donovan GH & Butry DT (2010) Trees in the city: Valuing street trees in Portland,
Oregon. Landscape and Urban Planning 94: 77–83.
Donovan GH & Butry DT (2011) The effect of urban trees on the rental price of singlefamily homes in Portland, Oregon. Urban Forestry & Urban Greening 10: 163–168.
Donovan GH, Michael YL, Butry DT, Sullivan AD & Chase JM (2011) Urban trees and
the risk of poor birth outcomes. Health & Place 17: 390–393.
Doran PT & Zimmerman MK (2009) Examining the scientific consensus on climate
change. EOS, Transactions of the American Geophysical Union 90: 22–23.
Ellis CJ, Coppins BJ & Dawson TP (2007a) Predicted response of the lichen epiphyte
Lecanora populicola to climate change scenarios in a clean-air region of Northern
Britain. Biological Conservation 135: 396–404.
Ellis CJ, Coppins BJ, Dawson TP & Seaward MRD (2007b) Response of British lichens
to climate change scenarios: Trends and uncertainties in the projected impact for
contrasting biogeographic groups. Biological Conservation 140: 217–235.
Erisman JW & Draaijers GPJ (1995) Atmospheric deposition in relation to acidification
and eutrophication, Amsterdam, Oxford, Elsevier.
Esseen P-A & Renhorn K-E (1998) Edge effects on an epiphytic lichen in fragmented
forests. Conservation Biology 12: 1307–1317.
Farrar JF (1974) A method for investigating lichen growth rates and succession. The
Lichenologist 6: 151–155.
Feuerer T (ed.) (2009) Checklist of lichens and lichenicolous fungi, Version 1
September 2009. http:www.checklist.de
Foucard T (2001) Svenska skorplavar och svampar som växer på dem. Interpublishing,
Stockholm.
30
Foley JA, DeFries R, Asner GP, Barford C, Bonan G, Carpenter SR, Chapin FS, Coe
MT, Daily GC, Gibbs HK, Helkowski JH, Holloway T, Howard EA, Kucharik CJ,
Monfreda C, Patz JA, Prentice C, Ramankutty N & Snyder PK (2005) Global
consequences of land use. Science 309: 570–574.
Gabre RM, Adham FK & Chi H (2005) Life table of Chrysomya megacephala
(Fabricius) (Diptera: Calliphoridae). Acta Oecologica 27: 179–183.
Gaston KJ, Smith RM, Thompson K & Warren PH (2005) Urban domestic gardens (II):
experimental tests of methods for increasing biodiversity. Biodiversity and
Conservation 14: 395–413.
Gauslaa Y, Lie M, Solhaug KA & Ohlson M (2006) Growth and ecophysiological
acclimation of the foliose lichen Lobaria pulmonaria in forests with contrasting
light climates. Oecologia 147: 406–416.
Gilbert OL (1968) Bryophytes as indicators of air pollution in the Tyne valley. New
Phytologist 67: 15–30.
Gilbert OL (1990) The lichen flora of urban wasteland. The Lichenologist 22: 87–101.
Giordani P (2007) Is the diversity of epiphytic lichens a reliable indicator of air
pollution? A case study from Italy. Environmental Pollution 146: 317–323.
Gonzales A, Lawton JH, Gilbert FS, Blackburn TM & Evans-Freke I (1998)
Metapopulation dynamics, abundance, and distribution in a microecosystem.
Science 281: 2045–2047.
Gorbach NV & Kobzar NN (1981) Determination of the age of epiphyte lichen
Hypogymnia physodes (L.) Nyl. Doklady Akademii Nauk BSSR [Minsk] 25: 848–
851.
Gosling SN, Warren R, Arnell NW, Good P, Caesar J, Bernie D, Lowe JA, van der
Linden P, O’Hanley JR & Smith SM (2011) A review of recent developments in
climate change science. Part II: The global-scale impacts of climate change.
Progress in Physical Geography 35: 443–464.
Gregory PH (1978) Distribution of airborne pollen and spores and their long distance
transport. Pure and Applied Geophysics 116: 309–315.
Gries C (1997) Lichens as indicators of air pollution. In: Nash III TH (Ed.) Lichen
biology. University Press, Cambridge.
Grindon LH (1859) The Manchester flora. W. White, London.
Grüninger W (1988) Use of the temperate lichen Hypogymnia physodes (Parmeliaceae)
to evaluate air pollution in the Tropics. Revista de Biologia Tropical 36: 545–547.
Gustafsson A & Milberg P (2008) Changes in the abundance of Lobaria pulmonaria in
south-eastern Sweden from 1994 to 2007. Graphis Scripta 20: 44–51.
Gärdenfors U (2010) Rödlistade arter i Sverige 2010. The 2010 red list of Swedish
species. ArtDatabanken, Uppsala.
Hahs AK, McDonnell MJ, McCarthy MA, Vesk PA, Corlett RT, Norton BA, Clemants
SE, Duncan RP, Thompson K, Schwartz MW & Williams NSG (2009) A global
synthesis of plant extinction rates in urban areas. Ecology Letters 12: 1165–1173.
Hale ME (1973) Growth. In: Ahmadjian V & Hale ME (Eds.) The Lichens. Academic
Press, London.
Hale ME Jr (1974) The biology of lichens, 2nd Edn. A series of student texts in
contemporary biology. Edward Arnold, London.
Hannah L, Carr JL & Lankerani A (1995) Human disturbance and natural habitat: a
biome level analysis of a global data set. Biodiversity and Conservation 4: 128–
155.
31
Hansen ES (2010) A review of lichen growth and applied lichenometry in southwest
and southeast Greenland. Geografiska Annaler 92 A: 65–79.
Hansson L, Söderström L & Solbreck C (1992) The ecology of dispersal in relation to
conservation. In: Hansson L (Ed.) The ecological principles of nature conservation.
Elsevier Applied Science, London.
Harmata K & Olech M (1991) Transect for aerobiological studies from Antarctica to
Poland. Grana 30: 458–463.
Hauck M (2009) Global warming and alternative causes of decline in arctic-alpine and
boreal-montane lichens in North-Western Central Europe. Global Change Biology
15: 2653–2661.
Hawksworth DL & Rose F (1970) Qualitative scale for estimating sulphur air pollution
in England and Wales using epiphytic lichens. Nature 227: 145–148.
Hawksworth DL & Hill DJ (1984) The lichen-forming fungi. Blackie, Glasgow, United
Kingdom.
Hawksworth DL & McManus PM (1989) Lichen recolonization in London under
conditions of rapidly falling sulphur-dioxide levels, and the concept of zone
skipping. Botanical Journal of the Linnean Society 100: 99–109.
Hawksworth D (2005) To be or not to be a lichen. Nature 433: 468.
Haworth LA, Calkin PE & Ellis JM (1986) Direct measurement of lichen growth in the
central brooks range, Alaska, U.S.A., and its application to lichenometric dating.
Arctic and Alpine Research 18: 289–296.
Hedwall P-O, Brunet J, Nordin A & Bergh J (2012) Changes in the abundance of
keystone forest floor species in response to changes of forest structure. Journal of
Vegetation Science in press.
Heinken T (1999) Dispersal patterns of terricolous lichens by thallus fragments. The
Lichenologist 31: 603–612.
Hill D (1981) The growth of lichens with special reference to the modelling of circular
thalli. The Lichenologist 13: 265–287.
Hilmo O & Såstad SM (2001) Colonization of old-forest lichens in a young and an old
boreal Picea abies forest: An experimental approach. Biological Conservation 102:
251–259.
Hirst JM, Stedman OJ & Hogg WH (1967) Long-distance spore transport: Methods of
measurement, vertical spore profiles and the detection of immigrant spores. Journal
of General Microbiology 48: 329–355.
Hollister RD, Webber PJ & Tweedie CE (2005) The response of Alaskan arctic tundra
to experimental warming: Differences between short- and long-term responses.
Global Change Biology 11: 525–536.
Honegger R, Conconi S & Kutasi V (1996) Field studies on growth and regenerative
capacity in the foliose macrolichen Xanthoria parietina (Teloschistales,
Ascomycotina). Botanica Acta 109: 187–193.
Honour SL, Bell JNB, Ashenden TW, Cape JN & Power SA (2009) Responses of
herbaceous plants to urban air pollution: Effects on growth, phenology and leaf
surface characteristics. Environmental Pollution 157: 1279–1286.
Hudson JMG & Henry GHR (2009) Increased plant biomass in a High Arctic heath
community from 1981 to 2008. Ecology 90: 2657–2663.
Hudson JMG & Henry GHR (2010) High Arctic plant community resists 15 years of
experimental warming. Journal of Ecology 98: 1035–1041.
32
Hughes L (2000) Biological consequences of global warming: Is the signal already
apparent? Trends in Ecology and Evolution 15: 56–61.
Hultengren S (1995) Något om lavfloran på en västsvensk ek. Svensk Botanisk Tidskrift
89: 165–170.
Högberg N, Kroken S, Thor G & Taylor JW (2002) Reproductive mode and genetic
variation suggest a North American origin of European Letharia vulpina. Molecular
Ecology 11: 1191–1196.
Janzen DH (1976) Why bamboos wait so long to flower. Annual Review of Ecology
and Systematics 7: 347–391.
Jenkins M (2003) Prospects for biodiversity. Science 302: 1175–1177.
Jim CY (1998) Impacts of intensive urbanization on trees in Hong Kong. Environmental
Conservation 25: 146–159.
Johansson P & Ehrlen J (2003) Influence of habitat quantity, quality and isolation on the
distribution and abundance of two epiphytic lichens. Journal of Ecology 91: 213–
221.
Johansson V, Bergman K-O, Lättman H & Milberg P (2009) Tree and site quality
preferences of six epiphytic lichens growing on oaks in southeastern Sweden.
Annales Botanici Fennici 46: 496–506.
Johst K & Brandl R (1997) The effect of dispersal on local population dynamics.
Ecological Modelling 104: 87–101.
Kappen L & Smith CW (1980) Heat tolerance of two Cladonia species and Campylopus
praemorsus in a hot steam vent area of Hawaii. Oecologia 47: 184–189.
Kari K (2008) Species-specific responses of an alpine plant community under simulated
environmental change. Journal of Vegetation Science 19: 363–372.
Karlén W & Black JL (2002) Estimates of lichen growth-rate in northern Sweden.
Geografiska Annaler 84 A: 225–232.
Keon DB & Muir PS (2002) Growth of Usnea longissima across a variety of habitats in
the Oregon coast range. The Bryologist 105: 233–242.
Krempelhuber A (1867) Geschichte und litteratur der lichenologie von den ältesten zeit
bis zum schlusse des jahres 1865. Selbstverlag, München.
Kärnefelt I & Lättman H (2001) An unusual discovery of Parmelina tiliacea in central
Malmö, Sweden. Graphis Scripta 11: 9–12.
Larsen RS, Bell JNB, James PW, Chimonides PJ, Rumsey FJ, Tremper A & Purvis OW
(2007) Lichen and bryophyte distribution on oak in London in relation to air
pollution and bark acidity. Environmental Pollution 146: 332–340.
Leopold LB (1968) Hydrology for urban land planning ― A guidebook of the
hydrologic effects of urban land use. U.S. Geological Survey Circular 554,
Washington, D.C.
Lindblom L & Ekman S (2006) Genetic variation and population differentiation in the
lichen-forming ascomycete Xanthoria parietina on the island Storfosna, central
Norway. Molecular Ecology 15: 1545–1559.
Lindblom L & Ekman S (2007) New evidence corroborates population differentiation in
Xanthoria parietina. The Lichenologist 39: 259–271.
Lisowska M (2011) Lichen recolonisation in an urban-industrial area of southern Poland
as a result of air quality improvement. Environmental Monitoring Assessment 179:
177–190.
33
Lohr VI, Pearson-Mims CH, Tarnai J & Dillman DA (2004) How urban residents rate
and rank the benefits and problems associated with trees in cities. Journal of
Arboriculture 30: 28–35.
Lorentsson S & Mattsson J-E (1999) New reports of soredia dispersed by ants, Formica
cunicularia. The Lichenologist 31: 204–207.
Lücking R (2003) Takhtajan's floristic regions and foliicolous lichen biogeography: A
compatibility analysis. The Lichenologist 35: 33–54.
Maclean IMD, Austin GE, Rehfisch MM, Blew J, Crowe O, Delany S, Devos K,
Deceuninck B, Günther K, Laursen K, van Roomen M & Whal J (2008) Climate
change causes rapid changes in the distribution and site abundance of birds in
winter. Global Change Biology 14: 2489–2500.
Malcolm JR, Markham A, Neilson RP & Garaci M (2002) Estimated migration rates
under scenarios of global climate change. Journal of Biogeography 29: 835–849.
Matthews JA & Trenbirth HE (2011) Growth rate of a very large crustose lichen
(Rhizocarpon subgenus) and its implications for lichenometry. Geografiska Annaler
93 A: 27–39.
Mattsson J-E, Lättman H & Milberg P (2006) Rapid changes in the epiphytic
macrolichen flora on sites in southern Sweden. The Lichenologist 38: 323–329.
McCune B & Caldwell BA (2009) A single phosphorus treatment doubles growth of
cyanobacterial lichen transplants. Ecology 90: 567–570.
McKinney ML (2006). Urbanization as a major cause of biotic homogenization.
Biological Conservation 127: 247–260.
Middleton JT, Darley EF & Brewer RF (1958) Damage to vegetation from polluted
atmospheres. Journal of the Air Pollution Control Association 8: 9–15.
Montero Alvarez A, Estévez Alvarez JR, Iglesias Brito H, Pérez Arriba O, López
Sánchez D & Wolterbeek HT (2006) Lichen based biomonitoring of air quality in
Havana City west side. Journal of Radioanalytical and Nuclear Chemistry 270: 63–
67.
Muhammadi UH (2011) Spatial pattern of occurrence of eleven epiphytic lichen species
in a heterogeneous landscape. Master thesis, Linköping University, LiTH-IFM-Ex-11/2441--SE.
Munoz J, Felicisimo AM, Cabezas F, Burgaz AR & Martinez I (2004) Wind as a longdistance dispersal vehicle in the southern hemisphere. Science 304: 1144–1147.
Myllys L, Stenroos S, Thell A & Ahti T (2003) Phylogeny of bipolar Cladonia
arbuscula and Cladonia mitis (Lecanorales, Euascomycetes). Molecular
Phylogenetics and Evolution 27: 58–69.
Mylona S (1996) Sulphur dioxide emissions in Europe 1880–1991 and their effect on
sulphur concentrations and depositions. Tellus Series B-Chemical and Physical
Meteorology 48: 662–689.
Nash III TH (ed.) (1996): Lichen Biology. 1st Ed. – Cambridge University Press,
Cambridge.
Nash III TH & Gries C (2002) Introduction. In: Nash III TH, Ryan BD, Gries C &
Bungartz F (Eds.) Lichen flora of the Greater Sonoran Desert Region. Vol. I, pp. 1–
53. Lichens Unlimited, Arizona State University.
Nathan R (2006) Long-distance dispersal of plants. Science 313: 786–788.
NBF (2007) Statistical Yearbook of Forestry 2007. NBF, Jönköping, Sweden.
Nordén B & Appelqvist T (2001) Conceptual problems of ecological continuity and its
bioindicators. Biodiversity and Conservation 10: 779–791.
34
Nowak DJ, Crane DE & Dwyer JF (2002) Compensatory value of urban trees in the
United States. Journal of Arboriculture 28: 194–199.
Nylander W (1866) Les lichens du Jardin du Luxembourg. Bulletin de la Société
Botanique de France 13: 364–372.
Palice Z & Printzen C (2004) Genetic variability in tropical and temperate populations
of Trapeliopsis glaucolepidea: Evidence against long-range dispersal in a lichen
with disjunct distribution. Mycotaxon 90: 43–54.
Pandey V, Upreti DK, Pathak R & Pal A (2002) Heavy metal accumulation in lichens
from the Hetauda industrial area Narayani zone Makwanpur District, Nepal.
Environmental Monitoring and Assessment 73: 221–228.
Parmesan C (2006) Ecological and evolutionary responses to recent climate change.
Annual Review of Ecology Evolution and Systematics 37: 637–669.
Parmesan C & Yohe G (2003) A globally coherent fingerprint of climate change
impacts across natural systems. Nature 421: 37–42.
Perry AL, Low PJ, Ellis JR & Reynolds JD (2005) Climate change and distribution
shifts in marine fishes. Science 308: 1912–1915.
Pharo EJ & Beattie AJ (1997) Bryophyte and lichen diversity: A comparative study.
Australian Journal of Ecology 22: 151–162.
Pharo EJ, Beattie AJ & Binns D (1999) Vascular plant diversity as a surrogate for
bryophyte and lichen diversity. Conservation Biology 13: 282–292.
Pisani T, Paoli L, Gaggi C, Pirintsos SA & Loppi S (2007) Effects of high temperature
on epiphytic lichens: Issues for consideration in a changing climate scenario. Plant
Biosystems 141: 164–169.
Platt RB & Amsler FP (1955) A basic method for the immediate study of lichen growth
rates and succession. Journal of the Tennessee Academy of Science 30: 177–183.
Printzen C, Ekman S & Tønsberg T (2003) Phylogeography of Cavernularia hultenii:
Evidence of slow genetic drift in a widely disjunct lichen. Molecular Ecology 12:
1473–1486.
Proctor MCF (1983) Sizes and growth-rates of thalli of the lichen Rhizocarpon
geographicum on the moraines of the Glacier de Valsorey, Valais, Switzerland. The
Lichenologist 15: 249–261.
Ranius T, Johansson P, Berg N & Niklasson M (2008) The influence of tree age and
microhabitat quality on the occurrence of crustose lichens associated with old oaks.
Journal of Vegetation Science 19: 653–662.
Ranta P & Viljanen V (2011) Vascular plants along an urban-rural gradient in the city
of Tampere, Finland. Urban Ecosystems 14: 361–376.
Rasmussen P (2005) Mid- to late-Holocene land-use change and lake development at
Dallund Sø, Denmark: vegetation and land-use history inferred from pollen data.
The Holocene 15: 1116–1129.
Rogers RW (1971) Distribution of the lichen Chondropsis semiviridis in relation to its
heat and drought resistance. New Phytologist 7: 1069–1077.
Romeike J, Friedl T, Helms G & Ott S (2002) Genetic diversity of algal and fungal
partners in four species of Umbilicaria (lichenized Ascomycetes) along a transect
of the Antarctic peninsula. Molecular Biology and Evolution 19: 1209–1217.
Root TL, Price JT, Hall KR, Schneider SH, Rosenzweig C & Pounds JA (2003)
Fingerprints of global warming on wild animals and plants. Nature 421: 57–60.
Rose F (1974) The epiphytes of oak. In Morris MG & Perring FH (Eds.) The British
oak: its history and natural history, Faringdon, Classey.
35
Rosenzweig C, Karoly D, Vicarelli M, Neofotis P, Wu Q, Casassa G, Menzel A, Root
TL, Estrella N, Seguin B, Tryjanowski P, Liu C, Rawlins S & Imeson A (2008)
Attributing physical and biological impacts to anthropogenic climate change.
Nature 453: 353–358.
Sancho LG, de la Torre R, Horneck G, Ascaso C, de Los Rios A, Pintado A, Wierzchos
J & Schuster M (2007a) Lichens survive in space: Results from the 2005 LICHENS
experiment. Astrobiology 3: 443–454.
Sancho LG, Green TGA & Pintado A (2007b) Slowest to fastest: Extreme range in
lichen growth rates supports their use as an indicator of climate change in
Antarctica. Flora 202: 667–673.
Sanders WB, Moe RL & Ascaso C (2004) The intertidal marine lichen formed by the
pyrenomycete fungus Verrucaria tavaresiae (Ascomycotina) and the brown alga
Petroderma maculiforme (Phaeophyceae): Thallus organization and symbiont
interaction. American Journal of Botany 91: 511–522.
Sanz-Elorza M, Dana ED, Gonzalez A & Sobrino E (2003) Changes in the highmountain vegetation of the central Iberian peninsula as a probable sign of global
warming. Annals of Botany 92: 273–280.
Saxe H, Cannell MGR, Johnsen B, Ryan MG & Vourlitis G (2001) Tree and forest
functioning in response to global warming. New Phytologist 149: 369–399.
Schopp W, Posch M, Mylona S & Johansson M (2003) Long-term development of acid
deposition (1880–2030) in sensitive freshwater regions in Europe. Hydrology and
Earth System Sciences 7: 436–446.
Seaward MRD (1976) Performance of Lecanora muralis in an urban environment. In:
Brown DH, Hawksworth DL & Bailey RH (Eds.) Progress and problems in
lichenology. Academic Press, London.
Shukla V & Upreti DK (2011) Changing lichen diversity in and around urban
settlements of Garhwal Himalayas due to increasing anthropogenic activities.
Environmental Monitoring Assessment 174: 439–444.
Sillett SC, McCune B, Peck JE, Rambo TR & Ruchty A (2000) Dispersal limitations of
epiphytic lichens result in species dependent on old-growth forests. Ecological
Applications 10: 789–799.
SMHI (1986–2003) Väder och vatten. -Swedish Meteorological and Hydrological
Institute, Norrköping, Sweden.
Spier L, van Dobben H & van Dort K (2010) Is bark pH more important than tree
species in determining the composition of nitrophytic or acidophytic lichen floras?
Environmental Pollution 158: 3607–3611.
Stubbs CS (1995) Dispersal of soredia by the oribatid mite, Humerobates arborea.
Mycologia 87: 454–458.
Søchting U & Olech M (1995) The lichen genus Caloplaca in polar regions. The
Lichenologist 27: 453–471.
Tapper R (1976) Dispersal and changes in the local distributions of Evernia prunastri
and Ramalina farinacea. New Phytologist 77: 725–734.
Taylor TN (1995) The oldest fossil lichen. Nature 378: 244.
The World Conservation Union (2010) IUCN Red list of threatened species. Summary
Statistics for Globally Threatened Species. Table 1: Numbers of threatened species
by major groups of organisms (1996–2010).
Tibell LB (1994) Distribution patterns and dispersal strategies of Caliciales. Botanical
Journal of the Linnean Society 116: 159–202.
36
Thomas CD (2010) Climate, climate change and range boundaries. Diversity and
Distributions 16: 488–495.
Thomas CD, Cameron A, Green RE, Bakkenes M, Beaumont LJ, Collingham YC,
Erasmus BFN, Ferreira de Siqueira M, Grainger A, Hannah L, Hughes L, Huntley
B, van Jaarsveld AS, Midgley GF, Miles L, Ortega-Huerta MA, Townsend
Peterson A, Phillips OL & Williams SE (2004) Extinction risk from climate
change. Nature 427: 145–148.
Thuiller W, Lavorel S, Araujo MB, Sykes MT & Prentice IC (2005) Climate change
threats to plant diversity in Europe. Proceedings of the National Academy of
Sciences of the United States of America 102: 8245–8250.
Tripathi AK & Gautam M (2007) Biochemical parameters of plants as indicators of air
pollution. Journal of Environmental Biology 28: 127–132.
Turner IM (1996) Species loss in fragments of tropical rain forest: A review of the
evidence. The Journal of Applied Ecology 2: 200–209.
UNECE (1999) Protocol to the 1979 convention on long-range transboundary air
pollution to abate acidification, eutrophication, and ground-level ozone. United
Nations Economic Commission for Europe, Geneva, Switzerland.
Vitousek PM (1994) Beyond global warming: Ecology and global change. Ecology 75:
1861–1876.
Vitousek PM, Mooney HA, Lubchenco J & Melillo JM (1997) Human domination of
earth’s ecosystems. Science 277: 494–499.
Wagner HH, Werth S, Kalwij JM, Bolli JC & Scheidegger C (2006) Modelling forest
recolonization by an epiphytic lichen using a landscape genetic approach.
Landscape Ecology 21: 849–865.
Walser JC (2004) Molecular evidence for limited dispersal of vegetative propagules in
the epiphytic lichen Lobaria pulmonaria. American Journal of Botany 91: 1273–
1276.
Walser JC, Holderegger R, Gugerli F, Hoebee SE & Scheidegger C (2005)
Microsatellites reveal regional population differentiation and isolation in Lobaria
pulmonaria, an epiphytic lichen. Molecular Ecology 14: 457–467.
Walther G-R, Post E, Convey P, Menzel A, Parmesank C, Beebee TJC, Fromentin J-M,
Hoegh-Guldberg O & Bairlein F (2002) Ecological responses to recent climate
change. Nature 416: 389–395.
Warren MS, Hill JK, Thomas JA, Asher J, Fox R, Huntley B, Roy DB, Telfer MG,
Jeffcoate S, Harding P, Jeffcoate G, Willis SG, Greatorex-Davies JN, Moss D &
Thomas CD (2001) Rapid responses of British butterflies to opposing forces of
climate and habitat change. Nature 414: 65–69.
Washburn SJ & Culley TM (2006) Epiphytic macrolichens of the greater Cincinnati
metropolitan area―Part II: Distribution, diversity and urban mycology. The
Bryologist 109: 516–526.
Wedin M, Döring H & Gilenstam G (2004) Saprotrophy and lichenization as options for
the same fungal species on different substrata: Environmental plasticity and fungal
lifestyles in the Stictis-Conotrema complex. New Phytologist 164: 459–465.
Werth S, Wagner HH, Gugerli F, Holderegger R, Csencsics D, Kalwij JM &
Scheidegger C (2006a) Quantifying dispersal and establishment limitation in a
population of an epiphytic lichen. Ecology 87: 2037–2046.
37
Werth S, Wagner HH, Holderegger R, Kalwij JM & Scheidegger C (2006b) Effect of
disturbances on the genetic diversity of an old-forest associated lichen. Molecular
Ecology 15: 911–921.
Wielgolaski FE (1975) Biological indicators on pollution. Urban Ecology 1: 63–79.
Zalasiewicz J, Williams M, Smith A, Barry TL, Coe AL, Bown PR, Brenchley P,
Cantrill D, Gale A, Gibbard P, Gregory FJ, Hounslow MW, Kerr AC, Pearson P,
Knox R, Powell J, Waters C, Marshall J, Oates M, Rawson P & Stone P (2008) Are
we now living in the Anthropocene? Geological Society of America 2: 4–7.
Öckinger E, Niklasson M & Nilsson SG (2005) Is local distribution of the epiphytic
lichen Lobaria pulmonaria limited by dispersal capacity or habitat quality?
Biodiversity and Conservation 14: 759–773.
38
PAPER I
The Lichenologist 41(5): 557–559 (2009) © 2009 British Lichen Society
doi:10.1017/S0024282909990259 Printed in the United Kingdom
Short Communication
Generation time estimated to be 25–30 years in Cliostomum
corrugatum (Ach.) Fr.
Knowledge of spore to spore generation time
is extremely important for several reasons. As
it is the shortest generation time, it indicates
the maximum nucleotide substitution rate
over time and provides a rate limit for the
evolution of a species. In population genetics
most calculations involving time use ‘generations’ as the unit of measurement and in
order to convert these ‘generations’ into
‘years’, knowledge of generation time is
needed but rarely available. Knowledge of
generation time may also be essential for
conservation purposes and assessments of
migration history. This knowledge also
makes it possible to estimate both the age of a
population and also to determine to what
extent a population represents the genetic
diversity of a species (Rosenberg & Nordborg
2002). In this paper we present a method for
assessing generation length for lichens using
Cliostomum corrugatum (Ach.) Fr. as an
example. This lichen was selected for investigation because it is restricted to forests with
long temporal continuity (Lättman et al.
2009) and information on generation time is
essential to estimate the rate of dispersal at
the landscape level.
Cliostomum corrugatum is an epiphytic
crustose lichen with distinctive morphology
and clearly delimited habitats making it relatively easy to find and identify in the field. It
has a greyish thallus which bears conspicuous, light yellow to light brown, apothecia,
0·5–1·2 mm wide, as well as black pycnidia,
0·2–0·5 mm wide. Its geographical distribution in Sweden largely reflects that of
Quercus robur L. as it is chiefly confined to the
coarse bark of old trunks in old stands of
this tree species, often in relatively dry and
semi-open forests or parklands (Thor &
Arvidsson 1999). It occurs mainly on the flat
outer parts of the bark but is occasionally
found on the coarse bark of other deciduous
trees and on decorticated stumps, old wood,
and twigs of Picea abies L. In northern
Europe it is rare and red-listed, for example,
in Sweden, Near Threatened; in Norway,
Critically Endangered, and in Finland, Near
Threatened (Lättman et al. 2009). Cliostomum corrugatum produces sexual spores but
not vegetative diaspores, such as soredia or
isidia, and we assume that new thalli establish from sexually produced spores. Based on
results of Lättman et al. (2009) we assume no
dispersal limitation at the spatial scale
studied. Suitable habitats, i.e. old oaks will
probably be colonized as soon as their bark is
coarse enough to be a suitable habitat.
On a site in the province of Östergötland, (southcentral Sweden) all Quercus robur trunks >200 cm circumference were examined in April 2006 from ground
level to 2 m in a search for Cliostomum corrugatum.
Preliminary studies had shown a very low occurrence of
C. corrugatum on trees with a circumference < 200 cm in
agreement with later studies (Johansson et al. 2009;
Ranius et al. 2008). The occurrence and the size of the
largest C. corrugatum thalli with apothecia and the size of
the largest apothecia were recorded. Data for the largest
thallus on each trunk was used to determine the smallest
trunk size colonized by C. corrugatum. The relationship
between the square root of the thallus area and trunk
circumference was examined using linear regression
analysis in R 2.7.2 (R development core team, 2005)
and the average circumference at the time of spore
germination was determined using the point at which
the regression line intercepts the x-axis. The minimum
circumference value was compared to that of the smallest trunk with fertile lichens. The time (in years)
between spore germination and spore production
was subsequently calculated by dividing the difference
(in mm) between this circumference value and the
558
THE LICHENOLOGIST
Vol. 41
circumference of the thinnest trunk carrying lichen by 2
multiplied by the annual radial growth.
Carlsson (2004) found an annual radial growth of
1·6–2·4 mm of oaks at a single location with a circumference of 200 cm. Compared to this Carbonnier (1975)
gives a value of 2·4 mm on good soils in southern Sweden and an assessment based on data in Tree-ring Database, NOOA Paleoclimatology (http://www.ncdc.noaa.
gov/paleo/paleo.html) gives a value of 1.6 mm, with a
standard deviation of 1.0 mm (28088 measurements
from 9 data sets referring to Europe north of latitude 54
after the year 1700). As the soil conditions on the
studied site are good it seems reasonable to assume an
annual radial growth of 2 mm.
Seventeen large Quercus robur trunks were
examined. Trunks with a circumference
>350 cm were not included because cracks in
the bark that limit the growth of thalli in
some directions, are most frequent on older
trees. In addition, as the life span of lichens is
limited, older trees may support thalli representing later colonizations. Also, the comparatively small thallus (40 cm2) on a trunk
with a perimeter of 280 cm was omitted as it
probably represents a late colonization event.
The linear regression analysis gave a minimum value for oak circumference at which
colonization by C. corrugatum occurs of
215 ± 16 cm, P=0·001, r2=0·826 (Fig. 1).
The smallest circumference of trunks with
fertile specimens of C. corrugatum was 240
cm. During the time from colonization to
formation of apothecia an oak will increase
its perimeter by 25 ± 16 cm. The time span
from spore to spore may thus be calculated to
approximately 20 years (250/(2 × 2) = 20).
If the standard error of the average circumference (16 cm) is taken into account the
time span will be 7–25 years.
The observations supported the assumption of a correlation between thallus age, as
indicated by thallus size, and the formation
of apothecia because all fertile thalli were at
least 30 cm2. There was also a correlation
between tree age and occurrence of thalli
indicating good dispersal properties to appropriate substrata of a certain age, although
one tree was thought to have been colonized
late. Of course, other factors, such as microclimatic conditions might affect dispersal so
that specific values (e.g. of minimum circumference) might only be relevant to the site.
F. 1. The relationship of the size of the largest thallus
of Cliostomum corrugatum on each Quercus robur and the
circumference of the trunk.
Small thalli with well-developed apothecia
were not encountered, indicating a correlation between thallus size and formation of
apothecia. Further, as apothecia are absent
on the smallest trees this also indicates relatively rapid colonization when the trees
develop a suitable bark structure for colonization. All these observations suggest that it
would be appropriate to apply the results to a
larger geographic area.
The method described here uses observations on the smallest/youngest tree with sterile thalli, thus identifying the shortest
possible generation length, rather than average spore to spore generation length which is
probably significantly longer. Within population genetics the latter value would be more
useful. It seems reasonable to assume an
average length of about 25–30 years. On the
other hand, a minimum value also indicates
possibilities. It shows the maximum rate of
evolutionary processes and in this case also
indicates a fairly high average rate as all thalli
on bigger oaks are fertile.
The importance of other means of dispersal, for example, vegetative diaspores, should
be considered. Since no soredia or isidia have
been reported in C. corrugatum, it seems
reasonable to assume meiospore dispersal as
all larger thalli possess apothecia. There is of
course a possibility of production of nonviable spores in the apothecia observed but it
seems reasonable to assume functionality.
2009
Short Communication
The pycnidia of the thallus produce mitotic conidia that may also function as spermatia, fungal diaspores, or maybe both, as
has been known from studies since the 19th
century (Vobis 1980). As far as we know
none of these studies concerns C. corrugatum,
but the possibility of dispersal by pycnoconidia should not be neglected. If pycnoconidia act as dispersal units the shortest
generation time from spore to spore may be
underestimated by our method.
This study is part of the multi-disciplinary project “Ecological and Societal Systems in Interaction” a multidisciplinary project aiming at identifying and describing
processes behind environmental and conservation policies of public authorities at different levels in the Baltic
Sea area. The project is hosted at and is a part of the
strategic development of environmental sciences at
Södertörn University and is generously funded by
Östersjöstiftelsen. Economic support from the regional
county administration (Stockholm läns landsting) is also
acknowledged. The leader of the course Field Studies of
the Biological program of Södertörn University, Mikael
Lönn, made student participation and inclusion of this
project in the course possible. Finally, we are also grateful to Jörg Brunet for help with references and to Stefan
Ekman for the calculations of the average annual growth
from NOOA data. Together with two unknown referees,
the latter also contributed useful comments which
greatly improved the manuscript. Finally we thank
Anthony Wright for linguistic revision of the text.
R
Carbonnier, C. (1975) Produktionen i kulturbestånd av
ek i södra Sverige. Studie Forestalia Suecica 125.
Skogshögskolan, Stockholm.
Carlsson, J. (2004) Betydelsen av ålder och tillväxt hos
ek (Quercus robur) för förekomst av rödlistade arter i
Östergötland. LiU-IFM-Biol-Ex-1137. Student
dissertation, Linköping University.
559
Johansson, V., Bergman, K.-O., Lättman, H. &
Milberg, P. (2009) Tree and quality preferences of
six epiphytic lichens growing on oaks in south eastern Sweden. Annales Botanici Fennici (in press).
Lättman, H., Lindblom, L., Mattsson, J.-E., Millberg,
P., Skage, M. & Ekman, S. (2009) Estimating the
dispersal of the rare lichen Cliostomum corrugatum.
Biological Conservation (in press).
R development core team. (2005) R: a language and
environment for statistical computing. http://cran.rproject.org
Ranius, T., Johansson, P., Berg, N. & Niklasson, M.
(2008) The influence of tree age and microhabitat
quality on the occurrence of crustose lichens associated with old oaks. Journal of Vegetation Science 19:
653–662.
Rosenberg, N. A. & Nordborg, M. (2002) Genealogical
trees, coalescent theory and the analysis of genetic
polymorphisms. Nature Reviews Genetics 3: 380–
390.
Thor, G. & Arvidsson, L. (1999) Swedish Red Data Book
of Lichens. Uppsala: ArtDatabanken, SLU.
Vobis, G. (1980) Bau und Entwicklung der FlechtenPycnidien und ihrer Conidien. Bibliotheca Lichenologica 14.
Håkan Lättman, Anneli Brand,
Johanna Hedlund, Mikael Krikorev,
Niklas Olsson, Alexandra Robeck,
Fredrik Rönnmark and
Jan-Eric Mattsson
H. Lättman: School of Life Sciences, Södertörn University, SE-141 89 Huddinge, Sweden and . IFM Division
of Ecology, Linköping University, SE-581 83
Linköping, Sweden.
A. Brand, J. Hedlund, M. Krikorev, N. Olsson, A.
Robeck, F. Rönnmark and J.-E. Mattsson (corresponding author): School of Life Sciences, Södertörn University, SE-141 89 Huddinge, Sweden. Email: jan-eric.
[email protected]
PAPER II
Biological Conservation 142 (2009) 1870–1878
Contents lists available at ScienceDirect
Biological Conservation
journal homepage: www.elsevier.com/locate/biocon
Estimating the dispersal capacity of the rare lichen Cliostomum corrugatum
Håkan Lättman a,b, Louise Lindblom c,f, Jan-Eric Mattsson a, Per Milberg b,g, Morten Skage c,d, Stefan Ekman c,e,*
a
School of Life Sciences, Södertörn University College, SE-181 49 Huddinge, Sweden
IFM Division of Ecology, Linköping University, SE-581 83 Linköping, Sweden
Department of Biology, University of Bergen, P.O. Box 7800, NO-5020 Bergen, Norway
d
Department of Biology, University of Oslo, P.O. Box 1066, NO-0316 Oslo, Norway
e
Museum of Evolution, Evolutionary Biology Centre, Uppsala University, Norbyvägen 16, SE-752 36 Uppsala, Sweden
f
Museum of Natural History, University of Bergen, P.O. Box 7800, NO-5020 Bergen, Norway
g
Department of Crop Production Ecology, SLU, Box 7043, 750 07 Uppsala, Sweden
b
c
a r t i c l e
i n f o
Article history:
Received 31 October 2008
Received in revised form 6 March 2009
Accepted 21 March 2009
Available online 5 May 2009
Keywords:
Dispersal
Establishment
Ecological continuity
Old-growth forests
Quercus
Ascomycete
a b s t r a c t
The objective of this study was to estimate the dispersal rate in an organism assumed to be confined to
tree stands with unbroken continuity. We used the lichen-forming ascomycete Cliostomum corrugatum,
which is largely confined to old oak stands. Five populations, with pairwise distances ranging from 6.5
to 83 km, were sampled in Östergötland, south-eastern Sweden. DNA sequence data from an intron in
the small subunit nuclear ribosomal RNA gene was obtained from 85 samples. Nearly all molecular variance (99.6%) was found within populations and there were no signs of isolation-by-distance. The absolute number of immigrants per population per generation (estimated to 30 years), inferred by Bayesian
MCMC, was found to be between 1 and 5. Altogether, evidence suggests abundant gene flow in the history
of our sample. A simulation procedure demonstrated that we cannot know whether effective dispersal is
ongoing or if it ceased at the time when oaks started to decrease dramatically around 400 years BP. However, a scenario where effective dispersal ceased already at the time when the postglacial reinvasion of
oak had reached the region around 6000 years BP is unlikely. Vegetation history suggests that the habitat
of C. corrugatum was patchily distributed in the landscape since the early Holocene. Combined with the
high dispersal rate estimate, this suggests that the species has been successful at frequently crossing distances of at least several kilometres and possibly that it has primarily been limited by the availability of
habitat rather than by dispersal.
Ó 2009 Elsevier Ltd. All rights reserved.
1. Introduction
Many organisms, belonging to a variety of taxonomic groups
like wood-decay fungi, lichens, bryophytes, vascular plants, and insects, seem to be confined to habitat patches that have persisted
presumably unchanged over an extended period of time (Berg
et al., 1994; see Nordén and Appelqvist, 2001, p. 781 for references
to specific organism groups). The concept ‘ecological continuity’
(EC), coined by Rose (1974), has been used to refer to the temporally unbroken continuity of such habitat, often assumed to be primeval or old-growth forests. It has also been proposed that certain
species can be used as indicators of EC (e.g., Rose, 1974; Tibell,
1992; Selva, 1994, 2003; Kuusinen, 1996; Økland, 1996; Lindblad,
1998) when historical data are difficult to obtain. The EC concept
* Corresponding author. Address: Museum of Evolution, Evolutionary Biology
Centre, Uppsala University, Norbyvägen 16, SE-752 36 Uppsala, Sweden. Tel.: +46
18 471 28 21; fax: +46 471 27 94.
E-mail addresses: [email protected] (H. Lättman), [email protected]
(L. Lindblom), [email protected] (J.-E. Mattsson), [email protected]
(P. Milberg), [email protected] (M. Skage), stefan.ekman@evolmuseum.
uu.se (S. Ekman).
0006-3207/$ - see front matter Ó 2009 Elsevier Ltd. All rights reserved.
doi:10.1016/j.biocon.2009.03.026
has been criticized for often being vaguely defined on spatial and
temporal scales (Gauslaa and Ohlson, 1997; Nordén and Appelqvist, 2001; Sverdrup-Thygeson and Lindenmayer, 2003). In most
real applications, EC implicitly refers to the forest stand level.
Using indicator species to assess EC is also problematic, for several
reasons: the group of species claimed to indicate EC probably includes species with poor dispersal capacity as well as species with
particular microhabitat requirements. Their dispersal capacity and
habitat requirements are often poorly or not at all understood
(Nordén and Appelqvist, 2001; Rolstad et al., 2002). Forest history
and dynamics is poorly known and often judged from anecdotal
evidence (Rolstad et al., 2002). The spatial scale at which indicators
are assumed to work is usually undefined, the spatial precision of
the indicator species being dependent on dispersal capacity (Rolstad et al., 2002; Sverdrup-Thygeson and Lindenmayer, 2003; Kalwij et al., 2005). Finally, it cannot be uncritically assumed that
species richness or the number and abundance of rare species is
strictly positively correlated with temporal continuity (Ohlson
et al., 1997; Fenton and Bergeron, 2008; Lõhmus and Lõhmus,
2008). However, there is ample evidence that some species are
H. Lättman et al. / Biological Conservation 142 (2009) 1870–1878
indeed confined to EC habitats and that red-listed species or certain taxonomic groups are represented by more species in olderthan-average and unmanaged forests compared to younger and
managed ones (e.g., Gustafsson and Hallingbäck, 1988; Goward,
1994; Fritz and Larsson, 1996; Spence et al., 1996; Nilsson and Baranowski, 1997; Martikainen et al., 1999, 2000; Uliczka and Angelstam, 1999; Hedenås and Ericson, 2000; Cameron, 2002; Penttilä
et al., 2004; Tikkanen et al., 2006; Rivas Plata et al., 2008; Fritz
et al., 2008).
A central question is why some organisms are confined to sites
with long temporal continuity, or at least prefer them. There are
two main, not mutually exclusive, explanations for this phenomenon: (1) limitation by dispersal, dispersal primarily taking place
only at very short distances and being virtually absent at larger distances, and (2) limitation by habitat availability. If dispersal is the
primary limiting factor, occurrence in EC habitats is likely to be of a
relictual nature, whereas this is not necessarily the case when habitat availability is the primary limiting factor. Limitation by habitat
availability is indeed a realistic phenomenon, as the structural
complexity and consequently the number of microhabitats in a forest has been shown to be higher under old-growth conditions
(Zenner, 2004). In both cases, preserving currently occupied
patches and creating new habitat may be necessary for the longterm preservation of an organism restricted to EC habitats. However, the distances that can be allowed in this network of currently
and potentially occupied patches depend crucially on the dispersal
capabilities of the organism in question. Unfortunately, a scientifically based knowledge of effective dispersal (i.e. dispersal followed
by establishment) at the landscape level is currently missing in
most species. This includes also species restricted to EC habitats,
many of which are also red-listed and in need of proper management for long-term persistence.
The objective of this study was to estimate the rate of dispersal
at the landscape level in an organism restricted to forests with long
temporal continuity. We selected the lichen-forming ascomycete
Cliostomum corrugatum (Ach.) Fr. as a study species. C. corrugatum
is a rare lichen that is largely confined to very old stands of Quercus
robur (Berg et al., 2002) and has been suggested to be an indicator
of EC (Arup, 1997). We used DNA sequence data from a nuclear
marker, combined with a population genetics approach, to address
the question at hand. We are not aware of any previous investigations of genetic variation at small spatial scales in a crustose (crustforming) lichen.
2. Materials and methods
2.1. Study species
The epiphytic crustose lichen C. corrugatum (Ach.) Fr. (Ascomycota, Lecanoromycetes, Lecanorales, Ramalinaceae) possesses a
greyish thallus containing a green alga as its photosynthesizing
symbiont. The thallus bears conspicuous, light yellow to light
brown, 0.5–1.2 mm wide apothecia (fruiting bodies producing
putatively meiotic ascospores) as well as black, 0.2–0.5 mm wide
pycnidia (producing mitotic conidia that may function as spermatia, fungal diaspores, or both) (Thor and Arvidsson, 1999). In Sweden, its geographical distribution largely follows that of Quercus. Its
primary habitat is coarse bark of old trunks of Q. robur in relatively
dry and semi-open forests or parklands (Thor and Arvidsson,
1999), mainly on the flat terminal parts of the bark structure and
not on either side of the cracks in the bark (pers. obs.). C. corrugatum has occasionally been encountered on the coarse bark of other
deciduous trees as well as on wood of decorticated stumps, old
wood structures, and twigs of Picea abies (Thor and Arvidsson,
1999). No vegetative diaspores (soredia or isidia) containing tissue
from both symbionts are produced (Thor and Arvidsson, 1999).
1871
C. corrugatum, like all ascomycetes except some yeasts, is presumed to have a dominantly haploid life cycle. Dikaryotic and diploid stages appear only as very small amounts of hyphae confined
to the apothecia. C. corrugatum is rare in northern Europe and
red-listed in, e.g., Sweden (Near Threatened; Gärdenfors, 2005),
Norway (Critically Endangered; Kålås et al., 2006), Denmark (‘Vulnerable’ but not evaluated according to recent IUCN criteria; Stoltze and Pihl, 1998), Finland (Near Threatened; Rassi et al., 2001),
Germany (‘Critically Endangered’ but not evaluated according to
recent IUCN criteria; Ludwig and Schnittler, 1996), and the United
Kingdom (Vulnerable; Woods and Coppins, 2003). The lichen is
small but its distinctive morphology and habitat makes it relatively
easy to detect in the field.
2.2. Sample area and sites
Samples of C. corrugatum were collected on tree trunks of Q. robur up to 2 m above the ground between January 5 and February 4
2005 at five sites in central Östergötland, south-eastern Sweden
(Fig. 1; Table 1). Quercus colonized this area approximately
7000 years BP (Brewer et al., 2002). Central Östergötland supports
one of the highest densities of old oaks in Sweden. Altitudinal differences in this region are small and the soil is fertile and consists
of sedimentary silt and clay particles deposited during or after the
end of the last glaciation. The ice retreated from this region
10,000 years ago (Lundqvist, 1998), but due to land depression,
all sites were initially below sea level. Pairwise distances between
sites were evenly distributed and ranged from 6.5 to 83 km. The
smallest tree trunk inhabited by C. corrugatum, out of all investigated, was 0.65 m in diameter at breast height (dbh). Johansson
et al. (in press) demonstrated a positive correlation between the
probability of occurrence of C. corrugatum and tree trunk size of
Q. robur. Owing to difficulties determining the boundaries between
adjacent lichen thalli, only one sample was taken per tree to avoid
the risk of sampling the same individual twice. The minimum
number of samples per site was set to 15.
2.3. Laboratory methods and sequence editing
Methods for DNA extraction, PCR amplification, and sequencing
followed Lindblom and Ekman (2006), except that we mainly used
apothecial tissue (or tissue from the thallus or pycnidia, when apothecia were not available). We first targeted the internal transcribed spacer region (ITS) and the intergenic spacer (IGS) of the
nuclear ribosomal DNA. Because of low variability, we turned our
attention to the group 1 intron situated between positions 1516
and 1517 at the end of the small subunit (SSU) of the nuclear ribosomal RNA gene (Gargas et al., 1995). This region was amplified
using the forward primer ITS1F (White et al., 1990), situated at
the end of the SSU but upstream of the intron site, and the newly
designed reverse primer ITS1-Cc1-R, situated in the first part of
ITS1. The new primer was designed because the widely used combination of ITS1F and ITS4 (Gardes and Bruns, 1993) to amplify the
entire ITS region in many cases either failed or resulted in multiple
PCR products. The sequence of the new primer is 50 -ATG GTA AGG
TAA TCA CAG GGT GTA-30 . The amplification, PCR clean-up, and
sequencing procedures were identical to the ones used by Lindblom and Ekman (2006) for the ITS region. Sequencing was performed using both PCR primers. The technique used here,
particularly when using overlapping forward and reverse reads,
has by far the lowest error rate of any currently used sequencing
procedure (Johnson and Slatkin, 2007). Only sequences from reads
with a low level of noise relative to the signal were processed for
further analysis. Resulting sequences were manually edited and
aligned using BioEdit version 7.0.5.3 (Hall, 1999). The identity of
the sequences obtained with the new primer pair was confirmed
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H. Lättman et al. / Biological Conservation 142 (2009) 1870–1878
Fig. 1. The five sample sites of Cliostomum corrugatum in the province of Östergötland, southern Sweden, indicated with black dots.
Table 1
Geographical location and short description of the five sites where Cliostomum corrugatum was sampled in Östergötland, south-eastern Sweden.
Site
Latitude/longitude
Description
Bjärka-Säby
Orräng
Solberga
Bråborg
Stegeborg
58°160 29.9}N, 15°440 24.0}E
58°170 32.2}N, 15°510 19.9}E
58°200 19.2}N, 15°110 49.8}E
58°360 56.5}N, 16°220 00.4}E
58°260 17.2}N, 16°350 58.0}E
Parkland mixed with arable land
Grassland with patches of open oak stands
Forest adjacent to river
Mixed forest in NE facing slope to Baltic sea
Small patchy tree stands in agricultural landscape
by comparing with sequences initially generated with ITS1F and
ITS4. The ITS part of these sequences were in turn subjected to a
BLAST (megablast) search against the nr/nt database at the National
Center
for
Biotechnology
Information
(http://
www.ncbi.nlm.nih.gov) on 22 October 2008. The top 69 hits
against the amplified ITS were taxa in the Ramalinaceae (expectation scores 63 10137 and query coverage P72%). Each haplotype sequence was submitted to GenBank and given accession
numbers EU218541-EU218551.
2.4. Statistical analyses
In order to create a simple overview of haplotype relationships
and frequencies present in the sample, we constructed a haplotype
network using the 95% statistical parsimony criterion as implemented in the software TCS version 1.13 (Clement et al., 2000). A
likelihood model was fitted to the data using a hierarchical likelihood ratio test as implemented in MODELTEST version 3.7 (Posada
and Crandall, 1998). The best model reported was HKY85 without
rate heterogeneity. Using Arlequin version 3.1 (Schneider et al.,
2000), we first conducted a Ewens–Watterson–Slatkin exact test
of selective neutrality in the entire sample (Slatkin, 1994, 1996).
The null distribution under neutrality was obtained by simulating
the null hypothesis 10,000 times (Stewart, 1977). We conducted an
analysis of molecular variance (AMOVA) (Excoffier et al., 1992),
assessing significance by 10,000 permutations. A Mantel matrix
correlation test between pair-wise values of FST/(1 FST) (Slatkin’s
linearized FST) and pair-wise values of the natural logarithms of
geographic distance between populations (Rousset, 1997) was
used to check for the presence of isolation-by-distance (Wright,
1943), i.e. a spatial aggregation of genetically similar individuals.
Significance was assessed by 10,000 permutations. Wherever
applicable, Tamura distances were used, as they most closely corresponded to HKY85. In all calculation involving Arlequin, indels
were given weight 1 (i.e., they were counted as the ‘fifth character
state’).
The main part of our analysis, however, consisted of a direct
estimation of population parameters, including dispersal, using
Bayesian Markov chain Monte Carlo (MCMC), as implemented in
the software LAMARC 2.0.2 (Kuhner et al., 2005; Kuhner, 2006;
Beerli, 2006; Kuhner and Smith, 2007). In its official terminology,
LAMARC measures ‘migration’. This word is usually interpreted
as one or more individuals contributing to Ne (the effective population size) in one population leaving that population and entering
a new population. In our case, individuals are sessile and dispersal
is expected to occur via ascospores or perhaps conidia, without any
individuals contributing to Ne ever moving between populations.
However, migration in the true sense is not a prerequisite for population parameter estimates by LAMARC to be valid for dispersing
sessile organisms (Peter Beerli and Lucian Smith pers. comm.
2007). This is because the model in LAMARC decouples migration
from size fluctuations, and anyway restores population size to its
original size when individuals migrate. Bayesian MCMC has the
advantage of explicitly handling uncertainty in parameter
H. Lättman et al. / Biological Conservation 142 (2009) 1870–1878
estimates, which is all the more important when the amount of
data, as in this case, is small and uncertainty about estimates can
consequently be expected to be large. In the Bayesian MCMC, the
F84 model was used, because this is the model implemented in
LAMARC that most closely corresponds to HKY85. The transition
to transversion ratio, which is treated as fixed by LAMARC, was calculated under maximum likelihood using PAUP* 4.0b10 (Swofford,
2003) on a collapsed version of our dataset. For this estimate (ratio = 7.01), empirical nucleotide frequencies were used under the
F84 likelihood model. A void population was added to account
for ‘ghost populations’ (Slatkin, 2005) and unsampled populations,
following the recommendation by Kuhner (2003) and Beerli
(2004). Prior distributions were set to uniform in linear space on
the interval [103, 104] for migration and uniform in logarithmic
space on the interval [108, 10] for the population mutation rate
h (corresponding to the extreme lower and upper boundaries allowed by the software). These priors essentially meant assuming
that small and large values of migration are equally likely a priori,
whereas small values of h are more likely than large values a priori.
The prior distribution of population size fluctuations, when applicable, was uniform on the interval [500, 1000] in linear space
(LAMARC allows only linear priors for size fluctuations). The proposal rate for population parameters was set to ten times the proposal rate for genealogy rearrangements. This was necessary to
alleviate problems with poor effective sample sizes of population
parameter estimates, particularly h. Preliminary runs with adaptive
heating indicated that Metropolis coupling, the use of heated
chains, was unnecessary. However, we discovered that population
parameter estimates were more precisely repeatable when using
heating. Consequently, all subsequent runs were conducted with
one heated chain at a temperature of 1.1, allowing information
from the heated chain to be swapped into the cold chain every
10 generations. LAMARC by default treats population parameters
as unconstrained. This means that when h, migration, and size fluctuation are estimated jointly, there is one h and one size fluctuation
parameter for each population as well as one migration parameter
in each direction between each pair of populations (migration is
treated as asymmetric). With five sampled populations and one
void population, as in our case, this amounts to a large total number of parameters that may not be supported by the data. Therefore, we performed tests of model adequacy, which involved the
use of Bayes factor (Kass and Raftery, 1995) to compare models
based on the harmonic mean estimator (Newton et al., 1994).
Starting with simple models, we added parameters only if there
was ‘strong’ support (Kass and Raftery, 1995, p. 777) for a more
complex model, i.e. if twice the difference in harmonic mean ln
likelihood exceeded six. For h, we tested a model where all values
are identical against an unconstrained model. Size fluctuation was
either set to zero, treated as equal across all populations, or unconstrained. Migration was set to zero, treated as equal across all populations, as different between population pairs but symmetric, or
unconstrained (asymmetric between population pairs). In all cases,
the void population was treated as unconstrained, because we do
not know how many real-world populations it represents. Because
LAMARC 2.0.2 only reports the data ln likelihood for the last sample of the MCMC chain, we created a workaround by splitting the
analysis into several consecutive ‘initial chains’, the likelihood
being reported at the end of each such chain. We allowed 42,500
initial chains, each 200 generations long, i.e. a total of 8.5 million
generations. Software was written in RealBasic to extract data ln
likelihoods from the output (‘outsumfile’). Likelihoods were subsequently imported in a Microsoft Excel spreadsheet, likelihoods
plotted, and the harmonic mean ln likelihood calculated across
the stationary phase of the run. Plotting ln likelihood against generation indicated that the true burn-in was in the order of
100,000–150,000 generations, but we anyway discarded the first
1873
500,000 generations. Using this scheme, we arrived at a model
treating size fluctuations as absent (set to zero), and h and migration as equal across all populations. This does not mean that the
true scenario was this simple, only that the current data contained
no information to support a more complex model. Final estimates
of h and migration were obtained by summing results across three
identical runs, each 8.5 million generations in length and discarding the initial 500,000 generations as burn-in. In the haploid case,
h = 2Nel, where l is the per site mutation rate per generation.
Migration is measured as M = m/l, where m is the proportion of
immigrants into a population per generation. Software was written
in RealBasic in order to extract the posterior distribution of
Nem = hM/2, the absolute number of immigrants into a population
per generation, from the joint distribution of h and M. The unimodal posterior probability distributions were finally transformed
into 95% equal-tail credible intervals by removing 2.5% of the total
probability at each end of the posterior probability distribution.
Finally, we performed a simulation study using SimCoal 2.1.2
(Laval and Excoffier, 2004), with the purpose of evaluating the
temporal information contained in the migration rate estimates
obtained by LAMARC under different demographic histories. LAMARC assumes migration rates to be constant over time, from the
present to coalescence of the sample, but this is rarely the case
in real populations. Therefore, estimated migration rates might
be averages over recent evolutionary time, with limited information about ongoing migration. We wanted to answer two specific
questions: can we separate between a model with and a model
without migration after the start of the dramatic decrease of oaks
a few hundred years ago? Similarly, can we separate between a
model with and a model without migration after the immigration
of oaks to Östergötland 6000 years ago? We assumed a generation
time of 30 years for C. corrugatum, based on a combination of the
demography and phenology of the lichen as well as the growth rate
of the inhabited oak trees (Lättman et al., unpublished results). The
demographic history was divided into three phases (in backward
time): (1) the first 13 generations, corresponding to the time during which old oaks decreased dramatically in the region (Eliasson
and Nilsson, 2002); (2) generations 14–200, corresponding to the
period limited by the immigration of oak to the region; (3) generations 201 until coalescence, corresponding to the history of the
sample during which oaks had not yet immigrated into the region.
We assumed the most probable estimates of h and M obtained
from the LAMARC analysis and translated them into Ne and m
using an estimate of l. The (short-term) pedigree rate of mutation,
which should not be confused with the (long-term) phylogenetic
substitution rate (Howell et al., 2003; Ho et al., 2005, 2007), has
been found to be approximately 1–2 108 for several organisms
(Drake et al., 1998; Nachman and Crowell, 2000; Denver et al.,
2004). However, Lutzoni and Pagel (1997) reported up to 10-fold
higher mutation rates in lichen-forming fungi compared to nonmutualistic relatives that could not be ascribed to significantly relaxed negative selection. Therefore, we settled for l = 107 per site
per generation. The following scenarios were simulated (subscripts
of m refer to the three phases described above): (A)
m1 = m2 = m3 = 0.001 (migration has remained constant and is
ongoing), (B) m1 = 0, m2 = m3 = 0.001 (migration ceased 13 generations ago in connection with a decrease in available oak habitat),
and (C) m1 = m2 = 0, m3 = 0.002 (migration ceased once the postglacial expansion of oak reached the region). In scenario C, we doubled the migration rate in phase 3 in order to maintain an
approximate average migration rate of 0.001 over the entire time
span (assuming that coalescence occurred during the bottleneck
caused by the latest glaciation). In all three scenarios, we simulated
a 20-fold increase in population size from phase 1 to phase 2 (corresponding to a 95% population reduction in forward time). Two
hundred data sets were simulated per scenario. The transition to
1874
H. Lättman et al. / Biological Conservation 142 (2009) 1870–1878
transversion rate used in LAMARC was maintained in SimCoal.
Each data set simulated by SimCoal was analyzed using Arlequin
3.1, and three types of summary statistics were collected in order
to compare them with observed values: (1) the proportion of within-population variation inferred by an AMOVA, (2) Tajima’s D (Tajima, 1989) averaged over populations, and (3) the average number
of polymorphic (segregating) sites per population. Ninety five percentage ranges were constructed by removing the five most extreme values at each end of the distributions.
3. Results
We found three IGS haplotypes, two of which were represented
by a single individual each and the third by 79 individuals. ITS
proved to be difficult to amplify and sequence. Eight samples,
which were successfully sequenced, displayed no variation at all.
We found the variation in IGS and ITS to be insufficient and discarded this data in subsequent statistical analyses. The SSU intron,
on the other hand, was represented by 11 haplotypes (Table 2). Out
of the 96 samples, 85 were successfully extracted, the SSU intron
amplified, sequenced, and consequently included in the statistical
analyses. The SSU intron length varied from 612 to 613 nucleotides, and the resulting alignment was 614 positions including
gaps. Ten positions were variable. A haplotype network is presented in Fig. 2. This shows that the 11 haplotypes are separated
by single mutational steps and that the two common haplotypes
(represented by 30 and 46 thalli, respectively) are internal and
hence presumably older than the infrequent terminal haplotypes.
The neutrality test indicated no deviation from neutral conditions
(all populations p = 1.00, Bjärka-Säby p = 0.79, Orräng p = 0.68, Solberga p = 0.84, Bråborg p = 0.38, and Stegeborg p = 0.40). The AMOVA indicated that 0.4% of the variance is between populations and
99.6% within populations. The reported fixation index (UST = 0.004)
was not significant (p = 0.35). Consequently, the AMOVA provided
no evidence of significant neutral differentiation among populations. The Mantel test revealed no indication of significant isolation-by-distance, the correlation between Slatkin’s linearized FST
and the logarithm of geographic distance being non-significant
(p = 0.70). Estimates of h, M, and Nem obtained via LAMARC, including 95% equal-tail credible intervals, are reported in Table 3. The
absolute number of successful establishments per generation per
population was estimated to be between 1 and 5, with the median
of the posterior probability being two. With four other sampled
populations, from which dispersal and establishment into a population can take place, the total number of successful establishments from the sampled populations is four times higher, i.e.
Fig. 2. Unrooted haplotype network for Cliostomum corrugatum with the two
common haplotypes 1 (n = 30) and 2 (n = 46) in the centre. Remaining terminal
haplotypes are represented by one thallus each. One mutational step between
haplotypes is represented by a line.
Table 3
Population parameter estimates obtained using the coalescent in a Bayesian MCMC
framework, as implemented in LAMARC 2.0.2. h = 2Nel, M = m/l, and Nem = hM/2,
where Ne is the effective population size, l the per site mutation rate per generation,
and m the per generation proportion of immigrants into a population from another
population. MPE = most probable estimate, corresponding to the mode of the
posterior probability distribution. 95% CI = 95% equal-tail credible interval.
Parameter
MPE
95% CI
h
M
Nem
6.1 104
9541
1.9
3.2 104–1.2 103
4225–9988
1.1–4.8
between 4 and 20 with the median at eight successful establishments per generation from the other four populations. Dispersal
Table 2
Variable nucleotide sites in the alignment of 11 haplotypes of the position 1516 SSU intron. The alignment was 614 sites in length. GenBank accession numbers for each haplotype
are indicated. Dots denote nucleotides that are identical to haplotype 1.
Haplotype
1
2
3
4
5
6
7
8
9
10
11
Total
Sequence and base number
Origin
0
2
6
0
7
9
1
0
3
1
4
1
1
5
1
2
6
4
2
9
5
4
7
1
5
1
6
5
2
3
Bjärka-Säby
Orräng
Solberga
Bråborg
Stegeborg
GenBank accession nos
G
A
C
C
C
A
C
A
7
7
4
13
G
T
T
8
7
1
1
1
5
10
G
A
T
T
–
–
C
–
–
–
–
–
C
–
–
6
9
–
–
–
–
–
–
–
C
–
–
–
A
1
15
EU218541
EU218542
EU218543
EU218544
EU218545
EU218546
EU218547
EU218548
EU218549
EU218550
EU218551
17
G
G
C
G
C
C
1
1
1
1
1
18
17
18
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H. Lättman et al. / Biological Conservation 142 (2009) 1870–1878
from unsampled populations, the number of which is unknown but
presumably rather large, comes in addition. Migration (M) may
have been underestimated, or at least truncated, because much
of the posterior probability accumulated right below the highest
upper limit allowed by LAMARC for that parameter. Table 4 accounts for the simulation of three different demographic scenarios.
This simulation shows that the observed values of the proportion
of within-population variation, the average Tajima’s D across populations, and the average number of polymorphic sites across population is compatible with both a model of ongoing migration as
well as a model where migration ceased 13 generations ago. The
observed values are, however, incompatible with a model where
migration ceased 200 generations ago, because of the higher expected proportion of within-population variation.
4. Discussion
4.1. Dispersal in C. corrugatum
Knowledge about effective dispersal rates and dispersal distances are paramount to any scientifically-based conservation
measure. Yet, such knowledge is unavailable for most organisms.
We used DNA sequence data from an intron near the terminal
end of the nuclear small subunit ribosomal DNA to infer effective
dispersal rates between populations of C. corrugatum. Other markers, ITS and IGS, failed to produce useful amounts of variation. The
limited amount of data available to us, a single gene, made it
imperative to apply analytical methods that reveal the uncertainty
in parameter estimates (in addition to methods that calculate point
estimates). We handled uncertainty in our estimates by use of a
Bayesian as well as a simulation approach.
Indirect estimates of dispersal rates and dispersal distances
through a point estimate of population differentiation (AMOVA),
a point estimate of the correlation between interpopulational genetic and physical distances (Mantel test), as well as a Bayesian direct measure of dispersal (Table 3) all indicate that effective
dispersal between populations at this spatial scale has been substantial and without measurable restrictions. The Bayesian approach indicates that the most likely number of successful
establishments per generation between the sampled populations
is between 4 and 20. The number of immigrants needed to prevent
neutral divergence of populations has been suggested to be ca 5
(Lacy, 1987), 1–10 (Mills and Allendorf, 1996), or more than 10
(Vucetich and Waite, 2000) per generation. The very wide Bayesian
posterior distribution of dispersal rate also demonstrates that our
estimate is indeed associated with considerable uncertainty. However, although it remains unknown exactly how high the rate of
dispersal is at this spatial scale, the credible interval clearly excludes low dispersal rates even though we used a uniform prior
distribution ranging from no dispersal at all to very high dispersal
rates. A possible explanation for the high rates of successful dispersal is that dispersal is not as passive as one might think. Perhaps
dispersal in C. corrugatum is facilitated by winged insects carrying
ascospores, conidia, or pieces of lichen thallus. Ascospores and al-
gal cells have been shown to be viable after having passed through
the gut of mites (Meier et al., 2002). Mites, in turn, could be carried
over large distances with the help of mammals or birds. A number
of insects have been shown to be faithful to the kind of oak trees
that C. corrugatum inhabits (Niklasson and Nilsson, 2005).
An assumption of the Bayesian coalescent analysis was that dispersal rates have remained approximately constant from the present to coalescence in backward time, otherwise inferred rates will
reduce to averages over time. The simulation study, although a
simplistic picture of the real events, efficiently demonstrates that
we cannot separate between a model with constant and ongoing
dispersal from a model where migration ceased at the time when
oak habitat started to decrease dramatically around four centuries
ago (Table 4, scenario A and B). In other words, we cannot know
whether effective dispersal is ongoing or whether recent fragmentation, owing to human influence on landscape characteristics, has
caused connectivity between populations to decrease. On the other
hand, a scenario where effective dispersal ceased already at the
time when the postglacial reinvasion of oak had reached the region
is implausible (Table 4, scenario C).
Could the high inferred dispersal rates be a consequence of high
connectivity in large and effectively continuous populations of C.
corrugatum from the time of oak reinvasion until around 1600 AD?
Current knowledge of the vegetation history of southern Sweden
allows some inferences about the dispersal capabilities of C.
corrugatum, although we can say nothing about ongoing dispersal.
In the historic agricultural landscape of Sweden, we know that from
1558 until 1830 oaks were considered state property to meet the
needs of timber for the navy. This royal decree was increasingly
being disregarded by peasants, and the 1825 reinventory of oaks
made some 30 years earlier disclosed an 80% reduction of timber
oaks during this short period of time (Eliasson and Nilsson, 2002).
Another inventory of oaks in Östergötland in 1813 demonstrated
that more than 80% of the oaks were found in the enclosed meadows
and fields surrounding the villages (Eliasson and Nilsson, 2002). The
remainder of the oaks were found outside the village enclosures,
grazing intensity (and thereby the amount of sun-lit oaks) progressively decreasing with increasing distance from the villages. Before
and after the 1558–1830 period, oaks were probably uncommon inside village enclosures. The Swedish system of a clear division between areas inside and outside village enclosures has a tradition
that goes back at least 1000 years but probably as much as
2000 years (Ekstam et al., 1988; Niklasson and Nilsson, 2005). During this period, C. corrugatum habitat was probably patchily distributed, oaks occurring under semi-open and sun-lit conditions almost
exclusively being found in or near villages. Further back in time, prior
to the advent of agricultural landscape, human influence was primarily by cultivation of temporary clearings in the forest (Niklasson
and Nilsson, 2005). Recent developments in paleoecology (Mitchell,
2005; Birks, 2005) indicate that in the early and mid-Holocene, much
of lowland Europe was covered by closed-canopy forests, contrary to
earlier suggestions involving wood-pastures kept open by megaherbivores (Vera, 2000). In closed-canopy forests, C. corrugatum would
have been restricted to steep, south- or west-facing slopes (Ek
Table 4
Median and 95% ranges of the proportion of within-population variation, average value of Tajima’s D per population, and the average number of polymorphic sites per population
obtained when simulating three different demographic scenarios: A (constant and ongoing migration), B (migration ceased around the time when the oak habitat started to
decrease dramatically around four centuries ago), and C (migration ceased once the postglacial reinvasion of oak had reached the region). Details of the simulation parameters are
found in the text. The observed parameter values are included for comparison.
Observed
Scenario A
Scenario B
Scenario C
Proportion of within-population variation
Tajima’s D (average per population)
No. of polymorphic sites (average per population)
0.004
0.086 (0.010–0.230)
0.086 (0.001–0.274)
0.144 (0.014–0.307)
0.312
0.358 (0.690–1.717)
0.363 (0.692–1.441)
0.410 (0.615–1.306)
3.6
4.6 (1.0–12.2)
4.7 (1.4–12.2)
4.0 (1.0–12.6)
1876
H. Lättman et al. / Biological Conservation 142 (2009) 1870–1878
et al., 1995) and lake and river edges. This type of habitat is likely to
have been highly patchily distributed in the landscape. In conclusion, dispersal between patches suitable for C. corrugatum during
the last 6000 years in Östergötland must commonly have involved
crossing distances of at least several kilometres, even if oaks were
notably more common than in the present-day landscape of southern Sweden. Indeed, inferred dispersal rates are high enough to suspect that the postglacial occurrence of C. corrugatum was primarily
limited by the availability of habitat and not by dispersal. Limitation
by habitat availability has been suggested also for lichens in stands of
aspen (Populus tremula), inferred from a combination of occupancy
patterns, successional history, and stand characteristics (Hedenås
and Ericson, 2004).
4.2. Dispersal in lichens – the current state of knowledge
There is considerable disagreement in the literature concerning
the ability of lichens to disperse and establish. Like in most other
organisms, effective dispersal rates seem to be scale-dependent,
but this explains only part of the disagreement. Vegetative diaspore
dispersal at short distances, up to a few hundred meters, has been
suggested to be effective, although several studies did not investigate the success rate of establishment (Armstrong, 1987, 1990; Tapper, 1976; Heinken, 1999; Lorentsson and Mattsson, 1999).
Dispersal limitation has been reported for tree-living lichens within
tree stands, between tree stands in close proximity, or up to a few
kilometres apart (Dettki et al., 2000; Sillett et al., 2000; Hilmo and
Såstad, 2001; Johansson and Ehrlen, 2003; Walser, 2004; Öckinger
et al., 2005), as well as between populations of an asexual terrestrial
lichen at a distance of up to a few kilometres (Cassie and PierceyNormore, 2008). However, genetic studies of Xanthoria parietina
and Lobaria pulmonaria suggest otherwise: effective dispersal at this
scale shows no sign of being restricted, although ascospores have
been found to disperse, on average, at longer distances than heavier
vegetative diaspores (Lindblom and Ekman, 2006, 2007; Wagner
et al., 2006; Werth et al., 2006a,b). At large spatial scales, populations
being separated by hundreds of kilometres or more, genetic studies
of lichen populations revealed severe dispersal restrictions (Printzen
et al., 2003; Palice and Printzen, 2004; Walser et al., 2005), whereas
studies relying on biogeographic patterns (Munoz et al., 2004), trapping of lichen fragments in the atmosphere (Harmata and Olech,
1991), or observations of lichen fragments on bird feet (Coppins
and James, 1979) implicitly proposed effective dispersal to be frequent. Finally, the small size and weight of the ascospores has been
taken as indirect evidence in favour of lichens being able to disperse
‘‘widely” (Nordén and Appelqvist, 2001).
What is the conservation message contained in our results? We
have inferred high rates of dispersal at landscape level in the history of a set of populations of a red-listed crustose lichen confined
to EC habitats. The often-repeated claim that lichens confined to EC
habitats are poor dispersers at more than very local scales may be a
severe underestimate of their capabilities. As mentioned above,
there are indications that some rare taxa restricted to EC forests
are indeed poor dispersers at the landscape level, but that conclusion might not apply universally. Furthermore, there is a non-negligible risk that the species so far investigated are not
representative among the lichens, the majority of EC species being
crustose like C. corrugatum. Unfortunately, life-history traits might
not help us to accurately predict dispersal ability (Johansson and
Ehrlen, 2003; Duminil et al., 2007). Our knowledge of the dispersal
capabilities of lichenized fungi therefore remains in its infancy.
4.3. Methodological issues
There are two methodological issues that need to be discussed
briefly. Firstly, lateral transfer of group I introns between positions
(Bhattacharya et al., 2002) and even interspecific horizontal transfer
(Martin et al., 2003; Simon et al., 2005) in the nuclear SSU rRNA gene
has been implicated in a phylogenetic perspective. However our
haplotype network (Fig. 2), which is typically star-shaped and separates the 11 haplotypes by single mutational steps, strongly indicates that horizontal transfer did not affect our study. Secondly, we
primarily used fruiting bodies (apothecia) for DNA extraction, because extractions from the vegetative thallus were more likely to
be troubled with contamination by other lichenized or non-lichenized fungi present in the habitat. Apothecia contain very small
amounts of dikaryotic tissue as well as meiotic ascospores that could
potentially contain genetic material from another, presumably nearby, individual, the ‘father’. However, we did not experience problems
with multiple mixed PCR products as evidenced by chromatograms
with superimposed base calls. We cannot say whether this means
that C. corrugatum is homothallic (the haploid equivalent of self-fertilizing) or just that the amount of ‘father DNA’ was too small to be
detected among the dominant ‘mother DNA’ in the vegetative hyphae making up the vast majority of the apothecial tissue.
Acknowledgements
We thank Wala and Folke Danielssons stiftelse for financial support, Nicklas Jansson at the County Administrative Board in Linköping for suggesting potential sample sites for C. corrugatum, and
members of the plant ecology group at Linköping University for
comments on the manuscript and David Lawrence for last spelling
corrections.
References
Armstrong, R.A., 1987. Dispersal in a population of the lichen Hypogymnia physodes.
Environmental and Experimental Botany 27, 357–363.
Armstrong, R.A., 1990. Dispersal, establishment and survival of soredia and
fragments of the lichen, Hypogymnia physodes (L.) Nyl. New Phytologist 114,
239–245.
Arup, U., 1997. Skoglig kontinuitet. In: Arup, U., Ekman, S., Kärnefelt, I., Mattsson, J.E. (Eds.), Skyddsvärda lavar i sydvästra Sverige, SBF-förlaget, Lund, pp. 92–95.
Beerli, P., 2004. Effect of unsampled populations on the estimation of population
sizes and migration rates between sampled populations. Molecular Ecology 13,
827–836.
Beerli, P., 2006. Comparison of Bayesian and maximum-likelihood inference of
population genetic parameters. Bioinformatics 22, 341–345.
Berg, A., Ehnström, B., Gustafsson, L., Hallingbäck, T., Jonsell, M., Weslien, J., 1994.
Threatened plant, animal, and fungus species in Swedish forests – distribution
and habitat associations. Conservation Biology 8, 718–731.
Berg, A., Gärdenfors, U., Hallingbäck, T., Noren, M., 2002. Habitat preferences of redlisted fungi and bryophytes in woodland key habitats in southern Sweden –
analyses of data from a national survey. Biodiversity and Conservation 11,
1479–1503.
Bhattacharya, D., Friedl, T., Helms, G., 2002. Vertical evolution and intragenic spread
of lichen-fungal group I introns. Journal of Molecular Evolution 55, 74–84.
Birks, H.J.B., 2005. Mind the gap: how open were European primeval forests? Trends
in Ecology and Evolution 20, 154–156.
Brewer, S., Cheddadi, R., de Beaulieu, J.L., Reille, M., 2002. The spread of deciduous
Quercus throughout Europe since the last glacial period. Forest Ecology and
Management 156, 27–48.
Cameron, R.P., 2002. Habitat associations of epiphytic lichens in managed and
unmanaged forest stands in Nova Scotia. Northeastern Naturalist 9, 27–46.
Cassie, D.M., Piercey-Normore, M.D., 2008. Dispersal in a sterile lichen-forming
fungus, Thamnolia subuliformis (Ascomycotina: Icmadophilaceae). Botany 86,
751–762.
Clement, M., Posada, D., Crandall, K.A., 2000. TCS: a computer program to estimate
gene genealogies. Molecular Ecology 9, 1657–1659.
Coppins, B.J., James, P.W., 1979. Birds and the dispersal of lichen propagules.
Lichenologist 11, 105–106.
Denver, D.R., Morris, K., Lynch, M., Thomas, W.K., 2004. High mutation rate and
predominance of insertions in the Caenorhabditis elegans nuclear genome.
Nature 430, 679–682.
Dettki, H., Klintberg, P., Esseen, P.A., 2000. Are epiphytic lichens in young forests
limited by local dispersal? Ecoscience 7, 317–325.
Drake, J.W., Charlesworth, B., Charlesworth, D., Crow, J.F., 1998. Rates of
spontaneous mutation. Genetics 148, 1667–1686.
Duminil, J., Fineschi, S., Hampe, A., Jordano, P., Salvini, D., Vendramin, G.G., Petit, R.J.,
2007. Can population genetic structure be predicted from life-history traits?
American Naturalist 169, 662–672.
H. Lättman et al. / Biological Conservation 142 (2009) 1870–1878
Ek, T., Wadstein, M., Johannesson, J., 1995. Varifrån kommer lavar knutna till gamla
ekar? Svensk Botanisk Tidskrift 89, 335–343.
Ekstam, U., Aronsson, M., Forshed, N., 1988. Ängar. LTs förlag, Stockholm.
Eliasson, P., Nilsson, S.G., 2002. ‘You should hate young oaks and young noblemen’.
The environmental history of oaks in eighteenth- and nineteenth-century
Sweden. Environmental History 7, 659–677.
Excoffier, L., Smouse, P.E., Quattro, J.M., 1992. Analysis of molecular variance
inferred from metric distances among DNA haplotypes – application to human
mitochondrial-DNA restriction data. Genetics 131, 479–491.
Fenton, N.J., Bergeron, Y., 2008. Does time or habitat make old-growth forests
species rich? Bryophyte richness in boreal Picea mariana forests. Biological
Conservation 141, 1389–1399.
Fritz, Ö., Gustafsson, L., Larsson, K., 2008. Does forest continuity matter in
conservation? – A study of epiphytic lichens and bryophytes in beech forests
of southern Sweden. Biological Conservation 141, 655–668.
Fritz, Ö., Larsson, K., 1996. Betydelsen av skoglig kontinuitet för rödlistade lavar. En
studie av halländsk bokskog. Svensk Botanisk Tidskrift 90, 241–262.
Gärdenfors, U. (Ed.), 2005. Rödlistade arter i Sverige 2005. ArtDatabanken, SLU,
Uppsala.
Gardes, M., Bruns, T.D., 1993. ITS primers with enhanced specificity for
basidiomycetes – application to the identification of mycorrhizae and rusts.
Molecular Ecology 2, 113–118.
Gargas, A., Depriest, P.T., Grube, M., Tehler, A., 1995. Multiple origins of lichen
symbioses in fungi suggested by SSU rDNA phylogeny. Science 268, 1492–1495.
Gauslaa, Y., Ohlson, M., 1997. Et historisk perspektiv på kontinuitet og forekomst av
epifyttiske laver i norske skoger. Blyttia 55, 15–27.
Goward, T., 1994. Notes on oldgrowth-dependent epiphytic macrolichens in inland
British Columbia, Canada. Acta Botanica Fennica 150, 31–38.
Gustafsson, L., Hallingbäck, T., 1988. Bryophyte flora and vegetation of managed and
virgin coniferous forests in south-west Sweden. Biological Conservation 44,
283–300.
Hall, T.A., 1999. BioEdit: a user-friendly biological sequence alignment editor and
analysis program for Windows 95/98/NT. Nucleic Acids Symposium Series 41,
95–98.
Harmata, K., Olech, M., 1991. Transect for aerobiological studies from Antarctica to
Poland. Grana 30, 458–463.
Hedenås, H., Ericson, L., 2000. Epiphytic macrolichens as conservation indicators:
successional sequence in Populus tremula stands. Biological Conservation 93,
43–53.
Hedenås, H., Ericson, L., 2004. Aspen lichens in agricultural and forest landscapes:
the importance of habitat quality. Ecography 27, 521–531.
Heinken, T., 1999. Dispersal patterns of terricolous lichens by thallus fragments.
Lichenologist 31, 603–612.
Hilmo, O., Såstad, S.M., 2001. Colonization of old-forest lichens in a young and an
old boreal Picea abies forest: an experimental approach. Biological Conservation
102, 251–259.
Ho, S.Y.W., Phillips, M.J., Cooper, A., Drummond, A.J., 2005. Time dependency of
molecular rate estimates and systematic overestimation of recent divergence
times. Molecular Biology and Evolution 22, 1561–1568.
Ho, S.Y.W., Shapiro, B., Phillips, M., Cooper, A., Drummond, A.J., 2007. Evidence for
time dependency of molecular rate estimates. Systematic Biology 56, 515–522.
Howell, N., Smejkal, C.B., Mackey, D.A., Chinnery, P.F., Turnbull, D.M., Herrnstadt, C.,
2003. The pedigree rate of sequence divergence in the human mitochondrial
genome: there is a difference between phylogenetic and pedigree rates.
American Journal of Human Genetics 72, 659–670.
Johansson, P., Ehrlen, J., 2003. Influence of habitat quantity, quality and isolation on
the distribution and abundance of two epiphytic lichens. Journal of Ecology 91,
213–221.
Johansson, V., Bergman, K.-O., Lättman, H., Milberg, P., in press. Tree and site quality
preferences of six epiphytic lichens growing on oaks in south eastern Sweden.
Annales Botanici Fennici.
Johnson, P.L.F., Slatkin, M., 2007. Accounting for bias from sequencing error in
population genetic estimates. Molecular Biology and Evolution 25, 199–206.
Kålås, J.A., Viken, Å., Bakken, T. (Eds.), 2006. Norsk rødliste 2006. Artsdatabanken,
Trondheim.
Kalwij, J.M., Wagner, H.H., Scheidegger, C., 2005. Effects of stand-level disturbances
on the spatial distribution of a lichen indicator. Ecological Applications 15,
2015–2024.
Kass, R.E., Raftery, A.E., 1995. Bayes factors. Journal of the American Statistical
Association 90, 773–795.
Kuhner, M.K., 2003. LAMARC: estimating population genetic parameters from
molecular data. In: Salemi, M., Vandamme, A. (Eds.), The Phylogenetic
Handbook: A Practical Approach to DNA and Protein Phylogeny. Cambridge
University Press, Cambridge, pp. 379–399.
Kuhner, M.K., Yamato, J., Beerli, P., Smith, L.P., Rynes, E., Walkup, E., Li, C., Sloan, J.,
Colacurcio, P., Felsenstein, J., 2005. LAMARC v 2.0. University of Washington.
<http://evolution.gs.washington.edu/lamarc.html> (Cited 7 September 2006).
Kuhner, M.K., 2006. LAMARC 2.0: maximum likelihood and Bayesian estimation of
population parameters. Bioinformatics 22, 768–770.
Kuhner, M.K., Smith, L.P., 2007. Comparing likelihood and Bayesian coalescent
estimation of population parameters. Genetics 175, 155–165.
Kuusinen, M., 1996. Cyanobacterial macrolichens on Populus tremula as indicators
of forest continuity in Finland. Biological Conservation 75, 43–49.
Lacy, R.C., 1987. Loss of genetic diversity from managed populations: interacting
effects of drift, mutation, immigration, selection, and population subdivision.
Conservation Biology 1, 143–158.
1877
Laval, G., Excoffier, L., 2004. SIMCOAL 2.0: a program to simulate genomic diversity
over large recombining regions in a subdivided population with a complex
history. Bioinformatics 20, 2485–2487.
Lindblad, I., 1998. Wood-inhabiting fungi on fallen logs of Norway spruce: relations
to forest management and substrate quality. Nordic Journal of Botany 18, 243–
255.
Lindblom, L., Ekman, S., 2006. Genetic variation and population differentiation in
the lichen-forming ascomycete Xanthoria parietina on the island Storfosna,
central Norway. Molecular Ecology 15, 1545–1559.
Lindblom, L., Ekman, S., 2007. New evidence corroborates population differentiation
in Xanthoria parietina. Lichenologist 39, 259–271.
Lõhmus, A., Lõhmus, P., 2008. First-generation forests are not necessary worse than
long-term managed forests for lichens and bryophytes. Restoration Ecology 16,
231–239.
Lorentsson, S., Mattsson, J.-E., 1999. New reports of soredia dispersed by ants,
Formica cunicularia. Lichenologist 31, 204–207.
Ludwig, G., Schnittler, M. (Eds.), 1996. Rote Liste gefährdeter pflanzen Deutschlands.
Schriftenreihe für vegetationskunde 28, 1–744.
Lundqvist, J., 1998. Isavsmältning och israndlinjer i Sverige. In: Andersen, S.,
Pedersen, S.S. (Eds.), Israndslinier i Norden. TemaNord, København, pp. 61–82.
Lutzoni, F., Pagel, M., 1997. Accelerated evolution as a consequence of transitions to
mutualism. Proceedings of the National Academy of Sciences of the USA 94,
11422–11427.
Martikainen, P., Siitonen, J., Kaila, L., Punttila, P., Rauh, J., 1999. Bark beetles
(Coleoptera, Scolytidae) and associated beetle species in mature managed and
old-growth boreal forests in southern Finland. Forest Ecology and Management
116, 233–245.
Martikainen, P., Siitonen, J., Punttila, P., Kaila, L., Rauh, J., 2000. Species richness of
Coleoptera in mature managed and old-growth boreal forests in southern
Finland. Biological Conservation 94, 199–209.
Martin, M.P., Coucheron, D.H., Johansen, S., 2003. Structural features and
evolutionary considerations of group IB introns in SSU rDNA of the lichen
fungus Teloschistes. Fungal Genetics and Biology 40, 252–260.
Meier, F.A., Scherrer, S., Honegger, R., 2002. Faecal pellets of lichenivorous mites
contain viable cells of the lichen-forming ascomycete Xanthoria parietina and its
green algal photobiont, Trebouxia arboricola. Biological Journal of the Linnean
Society 76, 259–268.
Mills, L.S., Allendorf, F.W., 1996. The one-migrant-per-generation rule in
conservation and management. Conservation Biology 10, 1509–1518.
Mitchell, F.J.G., 2005. How open were European primeval forests? Hypothesis
testing using palaeoecological data. Journal of Ecology 93, 168–177.
Munoz, J., Felicisimo, A.M., Cabezas, F., Burgaz, A.R., Martinez, I., 2004. Wind as a
long-distance dispersal vehicle in the Southern Hemisphere. Science 304, 1144–
1147.
Nachman, M.W., Crowell, S.L., 2000. Estimate of the mutation rate per nucleotide in
humans. Genetics 156, 297–304.
Newton, M.A., Raftery, A.E., Davison, A.C., Bacha, M., Celeux, G., Carlin, B.P., Clifford,
P., Lu, C., Sherman, M., Tanner, M.A., Gelfand, A.E., Mallick, B.K., Gelman, A.,
Grieve, A.P., Kunsch, H.R., Leonard, T., Hsu, J.S.J., Liu, J.S., Rubin, D.B., Lo, A.Y.,
Louis, T.A., Neal, R.M., Owen, A.B., Tu, D.S., Gilks, W.R., Roberts, G., Sweeting, T.,
Bates, D., Ritter, G., Worton, B.J., Barnard, G.A., Gibbens, R., Silverman, B., 1994.
Approximate Bayesian-inference with the weighted likelihood bootstrap.
Journal of the Royal Statistical Society Series B – Methodological 56, 3–48.
Niklasson, M., Nilsson, S.G., 2005. Skogsdynamik och arters bevarande.
Studentlitteratur, Lund.
Nilsson, S.G., Baranowski, R., 1997. Habitat predictability and the occurrence of
wood beetles in old-growth beech forests. Ecography 20, 491–498.
Nordén, B., Appelqvist, T., 2001. Conceptual problems of ecological continuity and
its bioindicators. Biodiversity and Conservation 10, 779–791.
Öckinger, E., Niklasson, M., Nilsson, S.G., 2005. Is local distribution of the epiphytic
lichen Lobaria pulmonaria limited by dispersal capacity or habitat quality?
Biodiversity and Conservation 14, 759–773.
Ohlson, M., Söderström, L., Hörnberg, G., Zackrisson, O., Hermansson, J., 1997.
Habitat qualities versus long-term continuity as determinants of biodiversity in
boreal old-growth swamp forests. Biological Conservation 81, 221–231.
Økland, B., 1996. Unlogged forests: important sites for preserving the diversity of
mycetophilids (Diptera: Sciaroidea). Biological Conservation 76, 297–310.
Palice, Z., Printzen, C., 2004. Genetic variability in tropical and temperate
populations of Trapeliopsis glaucolepidea: evidence against long-range
dispersal in a lichen with disjunct distribution. Mycotaxon 90, 43–54.
Penttilä, R., Siitonen, J., Kuusinen, M., 2004. Polypore diversity in managed and oldgrowth boreal Picea abies forests in southern Finland. Biological Conservation
117, 271–283.
Posada, D., Crandall, K.A., 1998. MODELTEST: testing the model of DNA substitution.
Bioinformatics 14, 817–818.
Printzen, C., Ekman, S., Tønsberg, T., 2003. Phylogeography of Cavernularia hultenii:
evidence of slow genetic drift in a widely disjunct lichen. Molecular Ecology 12,
1473–1486.
Rassi, P., Alanen, A., Kanerva, T., Mannerkoski, I. (Eds.), 2001. The Red List of Finnish
Species. Ministry of the Environment, Finnish Environment Institute, Helsinki.
Rivas Plata, E., Lücking, R., Lumbsch, H.T., 2008. When family matters: an analysis of
Thelotremataceae (lichenized Ascomycota: Ostropales) as bioindicators of
ecological continuity in tropical forests. Biodiversity and Conservation 17,
1319–1351.
Rolstad, J., Gjerde, I., Gundersen, V.S., Stersdal, M., 2002. Use of indicator species to
assess forest continuity: a critique. Conservation Biology 16, 253–257.
1878
H. Lättman et al. / Biological Conservation 142 (2009) 1870–1878
Rose, F., 1974. The epiphytes of oak. In: Morris, M.G., Perring, F.H. (Eds.), The British
Oak: Its History and Natural History. Classey, Faringdon, pp. 250–273.
Rousset, F., 1997. Genetic differentiation and estimation of gene flow from Fstatistics under isolation by distance. Genetics 145, 1219–1228.
Schneider, S., Roessli, D., Excoffier, L., 2000. Arlequin: A Software for Population
Genetics Data Analysis. Version 2.000. Genetics and Biometry Laboratory,
Department of Anthropology, University of Geneva.
Selva, S.B., 1994. Lichen diversity and stand continuity in the northern hardwoods
and spruce-fir forests of northern New England and western New Brunswick.
Bryologist 97, 424–429.
Selva, S.B., 2003. Using calicioid lichens and fungi to assess ecological continuity in
the Acadian Forest Ecoregion of the Canadian Maritimes. Forestry Chronicle 79,
550–558.
Sillett, S.C., McCune, B., Peck, J.E., Rambo, T.R., Ruchty, A., 2000. Dispersal limitations
of epiphytic lichens result in species dependent on old-growth forests.
Ecological Applications 10, 789–799.
Simon, D., Moline, J., Helms, G., Friedl, T., Bhattacharya, D., 2005. Divergent histories
of rDNA group I introns in the lichen family Physciaceae. Journal of Molecular
Evolution 60, 434–446.
Slatkin, M., 1994. An exact test for neutrality based on the Ewens sampling
distribution. Genetical Research 64, 71–74.
Slatkin, M., 1996. A correction to the exact test based on the Ewens sampling
distribution. Genetical Research 68, 259–260.
Slatkin, M., 2005. Seeing ghosts: the effect of unsampled populations on migration
rates estimated for sampled populations. Molecular Ecology 14, 67–73.
Spence, J.R., Langor, D.W., Niemela, J., Carcamo, H.A., Currie, C.R., 1996. Northern
forestry and carabids: the case for concern about old-growth species. Annales
Zoologici Fennici 33, 173–184.
Stewart, F.M., 1977. Computer algorithm for obtaining a random set of allele
frequencies for a locus in an equilibrium population. Genetics 86, 482–483.
Stoltze, M., Pihl, S. (Eds.), 1998. Rødliste 1997 over planter og dyr i Danmark. Miljøog energiministeriet, Danmarks miljøundersøgelser, Skov- og naturstyrelsen,
København.
Sverdrup-Thygeson, A., Lindenmayer, D.B., 2003. Ecological continuity and assumed
indicator fungi in boreal forest: the importance of the landscape matrix. Forest
Ecology and Management 174, 353–363.
Swofford, D.L., 2003. PAUP. Phylogenetic Analysis Using Parsimony ( and other
Methods). Version 4. Sinauer Associates, Sunderland, Massachusetts.
Tajima, F., 1989. Statistical method for testing the neutral mutation hypothesis by
DNA polymorphism. Genetics 123, 585–595.
Tapper, R., 1976. Dispersal and changes in the local distributions of Evernia prunastri
and Ramalina farinacea. New Phytologist 77, 725–734.
Thor, G., Arvidsson, L. (Eds.), 1999. Rödlistade lavar i Sverige – artfakta.
ArtDatabanken, SLU, Uppsala.
Tibell, L., 1992. Crustose lichens as indicators of forest continuity in boreal
coniferous forests. Nordic Journal of Botany 12, 427–450.
Tikkanen, O.P., Martikainen, P., Hyvärinen, E., Junninen, K., Kouki, J., 2006. Red-listed
boreal forest species of Finland: associations with forest structure, tree species,
and decaying wood. Annales Zoologici Fennici 43, 373–383.
Uliczka, H., Angelstam, P., 1999. Occurrence of epiphytic macrolichens in relation to
tree species and age in managed boreal forest. Ecography 22, 396–405.
Vera, F.W.M., 2000. Grazing Ecology and Forest History. CABI International,
Wallingford, UK.
Vucetich, J.A., Waite, T.A., 2000. Is one migrant per generation sufficient for the
genetic management of fluctuating populations? Animal Conservation 3, 261–
266.
Wagner, H.H., Werth, S., Kalwij, J.M., Bolli, J.C., Scheidegger, C., 2006. Modelling
forest recolonization by an epiphytic lichen using a landscape genetic approach.
Landscape Ecology 21, 849–865.
Walser, J.C., 2004. Molecular evidence for limited dispersal of vegetative propagules
in the epiphytic lichen Lobaria pulmonaria. American Journal of Botany 91,
1273–1276.
Walser, J.C., Holderegger, R., Gugerli, F., Hoebee, S.E., Scheidegger, C., 2005.
Microsatellites reveal regional population differentiation and isolation in
Lobaria pulmonaria, an epiphytic lichen. Molecular Ecology 14, 457–467.
Werth, S., Wagner, H.H., Gugerli, F., Holderegger, R., Csencsics, D., Kalwij, J.M.,
Scheidegger, C., 2006a. Quantifying dispersal and establishment limitation in a
population of an epiphytic lichen. Ecology 87, 2037–2046.
Werth, S., Wagner, H.H., Holderegger, R., Kalwij, J.M., Scheidegger, C., 2006b. Effect
of disturbances on the genetic diversity of an old-forest associated lichen.
Molecular Ecology 15, 911–921.
White, T.J., Bruns, T., Lee, S., Taylor, J., 1990. Amplification and direct sequencing of
fungal ribosomal RNA genes for phylogenetics. In: Innis, M.A., Gelfand, D.H.,
Sninsky, J.J., White, T.J. (Eds.), PCR Protocols: A Guide to Methods and
Applications. Academic Press, San Diego, California, pp. 315–322.
Woods, R.G., Coppins, B.J., 2003. A Conservation Evaluation of British Lichens.
British Lichen Society, London.
Wright, S., 1943. Isolation by distance. Genetics 28, 114–138.
Zenner, E.K., 2004. Does old-growth condition imply high live-tree structural
complexity? Forest Ecology and Management 195, 243–258.
PAPER III
Nordic Journal of Botany 27: 413418, 2009
doi: 10.1111/j.1756-1051.2009.00425.x,
# The Authors. Journal compilation # Nordic Journal of Botany 2009
Subject Editor: Torbjörn Tyler. Accepted 18 June 2009
Changes in the distributions of epiphytic lichens in southern Sweden
using a new statistical method
Håkan Lättman, Per Milberg, Michael W. Palmer and Jan-Eric Mattsson
H. Lättman ([email protected]) and J.-E. Mattsson, School of Life Sciences, Södertörn Univ., SE181 89 Huddinge, Sweden. HL and
P. Milberg, IFM Biology, Division of Ecology, Linköping Univ., SE581 83 Linköping, Sweden. PM also at: Dept of Crop Production Ecology,
SLU, PO Box 7043, SE750 07 Uppsala, Sweden. M. W. Palmer, Dept of Botany, Oklahoma State Univ., Stillwater, Oklahoma 74078,
USA.
Past studies on changes in species distribution have mainly been based on analysis of range boundaries. In contrast, the
method used here evaluates shifts in species’ geographic centroids within a predefined area. We used presence/absence
data on epiphytic lichens collected 1986 and 2003 from 64 sites in southern Sweden. A centroid was calculated each year,
for each lichen species and substrate. The distance of centroid movement was evaluated in a permutation procedure. In
total, 56 lichen species on 22 tree species were involved in the analyses, yielding 30 cases that had sufficient sample sizes
both years to be evaluated. Out of these, three exhibited a significant movement of their centroid. The shift of lichen
centroids of Hypogymnia physodes (L.) Nyl. and Vulpicida pinastri (Scop.) J.-E. Mattsson & M. J. Lai on the tree species
Juniperus communis L. was 50 and 151 km with the direction 278 and 488, respectively. For Hypogymnia physodes on Pinus
sylvestris L., corresponding values were 41 km and 308. The northnortheast shifts of these species in Sweden could be a
response to a warming climate.
Large-scale environmental changes, such as global warming,
are likely to affect several organisms in a similar way, e.g. by
range shifts in similar directions, and also through a change
in the density of individuals within populations. So far,
most studies of distributional changes attributed to global
warming in the northern hemisphere have been examined as
the northward expansion of the studied organism (Thomas
and Lennon 1999, Warren et al. 2001, Parmesan and Yohe
2003, Root et al. 2003, Hickling et al. 2005). The number
of studies based on population densities within a species
distribution (Thomas and Lennon 1999, Warren et al.
2001, Hickling et al. 2005), or to the likelihood of encountering a species (Bridle and Vines 2007) is much lower.
As the appropriate spatial and temporal scales involved
are large, relatively few data sets exist. Further, available
data are often not straightforward and rarely detailed; for
example field notes from sites with long time gaps, data not
collected under a uniform protocol, or data with low
information content (e.g. presence/absence only). Data of
this type, however, still provide opportunities for interesting
questions to be addressed. Thus, simple methods for
analysis of uneven-quality data would be valuable.
Epiphytic lichens on trees, and other transient substrates,
are likely to respond relatively quickly to broad-scale
changes in the climate. van Herk et al. (2002) used
checklists, data from field meetings, herbarium material
and long-term monitoring data to study large-scale changes
of lichens in the Netherlands. Based on the documentation
of 329 lichen species collected in 1979, 1984, 1989, 1995
and 2001, they conclude that lichens can respond to global
warming over a time frame of only a few decades. Lichens
with a northern distribution and a southern border in the
Netherlands have declined. On the other hand, species with
a southern distribution and a northern border are at present
invading the Netherlands.
In the present study, we illustrate an approach using
presence/absence data to compare calculated centroids of
geographic location, within the area studied (also known as
‘centre of occurrence’, La Sorte and Thompson 2007), from
investigations at two points in time (1986 and 2003). The
changes of distance and direction of the centroid was easily
calculated and for distance, significance (null hypothesis of
no change in centroid) can be tested with a permutation
procedure. One advantage is that no assumption is needed
about the direction (most analyses so far test for changes in
northern and southern borders). Another is that our focus is
on the general distribution of the species within any
predefined area of any shape, not just on the boundaries.
The latter might be a poor representation of a large-scale
change in abundance, is sensitive to grid size, might be
highly influenced by single observations and be very
sensitive to data deficiencies (Dormann 2007). Furthermore, data for most species are not available to determine
clear distribution boundaries. We apply the method to a
data set on epiphytic lichens, a group likely to respond
relatively quickly, as both the lichen and its host might
413
respond to environmental changes on temporal and spatial
scales relevant for the data at hand. The data come from 64
sites in southern Sweden visited in 1986 and 2003.
Material and methods
Sampling
Epiphytic lichens on 64 sites in southern Sweden were
sampled in 1986 and 2003 (Fig. 1a). One of the authors
Jan-Eric Mattsson, (JEM) originally selected and visited the
sites for a study to which the extent Vulpicida juniperinus
(L.) J.-E. Mattsson and M. J. Lai and Vulpicida pinastri
(Scop.) J.-E. Mattsson and M. J. Lai were still present at
former localities (Mattsson 1988). The identification of
species and selection of the sites, from the first study, were
based on herbarium specimens of the genus Vulpicida
collected over a period of about one hundred years. Hence,
the sites are not a random selection but follow a previously
known occurrence of these target species. The sites span
over a wide range of ecological habitats with no apparent
bias towards a particular tree diameter (successional stage).
The sites were of different sizes ranging from a single tree to
one hectare and often delimited by natural boundaries such
as creeks or ridges, or by man-made borders such as fences,
or different land use. Different habitats common to
southern Sweden are represented at the sites, i.e. coniferous,
deciduous and mixed forests, wooded pastures, parks, and
on occasion single trees in towns, villages and farms.
Tree trunks, branches and twigs were examined to a
height of 2 m above the ground in search of all epiphytic
macrolichens. And on each site, all epiphytic lichen species
were collected on all substrates present. The number of tree
species per site varied from one to eleven and the aim was to
investigate approximately 50 trees with lichens at each site.
Tree trunks without visual lichens were observed, but were
not included in the study. Most of the lichen species
recorded are widespread and abundant with some exceptions, e.g. the redlisted Melanelia laciniatula (Flagey ex
H. Olivier) Essl. and Usnea barbata (L.) Weber ex F. H.
Wigg. (Thor and Arvidsson 1999). The species are, with a
few exceptions, easy to identify based on macro-characters
only. Most of the lichen species are foliose or fruticose.
Some of the species are habitat generalists and may occur on
several of the tree species surveyed, as well as on rock. One
observer (JEM) investigated the sites in 1986 and two
observers JEM and Håkan Lättman (HL) in 2003. Total
time, spent to record all species present, per site was similar
both times and the search method was identical. In 1986,
JEM used 40 minutes on each site and in 2003, JEM and
Figure 1. (a)(d) distribution of 64 sites in southern Sweden where collection of epiphytic lichen species were recorded 1986 and 2003,
and the results of the three significant cases of centroid movement (arrows) (Table 1). (a) the 64 investigated sites. The centroid shift of
(b) Hypogymnia physodes on Pinus sylvstris, (c) Hypogymnia physodes on Juniperus communis and (d) Vulpicida pinastri on Juniperus
communis. A cross indicates a site occupied by the lichen species 1986, a plus indicate a site occupied by the lichen species 2003 and the
star indicate a site occupied by the lichen species both years. The arrows indicate length and direction of centroid movements.
414
HL used 20 minutes each per site. Tree and lichen species
nomenclature follows Karlsson (1997) and Santesson et al.
(2004) respectively.
In total, the data included 64 sites, 22 tree species and
standing dead wood and 56 lichen species. Longitude and
latitude were recorded in field and converted to WGS 84
before analysis.
Analysis
We developed a permutation procedure to assess whether
the distance moved by the distributional centroid (the
arithmetic mean of coordinates of sites) was greater than
expected due to chance. Each combination of lichen and
tree species was analyzed separately. For a site to be
included for a particular combination of lichen and tree
species analysis, the tree species had to be recorded at the
site both years and the lichen species in at least one of the
years. From the information included per combination of
lichen and tree, we calculated the geographic centroid of the
lichen species in 1986 and 2003, and then the distance
between these centroids. We then randomly permutated the
status (present both times, present in 1986 only, and
present in 2003 only) for all those sites in which the lichen
occurred on that tree species at least once (Table 1). Thirty
combination of trees and lichen species were possible to
analyze, involving eight and 17 tree and lichens species,
respectively (Table 1).
We calculated the centroids and associated distance for
each of 5000 permutations, and compared these distances
to our measured value to obtain a p-value. This analysis was
performed using an Excel visual basic macro written by
MWP.
Although it may seem that a reasonable null hypothesis
for direction of movement is that all compass directions are
equally likely, irregularities in the distribution of samples
as well as the elongated shape of the study area in Sweden
(Fig. 1) mean that this is not the case. Therefore, we stored
the directions generated by the permuted data to compile an
expected distribution of directions. In addition, we calculated the directions between all possible pair-wise combinations of sites. The cumulative distribution of these two data
sets were compared with the corresponding of the observed
changes of directions of centroids (including those whose
distance had scored NS), using the KolmogorovSmirnov
(KS) test.
Results
Centroid movements of lichen species
We recorded a significant change (p B0.05) in distance
between the centroids calculated for 1986 and 2003 for
three of the 30 analysis possible to conduct (Table 1).
Centroid movements of the lichens Hypogymnia physodes
and Vulpicida pinastri on the tree species Juniperus communis
were 50 and 151 km, respectively, and corresponding
movement of Hypogymnia physodes on Pinus sylvestris was
41 km (Table 1). The direction of centroid movements of
these three significant cases were in a northeast direction
(Fig. 1). Furthermore, in these three cases there was a
general decrease in occurrence. Hypogymnia physodes and
Vulpicida pinastri on Juniperus communis decreased from 14
to 6 and from 14 to 3 observations, respectively. For
Hypogymnia physodes on Pinus sylvestris, the corresponding
values were 8 to 4 (Table 1).
If applying a control for false discovery rates (following
Benjamini and Hochberg 1995), due to multiple testing
(n 30), one of the three cases remained significant (i.e.
Vulpicida pinastri on Juniperus communis: p0.0002). It
can be argued, however, whether the 30 tests conducted
actually belong to the same family of tests (cf. Perneger
1998, Proschan and Waclawiw 2000).
The cumulative distribution of all possible directions
and of all permuted directions were similar, with two soft
bumps (Fig. 2); a consequence of the elongated shape of
the study area (Fig. 1a). The observed distribution of
the 30 analyses, that were possible to conduct (Fig. 2),
deviated from both of the above-mentioned in the KS tests
(pB0.025, pB0.005). Overall, the direction of movement
was, for the 30 evaluated cases, dominated by 16 in northeast
and 9 in southwest (Fig. 2).
Discussion
Movements and directions as judged by centroids
There are two main results in this study. First, it provides
evidence that movement of centroids (representing the
probability of finding a species, within an area), of epiphytic
lichens can be detected over a time frame of less than two
decades. The strength of the evidence is further discussed
below. This conclusion corroborates the findings of van
Herk et al. (2002); a study that was conducted on comparable spatial and temporal scales which suggested that
epiphytic lichens can respond very quickly to climatic
changes. It is easy to speculate on the cause for shifts in
distribution but more difficult to disentangle the possible
contributions of trends in, e.g. temperature, precipitation
and pollution. Yearly average temperature and precipitation
at sites within the investigated area increased slightly over the
study period (0.0568C year 1 (29 sites) and 5.0 mm year 1
(28 sites), respectively; SMHI 19872003). During the
same period air pollution decreased strongly and. e.g. NO2
(14 sites) and SO2 (6 sites) in the air dropped by
approximately 50 and 90%, respectively (IVL 2009). Other
studies of epiphytes in southern Sweden have documented a
local ‘reinvasion/recolonisation’ of lichens attributed to the
improved air quality (Hultengren et al. 2004). The three
significant cases documented in the present study involved
two lichens (Hypogymnia physodes and Vulpicida pinastri)
that both have mainly a northerly distribution in Sweden,
and that were shown to be generally on the retreat
(Table 1), and specifically so in the southwest (Fig. 1).
This region was also the one with the poorest air quality at
the onset of our study, so unless air pollutant concentrations
recorded in the 1980s turn out to be beneficial to the lichens
in question, it is difficult to see how the documented
patterns could be driven by a decrease in air pollution.
Hence, our tentative interpretation is that the shifts seen are
climate-driven.
415
416
Hypogymnia physodes L.
Hypogymnia physodes
Hypogymnia physodes
Hypogymnia physodes
Hypocenomyce scalaris (Ach.) M. Choisy
Pseudevernia furfuracea (L.) Zopf
Pseudevernia furfuracea
Platismatia glauca (L.) W. L. Culb. & C. F. Culb.
Parmeliopsis ambigua (Wulfen) Nyl.
Vulpicida pinastri (Scop.) J.-E. Mattsson & M. J. Lai
Parmelia sulcata Taylor
Vulpicida pinastri
Tuckermanopsis chlorophylla (Willd.) Hale
Hypogymnia tubulosa (Schaer.) Hav.
Usnea hirta (L.) Weber ex F. H. Wigg.
Vulpicida pinastri
Xanthoria parietina (L.) Th. Fr.
Ramalina farinacea (L.) Ach.
Evernia prunastri (L.) Ach.
Bryoria fuscescens (Gyeln.) Brodo & D. Hawskw.
Parmelia sulcata Taylor
Usnea hirta
Evernia prunastri
Usnea hirta
Usnea subfloridana Stirt.
Ramalina farinacea
Calicium viride Pers.
Ramalina fastgiata (Pers.) Ach.
Tuckermanopsis chlorophylla
Usnea subfloridana Stirt.
Lichen species
Betula spp.
Pinus sylvestris L.
Picea abies (L.) H. Karst
Juniperus communis L.
Pinus sylvestris
Betula spp.
Picea abies
Betula spp.
Betula spp.
Juniperus communis
Betula spp.
Pinus sylvestris
Picea abies
Juniperus communis
Pinus sylvestris
Picea abies
Populus tremula L.
Acer platanoides L.
Populus tremula
Betula spp.
Picea abies
Betula spp.
Picea abies
Juniperus communis
Betula spp.
Fraxinus excelsior L.
Quercus robur L.
Fraxinus excelsior
Pinus sylvestris
Picea abies
Tree species
9/35/15
8/31/4
9/29/3
14/23/6
6/10/15
16/6/12
9/9/13
10/7/9
5/3/19
14/5/3
7/1/14
8/3/2
6/1/7
2/1/10
8/2/2
13/0/1
1/2/8
0/4/4
5/1/4
6/2/0
1/2/5
7/1/1
2/1/5
5/2/0
8/0/1
1/0/6
1/1/2
1/0/4
0/1/3
3/0/2
Occurrences 1986 only/both
years/2003 only
94
74
70
66
41
40
40
33
30
27
23
16
15
14
14
14
13
12
11
10
10
10
9
9
9
7
5
5
5
5
Total occurrences
(max 128)
23
41
13
50
31
71
70
70
106
151
106
43
101
100
70
106
35
92
111
97
19
48
42
70
258
322
211
149
115
169
Distance (km)
44.3
30.5
348.4
26.9
41.7
43.2
93.2
35.8
45.0
47.7
51.6
27.2
71.2
16.6
212.8
44.6
94.7
222.9
197.9
59.0
224.8
195.5
222.6
189.0
53.7
16.3
169.3
302.1
223.4
204.0
Direction (03608)
0.2126
0.0066*
0.5594
0.0258*
0.5262
0.1028
0.0688
0.0808
0.1160
0.0002**
0.1132
0.4002
0.2192
0.4590
0.4262
0.6408
0.6994
0.0882
0.0604
0.3570
0.9146
0.5198
0.8922
0.7186
0.2196
0.2768
0.3292
0.8000
0.7524
0.0958
p-value
Table 1. Presence/absence of epiphytic lichens were recorded on 64 sites in southern Sweden in 1986 and 2003. The shift in centroid (Distance) was calculated and its direction could be calculated for 30
cases (lichen species on tree species). The statistical significance of the shift in centroid was evaluated in a permutation test. *p-valueB0.05, **p-value highly significant also after adjusting for false
discovery rate.
Methodological considerations
The current lichen data set and its low power
All data sets suffer from shortcomings. In our case, field
sampling was designed to be quick, recording only
presence/absence of lichens that, in most cases, did not
require a specialist for identification. Consequently, we
were able to include more sites than if including, e.g.
assessment of abundance or demographic data. Future
power analyses would be welcome to strike an appropriate
compromise between data quality and quantity when
setting up monitoring studies in general.
We lacked information about trees without lichens and
were therefore unable to discriminate between actual losses
of lichens on specific substrates and the loss of the substrate
on a site. An additional consequence of this is that it
prevented us from evaluating a possible change in substrate
preference (e.g. a species might, over time, occur on a wider
range of tree species).
Another drawback of the simple field method is that the
age or size of the tree species, and where the lichen was
found (trunk or branches), were not recorded. Although
most of the epiphytic species sampled are not sensitive to
tree species or tree size, some may, e.g. prefer thin branches
over trunks and thus suffer a bias due to tree size
composition.
Three out of 30 tests turned out to be significant, which
might be interpreted as negligible change. It must be
remembered, though, that many of the tests were based
on very few occurrences (Table 1). Thus, the statistical
power was, in most cases, low or very low and only large
movements would be possible to detect.
The relatively large number of tests conducted (n 30)
might imply a need for adjusting for the family-wise
type-I error (rejecting a true null hypothesis). Only one
of three significant tests was found significant after applying the adjustment for false discovery rate (Benjamini
and Hochberg 1995). It is mainly a matter of opinion
whether the current analyses should be considered to
belong to the same family of tests or not (i.e. a justified
need for correction or none), so we present both. But
we focus on the uncorrected p-values for two reasons.
Firstly, as generally in monitoring, type-II errors (accepting
a true null hypothesis) might be equally, or more, harmful than type-I errors (Legg and Nagy 2006). Secondly,
we were interested in including all species to be able to
consider the directions several cases, each with a weak
signal, might together indicate a trend. This was, in fact,
the outcome also non-significant cases contributed to
the evidence suggesting a prevailing northeast direction
(Fig. 2, Table 1).
In conclusion, despite the low power of the data our new
method proved to work well, and there were significant
case(s) of centroid movements with a discernable prevailing
northeast direction.
The versatile permutation procedure
The above-mentioned shortcomings are mainly related to
this particular data set and its small sample size, and the
method of analysis has some general advantages and might
be useful in a broader range of analyses. First of all, it is
possible to study changes within distribution areas without
any knowledge of distribution boundaries. These are often
difficult to determine and also probably have poor
statistical properties (Dormann 2007). Some species do
not have their outermost localities within an area studied,
severely reducing the data sets available for boundary
analysis. With the current method, many more types of
data set could be analyzed.
100
p<0.05
p 0.05-0.1
p 0.1-0.25
p 0.25-0.5
p>0.5
80
Cumulative percentage
The estimates of shift of the centroids presented, 310 km
year1, is in range with the weighted centroids of counts of
wintering shorebirds in western Europe (1.56.0, MacLean
et al. 2008). Although not comparable, it might be noted
that published estimates of northward movement show an
average of 6.1 km per decade (Parmesan and Yohe 2003,
several organisms), 9.5 km per decade (Thomas et al. 1999,
birds) and 98.8 km per decade (Perry et al. 2005, marine
fishes). Further, simulations on the range expansions needed
for different organism groups to keep track with expected
climate warming, suggest less than 1 km in most cases, and
rarely longer than 10 km per annum (Malcolm et al. 2002).
Second, our study also showed that the prevailing
direction of movements of lichens is likely to be in a
northeast direction, rather than north, in southern Scandinavia. Most studies of species evaluating range shifts
assume a northward movement of organisms in the
Northern Hemisphere (Parmesan 2006). It is not easy
to justify this particular direction when considering the
global air circulation; the prevailing northeast direction
documented in the present study are more in line with
the global atmospheric circulations (de Blij and Muller
1996) and the gradients in temperature and rainfall in the
study area (SMHI 19872003).
60
40
20
All pairwise distances
Distances from permutations
0
0
45
90
135
180
225
270
315
360
Angle
Figure 2. The cumulative distribution of the calculated directions
of centroid movement (Table 1), that were based on inventories of
epiphytic lichens on trees at 64 sites in southern Sweden in 1986
and 2003. The size of the mark shows the probability of a
movement of centroid between the two years. The dotted line
show the angles of all possible pair-wise combinations of sites and
the unbroken line show the angles generated in the 30 permutation
tests. The two soft bumps are due to the slightly elongated shape of
the study area (Fig. 1a).
417
With presence/absence data, power decreases when
approaching both zero and 100% frequency of the
phenomenon under study. Hence, species with intermediate
abundance, occurring in ca 50% of sample points, would
have the strongest power in the present analysis. Basing
centroids on abundance data, would allow a strong analysis
also of very frequent species (cf. the weighted centroids used
by MacLean et al. 2008).
Acknowledgements The study was funded by the Swedish Royal
Academy of Science, Lunds Botaniska Förening (1986) and the
County of Stockholm. PM and MWP were partly funded through
the research program ENGO sponsored by the Swedish Environmental Protection Agency. We thank Karl-Olof Bergman, Anders
Hargeby, Lars Westerberg and Mikael Lönn for discussions,
Christopher Zetterberg for drawing the maps and Jennifer Larsson
for linguistic revision. Thanks also Kristina Articus (Usnea) and
Anders Nordin (Bryoria and Melanelia) for your help to determine
some rare lichen species.
References
Benjamini, Y. and Hochberg, Y. 1995. Controlling the false
discovery rate: a practical and powerful approach to multiple
testing. J. R. Stat. Soc. Ser. B 57: 289300.
Bridle, J. R. and Vines, T. H. 2007. Limits to evolution at range
margins: when and why does adaptation fail? Trends Ecol.
Evol. 22: 140147.
de Blij, H. J. and Muller, P. O. 1996. Physical geography of the
global environment (2nd ed.). John Wiley and Sons.
Dormann, C. F. 2007. Promising the future? Global change
projections of species distributions. Basic Appl. Ecol. 8: 387
397.
Hickling, R. et al. 2005. A northward shift of range margins in
British Odonata. Global Change Biol. 11: 502506.
Hultengren, S. et al. 2004. Recovery of the epiphytic lichen flora
following air quality improvement in southwest Sweden.
Water Air Soil Pollut. 154: 203211.
IVL 2009. Swedish environmental research institute, Bwww.
ivl.se, accessed May 2009.
Karlsson, T. 1997. Förteckning över svenska kärlväxter. Sv. Bot.
Tidskr. 91: 241560.
418
La Sorte, F. A. and Thompson III, F. R. 2007. Poleward whifts in
winter ranges of North American birds. Ecology 88: 1803
1812.
Legg, C. J. and Nagy, L. 2006. Why most conservation
monitoring is, but need not be, a waste of time. J. Environ.
Manage. 78: 194199.
MacLean, I. L. M. D. et al. 2008. Climate change causes rapid
changes in the distribution and site abundance of birds in
winter. Global Change Biol. 14: 24892500.
Malcolm, J. A. et al. 2002. Estimated migration rates under
scenarios of global climate change. J. Biogeogr. 29: 835849.
Mattsson, J.-E. 1988. Några Cetraria-arters tillbakagång i södra
Sverige. Sv. Bot. Tidskr. 82: 2731.
Parmesan, C. 2006. Ecological and evolutionary responses
to recent climate change. Annu. Rev. Ecol. Evol. Syst. 37:
637669.
Parmesan, C. and Yohe, G. 2003. A globally coherent fingerprint
of climate change impacts across natural systems. Nature
421: 3742.
Perneger, T. V. 1998. What’s wrong with Bonferroni adjustments.
Brit. Med. J. 316: 12361238.
Perry, A. L. et al. 2005. Climate change and distribution shifts in
marine fishes. Science 308: 19121915.
Proschan, M. A. and Waclawiw, M. A. 2000. Practical guidelines
for multiplicity adjustment in clinical trials. Control. Clin.
Trials 21: 527539.
Root, T. L. et al. 2003. Fingerprints of global warming on wild
animals and plants. Nature 421: 5760.
Santesson, R. et al. 2004. Lichen-forming and lichenicolous fungi
of Fennoscandia. Uppsala Univ.
SMHI 19872003. Väder och vatten. Swedish Meteorological
and Hydrological Institute, Norrköping, Sweden.
Thor, G. and Arvidsson, L. 1999. Rödlistade lavar i Sverige,
artfakta. ArtDatabanken.
Thomas, C. D. and Lennon, J. J. 1999. Birds extend their ranges
northwards. Nature 399: 213213.
van Herk, C. M. et al. 2002. Long-term monitoring in the
Netherlands suggests that lichens respond to global warming.
Lichenologist 34: 141154.
Warren, M. S. et al. 2001. Rapid responses of British butterflies to
opposing forces of climate and habitat change. Nature 414:
6569.
PAPER IV
Biodiversity in the wake of urban sprawl: loss among
epiphytic lichens on large oaks
Håkan Lättman, Karl-Olof Bergman, Malin Rapp, Malin Tälle, Lars Westerberg and Per Milberg
Håkan Lättman, IFM Biology, Linköping University, SE-581 83 Linköping, Sweden, and School
of Life Sciences, Södertörn University, SE-181 49 Huddinge, Sweden
Karl-Olof Bergman, IFM Biology, Linköping University, SE-581 83 Linköping, Sweden
Malin Rapp, IFM Biology, Linköping University, SE-581 83 Linköping, Sweden
Malin Tälle, IFM Biology, Linköping University, SE-581 83 Linköping, Sweden
Lars Westerberg, IFM Biology, Linköping University, SE-581 83 Linköping, Sweden
Per Milberg, IFM Biology, Linköping University, SE-581 83 Linköping, Sweden
1
Abstract
Biodiversity often suffers from urbanisation. In the present study, we focused on how the age of
urbanisation affects the richness of 17 epiphytic lichens species and their cover on large oaks,
with a minimum spacing of 250 m, in urban environments in the city of Linköping (100,000
inhabitants), SE Sweden. We also surveyed trees in adjacent rural areas, selected to have similar
distributions of tree trunk circumference and oak density within 300 m. Lichen richness and
cover were significantly lower on urban trees compared to rural trees. Furthermore, richness and
cover decreased with the length of time that urban trees had been surrounded by houses. Roughly
one species is lost every 30 years. Most of the species that were analysed demonstrated a drop in
occurrence with respect to the duration of housing development. The reduction in the probability
of occurrence varied from 60% (Calicium viride, Evernia prunastri), 80% (Chrysotrix
candelaris) to 90% (Ramalina spp.) during the 160-year period of urbanisation considered.
Therefore, even if valuable trees survive over the course of development, their lichen flora are
likely to become depleted over time.
Key words: housing; landscape; Linköping; Quercus; spatial; urbanization
2
Introduction
Urbanisation alters the natural environment in a number of ways and causes local extinction of
species (McKinney, 2006). In the USA, urbanisation is considered to be one of the worst threats
to endangered species (Czech, Krausman, & Devers, 2000). Many studies have shown that when
moving from rural areas to the central parts of a city, there is a gradual decrease in species
richness in a number of organism groups, such as butterflies (Blair, 1999; Blair & Launer, 1997),
beetles (Ishitani, Kotze, & Niemelä, 2003; Niemelä & Kotze, 2009; Su, Zhang, & Qiu, 2011),
birds (Blair, 1999; Gagne & Fahrig, 2011), amphibians (Price, Dorcas, Gallant, Klaver &
Willson, 2006; McKinney, 2008), reptiles (McKinney, 2008), mammals (McKinney, 2008), and
plants (McKinney, 2008). However, there are a few studies that show that urbanisation has no
effect on biodiversity. For instance, Carabidae assemblages showed no changes across urban
gradients in three cities in Finland, Bulgaria and Canada with three different levels of
urbanisation (Niemelä et al., 2002). A study in the lower Florida Keys concluded that the native
ant fauna was not, at the moment, negatively affected by age of urbanisation (Forys & Allen,
2005). However, with a few exceptions, urbanisation generally seems to reduce biodiversity.
Urbanisation affects various environmental factors that in turn affect survival and
reproduction of species. The mechanisms involved in species loss vary, spanning from habitat
loss and fragmentation (Robinson, Newell, & Marzluff, 2005) to assumed shifts in biotic
interactions and possible adverse effects of pollution (Tarhanen, Poikolainen, Holopainen, &
Oksanen, 2000). One group that is particularly sensitive to pollution is lichens. When pollution
exceeds a critical level, some species become locally extinct (van Herk, 2001). In urban areas,
sulphur and nitrogen oxides primarily cause the greatest damage to lichens (Giordani, 2007).
Some lichens are less sensitive to air pollution, but if levels of sulphur oxide exceed 125 µg/m3,
very few lichens survive (Johnson, 1979; LeBlanc, Rao, & Comeau, 1972; de Wit, 1976). Based
on the fact that different lichen species have different tolerances for air pollutants, Hawksworth
and Rose (1970) created a method to assess the level of sulphur dioxide in the air based on
presence of epiphytic lichens.
Lichens also have different tolerances and preferences for substrates. Some lichens are
restricted to large trees (Washburn & Culley, 2006) that usually have a coarse bark structure.
Such specialised lichens have declined as the density of that substrate has diminished due to
agriculture and forestry (Hedenås & Ericson, 2008; Belinchón et al., 2009). However, trees are
often an important feature of cities and are often allowed to grow old for a number of reasons,
including human wellbeing, economic value, aesthetics, shade production, and in some
circumstances, fire prevention (Lohr, Pearson-Mims, Tarnai, & Dillman, 2004; Donovan &
Butry, 2010, 2011). An example is the large number of Quercus robur trees, which have been
maintained throughout urbanisation, in urban Linköping, a city with 100,000 inhabitants in
southeastern Sweden. Mobile organisms can escape urban habitats, while trees (and lichens) are
inevitably incorporated into the urban environment. Lichens on the remaining trees are thus
isolated, and their dispersal and establishment may be negatively affected. However, the future
3
prospects for epiphytic organisms are better than those of many other elements of urban
biodiversity as trees will continue to be an important feature of cityscapes.
In the present study, a selection of lichen species were surveyed on Quercus robur trees in
urban and rural environments. The objectives of this study can be divided into four questions: Is
there any difference with regard to (i) species richness and (ii) cover of epiphytic lichens on Q.
robur in urban or rural environment? If there are any differences are these affected by (iii)
species richness and (iv) cover of epiphytic lichens on Q. robur affected by the age of
urbanization? To make these comparisons, we matched a population of urban oaks with a
selection of rural oaks based on tree circumference and the density of surrounding oaks because
oak lichens are sensitive to both the size of their host tree (Johansson, Bergman, Lättman, &
Milberg, 2009) and the density of surrounding oaks (Muhammadi, 2011). The design of this
study allowed us to separate the effect of habitat loss from the effect from the age of
urbanization, a feature rarely considered.
4
Materials and Methods
Study area
The study was conducted in Linköping city and surrounding rural areas south of the city, in the
province of Östergötland, Sweden (Fig. 1). The area north of Linköping municipality was
avoided due to differences in soil and vegetation characteristics from the urban areas. The study
area is located at an altitude of 30 to 125 m above sea level and consists of the urban city centre
of Linköping, urban district and surrounding rural environments. The city of Linköping is the
seventh largest city in Sweden with a population of 104,232 in 2010 (Linköpings kommun,
2011). It was founded in the 11th century, but the majority of buildings have been built since the
mid-1800s. The annual average temperature at the nearest meteorological station (urban district
Malmslätt) is 6.1 ºC, with an average temperature of –3.2 ºC in January and 16.2 ºC in July; and
the average precipitation is 516 mm per year (Statistics Sweden, 2011). Linköping has expanded
concentrically (Fig. 2), and the city centre consists largely of three to five storey buildings with
shops at the street level. Blocks composed of flats and one to two-storey houses make up the
greater part of the residential areas in the urban district. Squares, parks and paved roads are both
in the city centre and urban district. The surrounding rural areas consist mainly of a mixed
landscape with deciduous forests, planted conifers, arable land and pastures.
Selection of Quercus robur and lichen species
We studied the lichen flora on pedunculate oak (Quercus robur). Quercus robur is common in
and around Linköping, and urban trees were selected for study from an existing tree database.
The database contains the position and circumference of all large Q. robur, as well as several
other tree species in the province of Östergötland, and was developed by the County
Administration Board in Linköping (available at http://gis.lst.se/lstgis/). Using ArcMap 9.3
(ESRI, 2011), all Q. robur within the municipality of Linköping with a circumference >250 cm
were chosen because these trees can be assumed to be old enough to host a high diversity of
lichen flora (e.g., Johansson et al., 2009). The age of the sampled trees can be estimated from
dendrological studies in the area (Berg, 2006) to be at least 180 to 240 years; therefore, the
establishment of most trees predates urbanisation by at least 100 to 150 years. A grid was added
over the urban area, and in each cell in the grid where Q. robur trees were present, the oak tree
closest to the centre of the cell was chosen; this method ensured that study trees were at least 250
m apart. From the database, we selected 105 urban trees, which were matched with a population
of rural trees (Fig. 1). The criteria for matching urban trees with rural trees were that (i) the
distribution of trunk circumference and (ii) the density of trees within 302 m should be similar
since Muhammadi (2011) has shown that species richness has a high degree of explanation at
this distance. Some selected rural trees were excluded during the fieldwork because they had
5
been cut down, or were in private gardens; others were substituted with similarly sized trees for
the same reasons. In the end, we used 105 and 109 trees in urban and rural environments.
The seventeen lichen species selected for our survey are shown together with their red-list
category (Gärdenfors, 2010) and, when appropriate, if the species are used as an indicator
species (Nitare, 2010), as well as with substrate preference (Santesson, Moberg, Nordin,
Tønsberg, & Vitikainen, 2004) (Table 1). Nine species were commonly found, and the majority
occur more or less throughout all of Sweden. Some of them are considered resistant to air
pollutants. Eight lichens are rare and red-listed by the threat categories near threatened (NT) or
vulnerable (VU), and/or indicator species for high nature conservation and are thought to be
sensitive to air pollution. Two additional lichen species were initially included but were excluded
in the beginning of the fieldwork as they turned out too difficult to reliably identify in the field.
After the fieldwork, it was concluded that three species (Calicium quercinum, Lecanographa
amylacea and Schismatomma decolorans) had never been registered.
Field survey
The fieldwork was conducted during April, May, October and November, 2011. For every oak,
circumference at breast height was measured to the nearest centimetre. The depth of tree trunk
bark crevices was measured in the north, east, south and west sides of the trunk to the nearest
millimetre using a ruler. Based on these four measurements, a mean bark crevice depth was
calculated for each tree. Sun exposure was estimated by assessing how much sun in per cent to
the trunk was received, taking into account the shade cast by nearby buildings, shrubs and trees.
On each tree trunk, lichens were searched for over a one-metre mantle area starting 50 cm above
the ground to avoid the influence of environmental effects at the base as they vary greatly
between individual trees. For each of the target lichen species, the area they covered was
estimated to the nearest cm2 and was expressed as the percentage of the inspected trunk area.
Additional parameters
Some complementary explanatory variables were also considered (Table 2). The number of Q.
robur trees around each tree, including the tree being examined, was used as an explanatory
variable. Six different radii were used: 150, 250, 350, 500, 700 and 1000 m. The same range of
radii was also used to determine the area of buildings around each studied Q. robur. The data on
buildings were drawn from the GSD-Topographic Map (Lantmäteriet,
http://www.lantmateriet.se/) and combined four different classes of buildings (industrial
buildings, one and two storey houses, apartment buildings and city blocks). Age of urbanisation
was estimated from the average year of construction of the five building closest to the target tree,
using a radius of approximately 200 m (data from the municipality of Linköping). Fewer
buildings were used if the year of construction was not available for all five buildings or if fewer
than five buildings were within 200 m of the tree. For the city centre (Fig. 2), where some older
buildings have been demolished and replaced with newer ones, the above approach would
underestimate the age. We therefore assigned the age 1900 to seven buildings that were erected
by 1950 or later (hence replacing older buildings). One tree associated with buildings in the area
dated to the eighteenth century (Fig. 2) was assigned the age 1850. In total, these age
assignments affected the urban age estimates of eight trees in the city centre, and are most likely
to be underestimates of the true ages.
6
Statistical analyses
The number of urban and rural trees of Q. robur used in the statistical analyses was 105 and 109,
respectively, and in total, 14 lichen taxa were observed: nine common and five rare.
To evaluate the data, regression analyses (generalised linear model, GLM) were run in
Statistica 10 (Statsoft, 2011). The first set of analyses involved the number of target species per
tree on urban vs. rural trees (normal distribution; identity link), and the total cover of target
lichen per tree for urban vs. rural trees (normal distribution; log link). Rural trees were not
included in any further analysis.
The second set of analyses evaluated the relationship between (i) the number of target
species per tree (normal distribution; identity link), and (ii) the total cover of target lichen per
tree (normal distribution; log link), as related to number of years in an urban setting (i.e., the
average age of the five closest buildings).
The third set of analyses involved the species-wise occurrence of the eight most frequent
species (binomial distribution; logit link). For each radius (150–1000 m), a model was made
including the following candidate explanatory variables: (i) age of urbanisation, (ii) tree
circumference, (iii) bark crevice depth, (iv) sun exposure, (v) density of oaks, and (iv) cover of
buildings. Using AIC, explanatory variables, including radii, were selected. Depth of bark
crevices was chosen as it is often considered a better proxy for tree age than is tree
circumference (Barkman, 1958; Pedersen, 1980; see Johansson et al., 2009). However, as the
depth and circumference were poorly correlated, both were considered as candidate variables.
Tree circumference and density of oaks were log transformed before the analysis.
Finally, five attributes of the fourteen species were analysed in relation to their average
cover (square-root transformed) on urban oaks: 1) spores (no, rarely, yes; ordinal multinomial;
logit); 2) diaspore (whether soredia and isidia occur or not; binomial; logit); 3); spore area (a
measure of spore size; normal distribution; identity link function); 4) pycnidium (no, rarely, yes;
ordinal multinomial; logit); and 5) growth form (crustose, fruticose, foliose; multinomial; logit)
(Foucard, 2001; Nash III, Ryan, Gries, & Bungartz, 2002; Nash III, Ryan, Diederich, Gries, &
Bungartz, 2004; Nash III, Gries, & Bungartz, 2007).
7
Results
In total, 214 trees of Quercus robur were studied, of which 105 and 109 were in urban and rural
environments, respectively. Out of the 17 lichen species that were searched for, three rare and
red-listed species (Calicium quercinum, Lecanographa amylacea and Schismatomma
decolorans) were not found (Table 1). The number of lichen species per oak varied between 0
and 9. The total number of observations of lichen species was 954, with 400 observations in
urban areas and 554 in rural areas.
Species richness and cover: urban vs. rural
There were clear differences in lichen species occurrence and cover of individual lichen species
on oaks in the urban area compared to oaks in rural environments (Fig. 3). Four taxa were only
found in the rural environment (Calicium adspersum, Cyphelium inquinans, Sclerophora
coniophaea, Usnea spp.).
All lichen taxa, except Lepraria incana, occurred more frequently on rural oaks than on
urban oaks (Fig. 3a). All lichen taxa (including Lepraria incana) had a higher cover on rural than
urban oaks, both when looking at averages over all trees (including zero cover observations) and
when considering only trees on which the species were present (Fig. 3b and c).
The number of target species per tree and the cover of target lichens were significantly
higher on oaks in rural environments than on oaks in urban area (P < 0.0001 in both cases) (Fig.
4a and c). The mean species richness was 34% higher on rural oaks than urban oaks (5.1 and 3.8
lichen species, respectively) (Fig. 4a). When looking at the total cover of the 14 lichen species,
the differences between urban oaks and rural oaks were even more pronounced, with a mean
cover many times larger on rural oaks than on urban oaks (0.041% and 0.007%, respectively)
(Fig. 4c).
Species richness and cover: effects from age of urbanisation
The number of lichen species per tree decreased significantly with the increasing age of the
surrounding buildings (P = 0.00031) (Fig. 4b). The average number of target species in the most
recently urbanised areas was four times higher than in the oldest parts (Fig. 4b), and a similar
decline was found in the cover of target lichens (Fig. 4d).
Species-wise responses to environmental factors in urban areas
Out of the 14 lichen species, eight were sufficiently frequent to be analysed individually in
consideration of all explanatory variables. The selected models proved significant for seven of
the eight species (Table 3). The majority of the species were affected by urbanisation factors or
oak size factors, while few species were affected by other biotic factors such as oak density. The
age of urbanisation had a significant negative effect on five species and the area covered by
buildings on two species. Chrysotrix candelaris and Hypogymnia physodes were positively
8
affected by the area covered by buildings (Table 3). The lichen species responded mostly to area
covered by buildings at larger scales, ≥350 m. A higher tree circumference or deeper bark
crevices had a significant positive effect on the occurrence of four species and a negative effect
on one. Sun exposure had a significant negative impact on two species (Table 3). Oak density
affected only one species significantly negatively and none positively.
For the four species where the age of urbanisation significantly affected the occurrence,
separate binomial (logit link) GLMs were conducted. All species showed a decrease in
prevalence in the older parts of the city in comparison to the newer parts. The models predicted a
reduction of 60% (Calicium viride, Evernia prunastri), 80% (Chrysotrix candelaris) and 90%
(Ramalina spp.) probability of occurrence during the 160 years of urbanisation (Fig. 5).
Out of the six tested species attributes (Table 4), only one had a significant association with
cover on urban trees: „spores‟, i.e., to what extent a species rely on spore dispersal (P = 0.039).
9
Discussion
This study shows a clear reduction in species richness and abundance of lichens on oaks in urban
areas compared to rural oaks, as well as with age of urbanisation in the urban areas, a finding that
is in line with studies on other groups (Blair, 1999; Blair & Launer, 1997; Gagne & Fahrig,
2011; Ishitani et al., 2003; McKinney, 2008; Niemelä & Kotze, 2009; Price et al., 2006; Su et al.,
2011). Our results indicate that the dispersal mode may be an important factor in predicting
which species will be affected by urbanisation.
Lower richness and cover on urban oaks
Species in urban areas are affected by a complex interaction between factors, such as pollution,
temperature, moisture, disturbance and habitat configuration (McDonnell et al., 1997). Generally
habitat loss and fragmentation are among the most important factors in urban areas (McKinney,
2006; Niemelä, 1999), which for most species is negative. A high density of large oak trees has
proved to be important for species richness and the occurrence of some lichens (Paltto,
Thomasson, & Nordén, 2010; Muhammadi, 2011), showing that habitat configuration (or
connectivity) is important. The sample populations of urban and rural oaks were matched for
both circumference and surrounding tree density, thereby concurrently allowing for comparison
while controlling for habitat quality (circumference) and connectivity. Several species are
generalist and not confined to oaks, but because previous studies emphasize the importance of
oak (Paltto et al., 2010; Muhammadi, 2011), we argue that our sample design can control for
differences in habitat connectivity. Thus, by controlling for habitat age and habitat composition
(density of surrounding oaks), we were able to single out the effect of urbanisation, from that of
a general decrease in habitat. In cases where it was possible to analyse the species, our study
showed that urbanisation had a clear effect on species richness and cover (Table 3). The
explanatory variable, i.e., age of urbanisation, showed that five species were negatively affected,
while area covered by buildings showed that three species were negatively affected and one
species was positively affected.
A decrease of lichens in urban environments was reported very early during industrialisation
in several places in Europe (Grindon, 1859; Nylander, 1866). The main factor for the decline is
sulphur dioxides from industries and traffic (Gilbert, 1968). Sulphur dioxides peaked during the
1960s and 1970s, and lichen deserts in urban areas were reported at these elevated emission
levels (Hawksworth & Rose, 1970). However, air quality has recently been shown to have
improved in many urban areas (Lisowska, 2011). In Tampere, Finland, sulphur dioxide levels
were reduced from 160 μg m–3 in 1973 to 2 μg m–3 in 1999, which coincided with an increase in
lichen epiphytic richness (>10 times) and cover (>200 times) since 1980 (Ranta, 2001). A
reduction has also occurred in the county of Östergötland, where sulphur dioxide levels have
decreased from 5290 tonnes per year in 1990 to 1588 tonnes per year in 2009, and from 2786 to
347 tonnes per year for Linköping municipality (RUS, 2012). In our study area, there is no
longer any ground for talking of a “lichen desert” in the city: There were no differences between
10
the number of oaks without target species in urban and rural areas (4 out of 105 urban oaks
without target lichens and 3 of 109 rural oaks). However, there were still large differences in
species diversity and the cover of individual species between urban and rural oaks. Our results
show that species that disperse by spores were often absent or had low cover, while those that
were not spore-dispersed were among those with the highest cover (Table 4). Species dependent
on the process of re-lichenisation with a photobiont after successful dispersal seem to be less
likely to succeed in urban areas. The species most sensitive to urbanisation in our study all
belong to this group. Calicium adspersum, Cyphelium inquinans, Sclerophora coniophaea and
Usnea spp. could not be found on oaks in the urban environment, but they were present in the
rural environment. Calicium adspersum, C. inquinans and S. coniophaea are classified as
indicators of forest areas with high conservation values (Nitare, 2010), while the taxon Usnea
spp. is known to be sensitive to air pollution (Hawksworth & Rose, 1970). In addition to the
dispersal mode, the photobiont may be important. Tarhanen et al. (2000) have demonstrated that
the green algae Trebouxia is sensitive to air pollution. Their results showed that high
concentrations of pollutants in the air increased plasmolysis and mitochondrial changes in cells.
Furthermore, degenerated cells showed altered chloroplasts and electron-translucent
pyrenoglobuli as far as 35 to 50 km from the pollutant source. The low cover on many urban
oaks compared to that on rural oaks indicates that something affects lichens in urban areas, even
though the air quality has improved. Armstrong and Bradwell (2010) reviewed 52 studies with
reported growth rates, and of which 33 of the surveys had been performed in Europe. If we adopt
an average European radial growth rate of 0.92 mm per year, which seems reasonable, the lichen
thallus would need 6 years to reach a size of one centimetre in diameter and thus become clearly
visible to the naked eye. For the five foliose and fruticose lichens, the growth rate is available for
Hypogymnia physodes: approximately 3 to 5 mm in diameter per year (Gorbach & Kobzar, 1981;
unpublished data), with a corresponding shorter time to reach visibility. Given that levels of
sulphur dioxide have decreased over the past 30 years, there should have been ample time for
regrowth for most lichens, given that they are not dispersal or photobiont limited and that current
SO2 levels are sufficiently low.
Lichens are also known to be sensitive to changes in temperature and to respond to global
warming (van Herk, Aptroot, & van Dobben, 2002), and most of the species favour a moist
environment. Therefore, shifts in temperature and moisture might be important for urban lichens.
Several studies have shown that the temperature is higher in urban than in rural areas (Bulut,
Toy, Irmak, Yilmaz, & Yilmas, 2008; George, Ziska, Bunce, & Quebedeaux, 2007; Liu, You, &
Dou, 2009), a phenomenon known as “urban heat island”. Hughes (2006) investigated the “urban
heat island effect” in four cities in the United Kingdom of varying population size. Norwich had
on average urban temperature of 0.5°C higher than the rural surroundings. Because the
population size of Norwich is similar to Linköping‟s, we can assume the latter has a similar
temperature difference. Relative humidity (RH) is, to a large extent, a function of temperature, so
we would expect the urban heat island effect to coincide with lower RH. Furthermore, hard
surfaces, efficient canalisation of runoff water, and small volumes of vegetation (low
evapotranspiration) in urban areas also suggest lower RH. Nevertheless, evidence from the
measurements is conflicting (George et al., 2007 found no differences, while Liu et al., 2009
did).
11
Loss rate over the course of urbanisation
Numerous studies have investigated how a spatial gradient from rural, suburban to urban areas
affects the number of organisms and their quantities. Most studies have concluded that the
number of species and their amount decreases with an increasing degree of urbanisation
(Mercado Cárdenas & Buddle, 2009), but there are exceptions where the cause is often alien
species (Dolan, Moore, & Stephens, 2011) and generalist species (Magura, Tóthmérész, &
Molnár, 2008; Tóthmérész, Máthé, Balázs, & Magura, 2011). The same reduction of species and
their amount toward the city centre can also be observed in temporal studies. For instance,
Fattorini (2012) made a reconstruction of the extinction trends of four insect groups in urban
Rome, Italy from 1885–1999. Her results showed a clear decline in species richness in each
group of insects. Price et al. (2006) examined the incidence of species from the family
Salamandridae near Davidson, North Carolina, USA, for a period of 30 years and found that the
populations declined over time.
Our results showed a clear effect on species number and cover with the length of time that
urbanisation had been taking place. Of all 14 lichens, eight were analysed individually (Table 3).
Age of urbanisation, or the area covered by buildings around the trees, affected seven species
negatively. It is of interest to know the rate of species loss, or cover, during urbanisation. We
estimated a decrease in species richness and the total cover of target lichen species per tree (Fig.
4b and 4d) and a lowering probability of occurrence by 3.7–5.6% per decade depending on
lichen species (Figure 5). In summary, the relatively slow but steady loss of biodiversity
confirms the assumption of Hahs et al. (2009) that modern cities potentially carry a large
extinction debt; however, the interpretation is complicated by the return of lichens following
recent improvements in air quality (see above).
Conclusion
We have shown major differences in lichen species richness and cover between urban and rural
environments, as well as a clear decrease with the age and degree of urbanisation on remnant
large oaks in a city. Therefore, even if the prospects are good for a continuous supply of
epiphytic substrate in cities – where trees are likely to exist – their value for epiphytic lichens
seems limited. For the urban planner, this means two things. First, retaining individual valuable
trees during urbanisation does not automatically preserve their lichen biodiversity. Second, the
preservation of lichen biodiversity is most likely best achieved if urban development leaves
valuable areas of trees with rich lichen populations intact and relatively large and plan for
connecting fragments of old trees instead of embedding them.
12
Acknowledgements
Thanks to Lena Samuelsson and Linköping municipality for assistance with obtaining maps of
Linköping‟s different urban district boundaries and the age of buildings.
13
References
Armstrong, R., & Bradwell T. (2010). Growth of crustose lichens: a review. Geografiska
Annaler Series A Physical Geography, 92(1), 3-17.
Barkman, J. J. (1958). Phytosociology and ecology of cryptogamic epiphytes: including a
taxonomic survey and description of their vegetation units in Europe. Assen, Van Gorcum.
Belinchón, R., Martínez, I., Otálora, M. A. G., Aragón, G., Dimas, J., & Escudero, A. (2009).
Fragment quality and matrix affect epiphytic performance in a Mediterranean forest
landscape. American Journal of Botany, 96, 1974-1982.
Berg, N. (2006). Age and size of hollow oaks and their associated lichen flora and beetle fauna.
Final thesis, Linköping University, LiTH-IFM-Ex--05/1605--S.
Blair, R. B. (1999). Birds and butterflies along an urban gradient: surrogate taxa for assessing
biodiversity? Ecological Applications, 9, 164-170.
Blair, R. B., & Launer, A. E. (1997). Butterfly diversity and human land use: species
assemblages along an urban gradient. Biological Conservation, 80, 113-125.
Bulut, Y., Toy, S., Irmak, M. A., Yilmaz, H., & Yilmas, S. (2008). Urban-rural climatic
differences over a 2-year period in the city of Erzurum, Turkey. Atmósfera, 21(2), 121-133.
Czech, B., Krausman, P. R., & Devers, P. K. (2000). Economic associations among causes of
species endangerment in the United States. BioScience, 50, 593-601.
de Wit, T. (1976). Epiphytic lichens and air pollution in the Netherlands. Bibliotheca
Lichenologica, 5, 1-227.
Dolan, R. W., Moore, M. E., & Stephens, J. D. (2011). Documenting effects of urbanization on
flora using herbarium records. Journal of Ecology, 99, 1055-1062.
Donovan, G. H., & Butry, D. T. (2010). Trees in the city: valuing street trees in Portland,
Oregon. Landscape and Urban Planning, 94, 77-83.
Donovan, G. H., & Butry, D. T. (2011). The effect of urban trees on the rental price of singlefamily homes in Portland, Oregon. Urban Forestry and Urban Greening, 10, 163-168.
ESRI. (2011). ArcGIS Desktop: Release 10. Redlands, CA: Environmental Systems Research
Institute.
Fattorini, S. (2012). Insect extinction by urbanization: a long term study in Rome. Biological
Conservation, 144, 370-375.
Forys, E. A., & Allen, C. R. (2005). The impacts of sprawl on biodiversity: the ant fauna of the
lower florida keys. Ecology and Society, 10(1), 25. [online] URL:
http://www.ecologyandsociety.org/vol10/iss1/art25/.
Foucard, T. (2001). Svenska skorplavar och svampar som växer på dem. Interpublishing,
Stockholm. [in Swedish]
Gagne, S. A., & Fahrig, L. (2011). Do birds and beetles show similar responses to urbanization?
Ecological Applications, 21, 2297-2312.
14
George, K., Ziska, L. H., Bunce, J. A., & Quebedeaux, B. (2007). Elevated atmospheric CO2
concentration and temperature across an urban–rural transect. Atmospheric Environment, 41,
7654-7665.
Gilbert, O. L. (1968). Bryophytes as indicators of air pollution in the Tyne Valley. New
Phytologist, 67(1), 15-30.
Giordani, P. (2007). Is the diversity of epiphytic lichens a reliable indicator of air pollution? A
case study from Italy. Environmental Pollution, 146, 317-323.
Gorbach, N V., & Kobzar, N. N. (1981). Determination of the age of epiphyte lichen
Hypogymnia physodes (L.) Nyl. Doklady Akademii Nauk BSSR [Minsk], 25, 848-851.
Grindon, L. H. (1859). The Manchester flora. W. White, London.
Gärdenfors, U. ed (2010). The 2010 red list of Swedish species. ArtDatabanken, SLU, Uppsala.
Hahs, A. K., McDonnell, M. J., McCarthy, M. A., Vesk, P. A., Corlett, R. T., Norton, B. A.,
Clemants, S. E., Duncan, R. P., Thompson, K., Schwartz, M. W., & Williams, N. S. G.
(2009). A global synthesis of plant extinction rates in urban areas. Ecology Letters, 12,
1165-1173.
Hawksworth, D. L., & Rose, F. (1970). Qualitative scale for estimating sulphur dioxide pollution
in England and Wales using epiphytic lichens. Nature, 227, 145-148.
Hedenås, H., & Ericson, L. (2008). Species occurrences at stand level cannot be understood
without considering the landscape context: cyanolichens on aspen in boreal Sweden.
Biological Conservation, 141, 710-718.
Hughes, K. (2006). The impact of urban areas on climate in the UK: a spatial and temporal
analysis, with an emphasis on temperature and precipitation effects. Earth & E-nvironment,
2, 54-83.
Ishitani, M., Kotze, D. J., & Niemelä, J. (2003). Changes in carabid beetle assemblages across an
urban-rural gradient in Japan. Ecography, 26, 481-489.
Johnson, D. W. (1979). Air pollution and the distribution of corticolous lichens in Seattle,
Washington. Northwest Science, 53, 257-263.
Johansson, V., Bergman, K.-O., Lättman, H., & Milberg, P. (2009). Tree and site quality
preferences of six epiphytic lichens growing on oaks in southeastern Sweden. Annales
Botanici Fennici, 46, 496-506.
LeBlanc, F., Rao, D. N., & Comeau, G. (1972). The epiphytic vegetation of Populus balsamifera
and its significance as an air pollution indicator in Sudbury, Ontario. Canadian Journal of
Botany, 50, 519-528.
Linköpings kommun. (2011). Linköpings kommun, Statistik & utredningar, Statistikinfo
(Linköping kommun, Statistics & reports, statistical information) 2011:7.
Lisowska, M. (2011). Lichen recolonisation in an urban-industrial area of southern Poland as a
result of air quality improvement. Environmental Monitoring and Assessment, 179, 177-190.
Liu, W., You, H., & Dou, J. (2009). Urban-rural humidity and temperature differences in the
Beijing area. Theoretical and Applied Climatology, 96, 201-207.
Lohr, V. I., Pearson-Mims, C. H., Tarnai, J., & Dillman, D. A. (2004). How urban residents rate
and rank the benefits and problems associated with trees in cities. Journal of Arboriculture,
30, 28-35.
Magura, T., Tóthmérész, B., & Molnár, T. (2008). A species-level comparison of occurrence
patterns in carabids along an urbanisation gradient. Landscape and Urban Planning, 86,
134-140.
15
McDonnell, M. J., Pickett, S. T. A., Groffman, P., Bohlen, P., Pouyat, R. V., Zipperer, W. C.,
Parmelee, R. W., Carreiro, M. M., & Medley, K. (1997). Ecosystem processes along an
urban-to-rural gradient. Urban Ecosystems, 1, 21-36.
McKinney, M. L. (2006). Urbanization as a major cause of biotic homogenization. Biological
Conservation, 127, 247-260.
McKinney, M. L. (2008). Effects of urbanization on species richness: a review of plants and
animals. Urban Ecosystems, 11, 161-176. doi:10.1007/s11252-007-0045-4
Mercado Cárdenas, A., & Buddle, C. M. (2009). Introduced and native ground beetle
assemblages (Coleoptera: Carabidae) along a successional gradient in an urban landscape.
Journal of Insect Conservation, 13, 151-163.
Muhammadi, U. H. (2011). Spatial pattern of occurrence of eleven epiphytic lichen species in a
heterogeneous landscape. Master thesis, Linköping University, LiTH-IFM-Ex--11/2441--SE.
Nash III, T. H., Ryan, B. D., Gries, C., & Bungartz, F. (eds.) (2002). Lichen flora of the greater
Sonoran Desert region. Vol. I. State University, Tepme, Arizona.
Nash III, T. H., Ryan, B.D., Diederich, P., Gries, C., & Bungartz, F. (eds.) (2004). Lichen flora
of the greater Sonoran Desert region. Vol. II. State University, Tepme, Arizona.
Nash III, T. H., Gries, C., & Bungartz, F. (eds.) (2007). Lichen flora of the greater Sonoran
Desert region. Vol. III. State University, Tepme, Arizona.
Niemelä, J. (1999). Is there a need for a theory of urban ecology? Urban Ecosystems, 3, 57-65.
Niemelä, J., & Kotze, D. J. (2009). Carabid beetle assemblages along urban to rural gradients: a
review. Landscape and Urban Planning, 92, 65-71.
Niemelä, J., Kotze, D. J., Venn, S., Penev, L., Stoyanov, I., Spence, J., Hartley, D., & de Oca, E.
M. (2002). Carabid beetle assemblages (Coleoptera, Carabidae) across urban-rural gradients:
an international comparison. Landscape Ecology, 17, 387-401.
Nitare, J. (2010). Signalarter. Indikatorer på skyddsvärd skog. Flora över kryptogamer (4th ed.)
(Indicators for valuable forest areas. A cryptogam flora). Jönköping: Skogsstyrelsen.
Nylander, W. (1866). Les lichens du Jardin du Luxembourg (Lichens in Jardin du Luxembourg).
Bulletin de la Société Botanique de France, 13, 364-372.
Paltto, H., Thomasson, I., & Nordén, B. (2010). Multispecies and multiscale conservation
planning: setting quantitative targets for red-listed lichens on ancient oaks. Conservation
Biology, 24, 758-768.
Pedersen, I. (1980). Epiphytic lichen vegetation in an old oak wood, Kaas skov. Botanisk
tidsskrift, 75, 105-120.
Price, S. J., Dorcas, M. E., Gallant, A. L., Klaver, R. W., & Willson, J. D. (2006). Three decades
of urbanization: estimating the impact of land-cover change on stream salamander
populations. Biological Conservation, 133, 436-441.
Ranta, P. (2001) Changes in urban lichen diversity after a fall in sulphur dioxide levels in the city
of Tampere, SW Finland. Annales Botanici Fennici, 38, 295-304.
Robinson, L., Newell, J. P., & Marzluff, J. M. (2005). Twenty-five years of sprawl in the Seattle
region: growth management responses and implications for conservation. Landscape and
Urban Planning, 71, 51-72.
RUS. (2012). Regional utveckling och samverkan i miljömålssystemet (Regional development
and cooperation in the environmental system). Nationella emissionsdatabasen.
http://www.rus.lst.se/nyheter.html.
16
Santesson, R., Moberg, R., Nordin, A., Tønsberg, T., & Vitikainen, O. (2004). Lichen-forming
and lichenicolous fungi of Fennoscandia. Museum of Evolution, Uppsala University,
Uppsala.
Statistics Sweden. (2011). Statistisk årsbok för Sverige 2011 (Statistical yearbook of Sweden
2011). Official statistics of Sweden, Örebro.
Statsoft, I. (2011). STATISTICA (Data Analysis Software System), Version 10.
www.statsoft.com.
Su, Z., Zhang, R., & Qiu, J. (2011). Decline in the diversity of willow trunk-dwelling weevils
(Coleoptera: Curculionoidea) as a result of urban expansion in Beijing, China. Journal of
Insect Conservation, 15, 367-377. doi:10.1007/s10841-010-9310-6
Tarhanen, S., Poikolainen, J., Holopainen, T., & Oksanen, J. (2000). Severe photobiont injuries
of lichens are strongly associated with air pollution. New Phytologist, 147, 579-590.
Tóthmérész, B., Máthé, I., Balázs, E., & Magura, T. (2011). Responses of carabid beetles to
urbanization in Transylvania (Romania). Landscape and Urban Planning, 101, 330-337.
van Herk, C. M. (2001). Bark pH and susceptibility to toxic air pollutants as independent causes
of changes in epiphytic lichen composition in space and time. The Lichenologist, 33, 419441.
van Herk, C. M., Aptroot, A., & van Dobben, H. F. (2002). Long-term monitoring in the
Netherlands suggests that lichens respond to global warming. The Lichenologist, 34(2), 141154.
Washburn, S. J., & Culley, T. M. (2006). Epiphytic macrolichens of the greater Cincinnati
metropolitan area―Part II: distribution, diversity and urban mycology. The Bryologist, 109,
516-526.
17
Fig. 1. The study area in and around Linköping, SE Sweden. The map in the middle shows the
position of the rural oaks (triangles) around Linköping (grey), while the bottom map shows
urban (circles) and rural oaks in and near Linköping (grey).
18
Fig. 2. The concentric expansion of Linköping, SE Sweden, over the past 300 years. Reproduced
with kind permission from the municipality of Linköping.
19
Fig. 3. (a)-(c) A comparison of the 14 investigated lichen species‟ occurrence and cover on oaks
in urban and rural environments. (a) shows the proportion of trees with occurrence. (b) shows the
average cover (%) on trees when present. (c) shows the average cover (%) based on all oaks.
Error bars show 95% confidence intervals (binomial ditto in 3a).
20
Fig. 4. The median of 14 target lichen species and the per cent of the trunks of Quercus robur
covered in rural and urban areas (a and c). In b and d, the cover of target lichen species on urban
oaks is shown as a function of when urbanisation occurred (based on the average age of the five
closest buildings).
21
Fig. 5. The predicted values based on results from a logistic regression of finding Calicium
viride, Chrysotrix candelaris, Evernia prunastri and Ramalina spp. on trees in relation to the age
of urbanisation.
22
Table 1The 17 target lichens searched for on Quercus robur. Nine of the lichens are common,
and eight are red-listed and/or indicator species with varying requirements of the substrate.
Lichen species
Calicium adspersum
Common
Calicium quercinum
Calicium viride
Status
Threat1
Indicator2
X
VU
X
Chaenotheca phaeocephala
Chrysothrix candelaris
Cliostomum corrugatum
X
X
NT
X
Cyphelium inquinans
X
Evernia prunastri
X
Hypocenomyce scalaris
X
Hypogymnia physodes
X
Lecanographa amylacea
Lepraria incana
X
Parmelia sulcata
X
Ramalina spp.
Schismatomma decolorans
Sclerophora coniophaea
X
Usnea spp.
X
VU
NT
NT
X
X
1
Gärdenfors (2010)
2
Nitare (2010)
3
Santesson Moberg, Nordin, Tønsberg, and Vitikainen (2004)
23
Substrate preference3
On rough bark of Quercus, rarely on
old wood.
Corticolous on Quercus, etc.,
sometimes on old wood.
Corticolous (Alnus, Quercus, Betula,
Pinus, etc.) and lignicolous.
On lignum and bark of various trees
(Quercus, etc.).
Corticolous and lignicolous.
On rough bark of Quercus, Ulmus,
etc., also lignicolous.
Lignicolous, rarely corticolous (on
Picea, Quercus and Betula) and
saxicolous.
Corticolous or sometimes lignicolous,
rarely saxicolous.
Corticolous (esp. on Pinus) and
lignicolous. Toxitolerant.
Corticolous, lignicolous and less often
saxicolous.
Corticolous on Quercus.
Corticolous and lignicolous, also
saxicolous (under overhangs).
Corticolous, lignicolous and
saxicolous, often on rocks manured by
birds.
Corticolous, esp. on old Quercus.
Corticolous on Quercus, Ulmus, etc.,
in northern areas usually lignicolous
on conifers.
-
24
Oaks circumference at breast height (cm)
Depth of bark crevices (mm)
Sun exposure (%)
Density of oaks (ha-1) within:
150 m
250 m
350 m
500 m
700 m
1000 m
Area covered by buildings (%) within:
150 m
250 m
350 m
500 m
700 m
1000 m
Urban (n=105)
Min
230
8.8
0
0.14
0.05
0.03
0.01
0.01
0.02
0.0
0.0
0.0
5.1
7.1
10.4
Average
341
24.8
36.8
0.85
0.41
0.29
0.20
0.18
0.13
35.9
39.8
40.8
40.6
38.6
35.8
88.2
78.7
77.0
71.8
75.6
66.3
3.96
1.58
1.01
0.71
0.56
0.36
Max
574
57.5
63.0
0.2
0.5
0.7
0.9
1.1
1.4
0.57
0.36
0.29
0.22
0.18
0.15
Average
351
29.2
32.6
0.0
0.0
0.0
0.0
0.0
0.0
0.14
0.05
0.03
0.01
0.01
0.00
Rural (n=109)
Min
250
12.6
4.0
15.8
24.8
28.3
35.3
30.9
19.8
2.55
1.53
1.35
1.04
0.91
0.71
Max
530
60.0
60.0
Table 2
Characterisation of the urban and rural oak populations studied, and the explanatory variables used when modelling the occurrence of
target lichens on Quercus robur in urban environments. The average depth of bark crevices is based on averages from each tree
measured from the north, east, south and west. The summarised number of oak trees (including the focal tree) and summarised density
of buildings; industrial area, high, low and enclosed buildings surrounding the focal tree on six different radius distances were
calculated from a tree database and maps, respectively.
25
Age of urbanisation
Tree circumference
Bark crevices
Sun exposure
Density of oaks within
150 m
250 m
350 m
500 m
700 m
1000 m
Area covered by buildings within
150 m
250 m
350 m
500 m
700 m
1000 m
P-value for the selected model
a
indicator lichen
0.078
+/0.014*
–/0.14
–/0.12
–/0.00076***
0.000043***
+/0.013*
Hypogymnia Chaenotheca
physodes
phaeocephalaa
–/0.13
0.00066***
–/0.055
Ramalina
spp.
–/0.013*
0.034*
–/0.039*
0.021*
–/0.044*
Parmelia Evernia
sulcata prunastri
–/0.046*
<0.00001***
+/0.099
Chrysothrix
candelaris
–/0.0043**
+/0.0014**
+/0.061
+/0.16
–/0.022*
+/0.036*
Lepraria
incana
0.00007*** 0.0037**
Calicium
viride
–/0.0033**
+/0.069
+/0.033*
–/0.017*
Table 3
The occurrence of eight lichens species on urban trees were analysed in relation to the explanatory variables shown. The P-value is
displayed for variables included in the best model for each species (based on AIC). Plus and minus signs indicate a positive or
negative association with the response variables. * indicates P-value <0.05; ** <0.01; *** <0.001.
26
Species
Cover (%)
Spores
Calicium adspersum1,3
0
Yes
Cyphelium inquinans1,3
0
Yes
Sclerophora coniophaea1
0
Yes
Usnea spp.4
0
Rarely
Cliostomum corrugatum1,3
6.66 E-08
Yes
Hypocenomyce scalaris1
5.62 E-07
Rarely
Hypogymnia physodes2
1.60 E-06
Rarely
Ramalina spp.3
2.02 E-06
Yes
Chaenotheca phaeocephala1
2.22 E-06
Yes
Parmelia sulcata2
4.75 E-06
Rarely
Evernia prunastri2
6.14 E-06
Rarely
Chrysothrix candelaris1
8.41 E-06
No
Calicium viride1,3
1.19 E-05
Yes
Lepraria incana1,3
2.89 E-05
No
P-value
0.039*
1
Foucard (2001)
2
Nash III, Ryan, Gries, and Bungartz (2002)
3
Nash III, Ryan, Diederich, Gries, and Bungartz (2004)
4
Nash III, Gries, and Bungartz (2007)
0.95
63.8
Spore area (µm2)
82.5
123
21.6
42.4
23.6
20.6
29.4
81.0
33.2
68.7
67.5
Diaspores
No
No
No
Yes
No
Yes
Yes
Yes
No
Yes
Yes
Yes
No
Yes
0.19
Pycnidium
Yes
Yes
No
No
Yes
Rarely
Yes
No
No
Rarely
Rarely
No
No
No
0.16
Photobiont
Trebouxia
Trebouxia
Trentepohlia
Trebouxia
Chlorococcus
Chlorococcus
Trebouxia
Trebouxia
Trebouxia
Trebouxia
Trebouxia
Chlorococcus
Trebouxia
Trebouxia
NA
Growth form
Crustose
Crustose
Crustose
Fruticose
Crustose
Crustose
Foliose
Fruticose
Crustose
Foliose
Fruticose
Crustose
Crustose
Crustose
0.27
Table 4
Per cent cover of 14 lichen species on urban trees as affected by six traits, i.e. if they mainly disperse by spores or diaspores, the
spores area, if pycnidium are present or not, the group to which the photobiont belong to and the growth form. The dispersal mode of
diaspores is for the included lichens soredium but with an addition in taxa Usnea spp. by isidium. * indicates P-value <0.05; ** <0.01;
*** <0.001.
27