Trait-based analysis of decline in plant species ranges during the 20th century: a regional...

Trait-based analysis of decline in plant species rangesduring the 20th century: a regional comparison betweenthe UK and EstoniaLAUR I LAAN I STO 1 , MAREK SAMMUL 2 , T I IU KULL 1 , P ETR MACEK 3 and

MICHAEL J . HUTCHINGS4

1Institute of Agricultural and Environmental Sciences, Estonian University of Life Sciences, Kreutzwaldi 5, 51014 Tartu, Estonia,2Institute of Ecology and Earth Sciences, University of Tartu, Vanemuise 46, 51014 Tartu, Estonia, 3Faculty of Science, University

of South Bohemia, Brani�sovsk�a 31, 370 05 �Cesk�e Bud�ejovice, Czech Republic, 4School of Life Sciences, University of Sussex, Falmer,

Brighton, Sussex BN1 9QG, UK

Abstract

Although the distribution ranges and abundance of many plant species have declined dramatically in recent decades,

detailed analysis of these changes and their cause have only become possible following the publication of second-

and third-generation national distribution atlases. Decline can now be compared both between species and in differ-

ent parts of species’ ranges. We extracted data from distribution atlases to compare range persistence of 736 plant

species common to both the UK and Estonia between survey periods encompassing almost the same years (1969 and

1999 in the UK and 1970 and 2004 in Estonia). We determined which traits were most closely associated with varia-

tion in species persistence, whether these were the same in each country, and the extent to which they explained dif-

ferences in persistence between the countries. Mean range size declined less in Estonia than in the UK (24.3% vs.

30.3%). One-third of species in Estonia (239) maintained >90% of their distribution range compared with one-fifth

(141) in the UK. In Estonia, 99 species lost >50% of their range compared with 127 species in the UK. Persistence was

very positively related to original range in both countries. Major differences in species persistence between the stud-

ied countries were primarily determined by biogeographic (affiliation to floristic element) and ecoevolutionary (plant

strategy) factors. In contrast, within-country persistence was most strongly determined by tolerance of anthropogenic

activities. Decline of species in the families Orchidaceae and Potamogetonaceae was significantly greater in the UK than

in Estonia. Almost all of the 736 common and native European plant species in our study are currently declining in

their range due to pressure from anthropogenic activities. Those species with low tolerance of human activity, with

biotic pollination vectors and in the families referred to above are the most vulnerable, especially where human popu-

lation density is high.

Keywords: biodiversity loss, common species, comparative analysis, conservation, plant floras, traits, vegetation change

Received 7 November 2014 and accepted 5 January 2015

Introduction

Anthropogenic activities cause irreversible change to

natural and semi-natural communities. It is well estab-

lished that activities associated with the increasing den-

sity of the human population have caused declines in

the sizes of populations of many species, in localized

species extinctions, and therefore in contractions in spe-

cies ranges’, and in the establishment of invasive aliens

(Drayton & Primack, 1996; Hooper et al., 2005; Kull &

Hutchings, 2006; Isbell & Wilsey, 2011; Chown, 2012;

Dullinger et al., 2013). Loss of biodiversity and of the

ecosystem services and benefits provided by species is

accelerating, potentially threatening the functional

integrity of communities as a consequence of increasing

anthropogenic impacts (Aguilar et al., 2006; Brook et al.,

2008; Isbell et al., 2013).

Factors causing declines in range and abundance

include habitat loss and fragmentation. The smaller

sizes and greater isolation of populations following

such disruptions place them at greater risk of further

decline and local extinction (Joshi et al., 2006; Laanisto

et al., 2013). In addition, the impact of invasive species

(Powell et al., 2011), the decrease in numbers, or total

loss, of pollinating species (Aguilar et al., 2006; Albrecht

et al., 2012) and other plant symbionts (Wagg et al.,

2011), soil degradation (Verbruggen et al., 2010) and

many other factors have been shown to be responsible

for local extinctions of plant populations. Furthermore,

changes in land use and climate have altered the loca-

tions at which species can find both optimum andCorrespondence: Lauri Laanisto, tel. +372 55636784, fax

+372 731 3988, e-mail: [email protected]

1© 2015 John Wiley & Sons Ltd

Global Change Biology (2015), doi: 10.1111/gcb.12887

acceptable conditions for their continued existence

(Brook et al., 2008), leading to formerly occupied sites

becoming unsuitable, and necessitating the colonization

of new locations.

Whereas much biogeographical research has

focused on predicting changes in the overall distribu-

tion ranges and range boundaries of species and

communities (Heikkinen et al., 2006; Peterson, 2011),

research in conservation biology has often concen-

trated more on small scale changes, such as those

occurring at the scale of single dots on distribution

maps (Kujala et al., 2011). These are often records of

species’ occurrences at scales of anything from 1 to

10 km2. To date, it has been difficult to connect

knowledge between these two approaches due to dif-

ferences in spatial scale (Guisan et al., 2013) and lack

of entailed biological information (Travis et al., 2013),

but the recent publication of second- and even third-

generation national atlases of species distribution

maps (Tamis et al., 2005) now enables analyses of

changes in species’ ranges between specified dates,

and comparison of changes in range within regions

or national territories for groups of species catego-

rized according to their evolutionary history, ecologi-

cal preferences and morphological, physiological and

reproductive characteristics. To date, most analyses

of changes in distribution ranges over time have

focused on single countries or regions (e.g. the Neth-

erlands (Tamis et al., 2005), Northamptonshire in the

UK (McCollin et al., 2000), Thi�erache in France (Van

Calster et al., 2008) and Flanders in Belgium (Van

Landuyt et al., 2008). To our knowledge, apart from a

comparison of range decline in the orchid species

common to Estonia and the United Kingdom (Kull &

Hutchings, 2006), only one comparative analysis of

changes in species’ distributions has been carried out

involving species common to the floras of more than

one region or country (Powney et al., 2014).

Although understanding of the limitations on species

distributions is poor (Peterson, 2011; Ara�ujo et al., 2013;

Guisan et al., 2013), much is known about the evolu-

tionary background of species, their ecology and their

tolerance of anthropogenic influences. Trait information

is, however, rarely taken into account when analyzing

species distributions and their changes from a conser-

vation point of view (van Kleunen et al., 2010; Wiens

et al., 2010; Chown, 2012; Saar et al., 2012). Although

accurate knowledge of the locations of populations has

a significant role to play in informing conservation

practices, forecasting the future dynamics of those pop-

ulations from presence/absence data alone might not

be reliable. The addition of trait-based information

might significantly improve predictions of future

behaviour (Peterson, 2011; Chown, 2012).

This study took advantage of the availability of sec-

ond-generation national plant distribution atlases for

the United Kingdom (hereafter the UK) and Estonia to

examine the relationships between a wide range of spe-

cies characteristics and change in species ranges. Our

purpose was to identify traits and factors associated

with change in distribution ranges and to determine

whether their influences have been similar in these two

countries. The national floras of the UK and Estonia

provide a unique opportunity to gain ecological

insights into the factors underlying species range

changes for the following reasons. Firstly, the national

floras have a large number of species (and a high pro-

portion of their total species) in common, enabling

direct comparison of the proportional changes of many

species ranges in both countries. Secondly, there have

been significant differences in the pressures to which

natural and semi-natural vegetation has been exposed

in the two countries. In particular, the UK has a very

high mean human population density, and the infra-

structure supporting this population has resulted in

considerable habitat fragmentation and isolation of the

remaining patches of natural and semi-natural vegeta-

tion, many of which are suffering considerable biodi-

versity loss (Thomas et al., 2004; Walker et al., 2009). In

contrast, Estonia has one of the lowest mean human

population densities in Europe, and, after the collapse

of the USSR in 1991, agriculture and forestry became

less intensive and the rate of wetland drainage declined

(Kull & Hutchings, 2006; Kurganova et al., 2014). There

was, for example, a 20-fold increase in the area of

unused former arable land between 1991 and 1995 (Pet-

erson & Aunap, 1998) (see Table 1 for further compari-

son of the UK and Estonia). Species range declines and

loss of biodiversity to date in Estonia might therefore

be expected to be less than in the UK. It is already

known that mean percentage decline of the distribution

range of orchid species in the UK during the second

half of the 20th century was twice as high (49.7%) as for

orchid species in Estonia (25.0%; Kull & Hutchings,

2006). The difference in decline for the 30 orchid species

common to both countries was also significant (mean

for the UK = 52.0%, mean for Estonia = 28.5%) (Kull &

Hutchings, 2006).

Species distributions depend on a wide variety of fac-

tors, ranging from life-history traits to their responses

to various abiotic environmental characteristics

(Chown, 2012). In social sciences, the standard method

of projecting population dynamics (e.g. population pro-

jections made by national census bureaus, http://

www.census.gov/population/projections/) is based on

analysis of a wide variety of information about the indi-

viduals in the population, including age, sex, ethnicity,

income and education. A similar approach is often used

© 2015 John Wiley & Sons Ltd, Global Change Biology, doi: 10.1111/gcb.12887

2 L. LAANISTO et al.

http://www.census.gov/population/projections/

http://www.census.gov/population/projections/

in evolutionary population and community ecology,

where evolutionary stable strategies or stable stage dis-

tributions are calculated based on the most general

traits describing the sample (Williams et al., 2011). As

temporal population projections are already available,

from successive distribution atlases, from two contrast-

ing countries for a large number of the same species,

we can apply the same method to investigate the main

factors behind the changes in distribution of those spe-

cies. For this purpose, a range of plant traits was

selected for use in analyses. These were the evolution-

ary age and origin of the species, their average size,

their tolerances towards potential environmental stres-

sors, the extent to which they are dependent upon other

species (e.g. pollinators), their reproductive strategy

and their tolerance of anthropogenic influences. Com-

parative analysis of such traits, for all the species in a

single model (for details of traits and related hypothesis

see Methods), allowed us to determine the main traits

and domains of traits associated with variation in spe-

cies persistence. We use the term ‘domain’ to signify

groups of traits that were considered likely to be associ-

ated with range decline.

Our study had three main goals: (1) to analyze and

compare species range changes during the second half

of the 20th century in the UK and Estonia. Besides doc-

umenting the overall changes, we examined whether

changes in persistence, both for the whole flora and for

individual species, were comparable between countries.

(2) To determine the traits and domains of traits that

are most closely associated with differences in species

persistence and whether these are the same in both

countries. (3) To analyse and compare persistence at

the level of plant families between the countries. We

describe the traits and domains of traits selected for

analysis, together with the specific hypotheses exam-

ined in connection with each of them in the Methods

section below.

Materials and methods

The United Kingdom and Estonia are both situated in north-

ern Europe. The UK is about 30 degrees west of Estonia, but

there is latitudinal overlap between the two countries, with

Estonia situated at similar latitude to northern Scotland and

the Orkney Islands. The land area of the UK is more than five

times that of Estonia, and its mean human population density

is more than twenty times as high (Table 1). About 3000 vas-

cular plant species have been recorded in the UK and Ireland

(Preston et al., 2002). The number of native species has been

reported to be 1254 (Thomas et al., 2004), 1446 (Pilgrim et al.,

2004) and 1515 (Crawley et al., 1996). The Estonian flora con-

sists of 1538 native species (Kukk, 1999; Kull et al., 2002) plus

880 alien species (€O€opik et al., 2013).

The databases used in the study were the New Atlas of the

British and Irish Flora (Preston et al., 2002) and the Atlas of

the Estonian Flora (Kukk & Kull, 2005). Note that data for the

Republic of Ireland were excluded; only the data for the UK

were used in all the analysis. In both atlases, the presence of

species was recorded on maps in 100 km2 grid squares. The

two survey periods for which distribution maps were pre-

pared were 1930–1969, and 1987–1999 in the UK and 1921–1970 and 1971–2004 in Estonia. Thus, there was an interval of

30 and 34 years between the ends of the first and second sur-

vey periods in the UK and Estonia, respectively, and the ends

of the two survey periods corresponded closely in both coun-

tries. For each species in each country, persistence in distribu-

tion range was calculated as the percentage increase or

decrease in the number of grid squares occupied by the end of

the second survey period compared with the number of grid

squares occupied by the end of the first survey period.

The available data do not include information on the abun-

dance of species in each grid square. For practical purposes, in

large-scale comparative analyses of changes in distribution

ranges such as this study, it has to be assumed that popula-

tions of all species are equally detectable in all habitats and in

all grid squares (K�ery, 2004). The number of species available

(i.e. species for which there were distribution maps for both

survey periods) from the initial data set was 1411 species in

the UK and 1115 species in Estonia. The number of species

common to both the UK and Estonia, for which data on

changes in distribution range were available, was 736 (see

Appendix S1 for species list). Pairwise comparisons of changes

in distribution range in the two countries were carried out for

all of these species.

Persistence, and differences in persistence between the two

countries, calculated as persistence in Estonia minus persis-

tence of the same species in the UK, were analysed using Type

III analysis of covariance (ANCOVA) in the GLM procedure in

STATISTICA 8.0 (StatSoft, Inc., Tulsa, OK, USA). We used this

Table 1 Comparison of relevant characteristics of the UK

and Estonia (UK data from Scott & Jones, 1995; Estonian land

use data from €O€opik et al., 2008). Climate data represent long-

term averages

UK Estonia

Area (km2) 243 610 45 227

Population (millions) 63.18 1.29

Mean density of human

population (km�2)

662 29

Land use (% of total area)

Agriculture 80 29

Forest 7 50

Wetland 2 13

Urban 11 8

Latitudinal range °N 49–60 57–59Annual precipitation (mm) 850 650

Nr of rainy days per year 133 115

Annual temperature °C 9.7 5.2

Temp. of warmest month °C 16.3 16.4

Temp. of coldest month °C 4.1 �5.7


PLANT SPECIES DECLINE DURING 20TH CENTURY 3

approach because of analytical transparency and lack of bias,

and because the results produced can be readily compre-

hended by the widest possible audience. To examine persis-

tence in plant families, we used two-way ANOVA Type III, with

family and country as factors, and we performed post hoc com-

parisons (Fisher LSD test).

To relate species persistences in Estonia and UK to species

traits and variables, we used the redundancy analysis proce-

dure (RDA) of multivariate analysis. RDA is an extension of

multivariate linear regression for a multivariate response vari-

able (Lep�s & �Smilauer, 2003), with the parametric test replaced

by a Monte Carlo permutation test to overcome problems with

distributional characteristics. RDA enables visualization of the

main trends in the data. Variables and trait states (for qualita-

tive variables) that contributed significantly to explaining per-

sistence were selected by forward selection (based on 499

permutations). The RDA analysis and graphical presentation

of the results were carried out using the Canoco 4.56 and

CanoDraw programs (Ter-Braak & �Smilauer, 2002).

Our analysis included a wide variety of variables, ranging

from historical factors, such as species origin, to functional

traits such as plant height. We assigned each variable into one

of several larger and more comprehensive domains related to

evolutionary history, adaptation, ecological preference and

tolerance of anthropogenic activities and disturbance, each of

which might play an important role in determining species

persistence. In addition, the analysis of larger sets of intercon-

nected variables enables determination of the key factors, both

within and between the domains, having the most direct effect

on species persistence. This approach also allows the domains

causing the greatest changes in distribution, both within and

between countries, to be identified. Appendix S2 gives details

of each variable, the qualitative or quantitative categories into

which each variable was divided, and the number of species

that fell within each category. The domains of traits, with

descriptions of the individual traits included, their relevance

to the analysis of species distribution change and related

hypotheses, were as follows:

Evolutionary history domain: Species age and origin

Floristic element. The geographic origin of the species, based

on Hult�en’s distribution atlas (Hult�en, 1971). Data were

extracted from the List of Estonian Vascular Plants (Kukk,

1999: summary in English: http://www.zbi.ee/~tomkukk/ni-

mestik/english.htm).

Age. The evolutionary age (in millions of years) of the crown

group of the family. Data were extracted from the Angiosperm

Phylogeny Webpage (http://www.mobot.org/MOBOT/

research/APweb/) and Pryer et al. (2004).

The information included in this domain is essentially his-

torical in nature; age and geographic origin of species can be

considered the basis for their ecological preferences (P€artel

et al., 2007; Ara�ujo et al., 2013; Marske et al., 2013). Niche con-

servatism– that is the degree to which taxa retain their ances-

tral ecological traits and geographical distributions– can

influence their current habitat and environmental preferences

(Wiens & Graham, 2005; Wiens et al., 2010; Crisp & Cook,

2012), and significantly affect plant species’ success at a local

scale, whether by limiting their capacity to reproduce (Laanis-

to et al., 2008), the ranges of species-specific pollen vectors

(McLeish et al., 2011), or some other functional property.

Species age and origin do not in themselves provide infor-

mation about species that will influence their persistence over

ecological timescales. These factors, especially age, would be

expected to have little influence on species persistence in

either country. On the other hand, climate change may not yet

have affected persistence significantly, but the evolutionary

origin of species and their persistence might be related. For

example, it has been shown that the ranges of species from

some floristic elements, such as arctic-alpine species, have

been in decline for some time (Lesica & McCune, 2004); we

therefore hypothesized that the time interval between the end

of the two surveys in each country might reveal changes in

persistence related to the floristic element to which species

belong.

Adaptation domain: Qualitative life-history traits

Strategy. Plant strategies based on their tolerance of stress

and disturbance, and their ruderality. Species were classified

into three main strategy types – competitors (C), stress tolera-

tors (S) and ruderals (R) and four intermediate strategies (CS,

CR, SR and CSR). Data were extracted from Grime et al.

(2007).

Diaspore type. Generative diaspores (units of dispersal) may

be seeds or seeds either embedded in additional structures or

with additional structures attached. Data to describe diaspore

type were obtained from the BiolFlor database (http://

www2.ufz.de/biolflor/overview/merkmale.jsp).

Pollen vector. Pollen vector or mode of pollen transfer. Data

were obtained from the BiolFlor database (http://

www2.ufz.de/biolflor/overview/merkmale.jsp) (Mode of fer-

tilization, as opposed to mode of pollen transfer, is considered

within Ecological preferences domain).

A key requirement for understanding the processes behind

vegetation change is the availability of functional autecologi-

cal data about the species involved (Hodgson et al., 1999).

Qualitative life-history traits and trade-offs, as reflected in

CSR plant strategy categories, are useful reflections of species

fundamental niches (Tilman, 1994; Kneitel & Chase, 2004;

Hubbell, 2005; de Bello et al., 2013), and determine, in broad

terms, species0 competition strategy.

All factors in this domain are qualitative in nature and they

are essentially trade-offs. Plant strategy reflects behaviour in

the established phase of the life history of the species and is

based on a trade-off between tolerance of stress and tolerance

of disturbance (Hodgson et al., 1999; Grime et al., 2007). Selec-

tion of strategy is based on long-term presence in, and adapta-

tion to, particular habitat types. Plant strategies can therefore

indicate whether local vegetation composition conforms with

expectation under current land use and climatic conditions,

and enable us to predict changes in vegetation if conditions



http://www.zbi.ee/~tomkukk/nimestik/english.htm

http://www.zbi.ee/~tomkukk/nimestik/english.htm

http://www.mobot.org/MOBOT/research/APweb/

http://www.mobot.org/MOBOT/research/APweb/

http://www2.ufz.de/biolflor/overview/merkmale.jsp




alter (Lep�s et al., 1982; Hodgson et al., 1999; Vicente et al.,

2013). We hypothesized that fragmentation and destabiliza-

tion of habitats due to anthropogenic activities will be

reflected in different plant strategies influencing species per-

sistence, especially in the UK, where human density is far

greater than in Estonia.

Similarly, diaspore type and pollen vector are determined

by the long-term dynamics of the habitats species occupy and

can provide information on traits that have been successful in

promoting persistence during the interval between the two

surveys in each country. For example, populations of plant

species that use a limited range of animal species for pollina-

tion and seed dispersal could be at greater danger of extinc-

tion as a result of habitat fragmentation and disturbance, than

species that use wind to disperse their pollen and diaspores

(Schleicher et al., 2011). On the assumption that disturbance

due to human impact has been greater in the UK than in Esto-

nia due to contrasting human population densities, we

hypothesized that both diaspore type and pollen vector will

have had a greater impact on persistence in the UK than in

Estonia.

Ecological preferences domain: Quantitative and plastictraits

Height. Mean heights of species (cm) were obtained from the

Flora of Estonia (Flora of Estonia SSR 1953–1984). The height

of trees (38 species) and aquatic plants (37 species) was not

included in this analysis because the inclusion of very tall spe-

cies and floating species raised variance to an unacceptable

level in the analysis.

Flowering phenology. This variable records the start, end and

duration of the flowering period of species (in months). Data



Type of reproduction. Reproduction can be by seed, by vege-

tative propagation, or include both of these possibilities. Data



Plasticity. The capacity to alter phenotype in response to

environmental conditions (e.g. being able to alter height, flow-

ering for different periods and propagate either generatively

or vegetatively) has been shown to have a significant positive

impact on species persistence in the face of fluctuating envi-

ronmental conditions (Lande, 2009), increasing human impact

(Crispo et al., 2010) and changing climate (Nicotra et al., 2010).

Plant height is considered one of the strongest predictors of

fitness (Falster & Westoby, 2003; Nicotra et al., 2010). It is one

of the most plastic traits in response to changes in habitat con-

ditions (Pan et al., 2013). Alteration of height can change many

aspects of plant behaviour, including the balance between sex-

ual and vegetative propagation, water use and light harvest-

ing efficiency. It can also determine successional status

(Falster & Westoby, 2003).

Climate change is strongly affecting plant phenology pat-

terns worldwide (Anderson et al., 2012; Wolkovich et al.,

2012). Shifts in the phenology of flowering caused by climate

change can benefit species with long flowering periods,

whereas species with shorter flowering periods may suffer,

for example by losing synchrony with the critical periods of

activity of their pollinators.

Type of reproduction affects local persistence of species in

various ways, for example by altering competition (Zobel,

2008). While the ability to propagate vegetatively can sustain

populations through adverse conditions, or allow species to

re-establish quickly after severe disturbance, such effects of

clonality normally manifest themselves only at a local scale

(Laanisto et al., 2008). Wind-pollinated species usually have

wider dispersal potential and are less pollen-limited than spe-

cies that are pollinated by animals, yet their pollination suc-

cess rate depends more on plant size (Friedman & Barrett,

2009). Self-pollination can be a very effective strategy over a

short time frame, but, like vegetative propagation, it is inferior

to sexual reproduction for genetic recombination, wide dis-

persal and resistance to parasites and viruses (Barrett, 1998).

However, if the abiotic environment changes, self-pollination

rates in plant communities tend to increase (Jones et al., 2013).

We hypothesized that these quantitative traits would play

an important role in determining changes in species distribu-

tion ranges in both the UK and Estonia during the period

under investigation, because species with greater phenotypic

plasticity are expected to be better competitors in unstable and

fragmented habitats (Callaway et al., 2003; Laanisto et al.,

2008; Anderson et al., 2012).

Human influence domain

Urbanity. Urbanity reflects the ability of plant species to per-

sist in urban areas. Species are categorized from urbanophile

(species recorded mainly in cities) to urbanophobe (species

recorded rarely in cities). Data were obtained from the Biol-

Flor database (http://www2.ufz.de/biolflor/overview/merk-

male.jsp).

Hemeroby. Hemeroby is a measure of departure from natu-

ralness of the habitats in which a species is recorded. Habitats

and vegetation types are classified on a scale from ahemerobe

(species occurring exclusively in natural habitats) to poly-

hemerobe (species occurring exclusively in non-natural habi-

tats). Data were obtained from the BiolFlor database (http://


Nitrogen requirement. Ellenberg indicator value for species’

nitrogen requirement was obtained from Ellenberg et al.

(1992).

The ability to tolerate pollution, land use changes, habitat

disturbance and fragmentation, and other stresses resulting

from human activities, has become increasingly important for

plant and other species (Garay-Narvaez et al., 2013). The for-

mation of novel ecosystems and communities with altered soil

conditions and invasive biota influences biodiversity and spe-











cies composition at a variety of spatial scales (Tamis et al.,

2005).

Some species tolerate proximity to human populations bet-

ter than others (Bernhardt-Romermann et al., 2011), and such

species are also more likely to be successful aliens (Dawson

et al., 2012). Generating a suitable measure of species’ toler-

ance of anthropogenic influences is difficult (Hill et al., 2002),

and therefore, we also used an indirect indicator. The last gla-

cial period destroyed almost all the high-nutrient habitats

from northern parts of the northern hemisphere (Graham

et al., 2003), including the UK and Estonia, and this caused the

extinction of many species with high-nutrient requirements

(Birks & Birks, 2004; P€artel et al., 2007). Current human activi-

ties both generate nutrient-rich novel ecosystems and enrich

existing habitats with nutrients (Hoover et al., 2012), so that

plant that are adapted to grow on substrates with higher nutri-

ent content are favoured, and plants with lower nutrient pref-

erences are more prone to local extinction (Saar et al., 2012;

Powney et al., 2014). Among the essential plant macronutri-

ents, a low availability of nitrogen is generally associated with

high plant diversity (Bobbink et al., 2010).

We hypothesized that all three traits in this domain would

play significant roles in determining local plant population

persistence both in the UK and Estonia. As the traits influence

vegetation primarily at a local level, we also hypothesized that

they would have a similar impact on species persistence in the

two countries.

Results

Mean decline of the 736 species common to both the

UK and Estonia during the period between the two sur-

veys (approximately 30 years in both countries) was

significantly lower (t = 5.2874; P < 0.0001) in Estonia

than in the UK: average persistence (�SE) was

75.7 � 0.008% for species in Estonia compared to

69.7 � 0.008% in the UK.

At the end of the second survey, nearly one-third of

the 736 species in Estonia (239 species) were still

recorded in at least 90% as many grid squares as they

had been recorded in at the end of the first survey,

whereas less than one-fifth of the species in the UK

(141) had such high persistence (Fig. 1). In Estonia, 99

species (13.5% of the 736 species) had lost more than

half of their local populations compared with 127

(17.3%) in the UK (Fig. 1); 13 species had apparently

become nationally extinct in Estonia; and 5 had suf-

fered the same fate in the UK. Comparing species per-

sistence between the countries showed that 258 species

had higher persistence in the UK, while 478 species had

higher persistence in Estonia (Fig. 2). Persistence in

both countries was strongly positively correlated with

the number of grid squares occupied at the end of the

first survey (Fig. 3).

Although we had predicted that species age and ori-

gin (the variables within the Evolutionary history

domain) would not play a significant role in determin-

ing persistence, especially at a local (i.e. within coun-

tries) scale, affiliation to floristic element was a

significant factor both at local and regional (i.e. between

the two countries) scales (Tables 2–4).Of the variables in the Adaptation domain, plant

strategy had a significant effect on both within- and

between-country differences in persistence (Tables 2–4). Contrary to our prediction, neither diaspore type

2%

0%

3%

6% 6%

11%

16%

21%

32%

2%3%

5%

7%

12%

14%

16%

20%19%

0.1 0.2 0.3 0.4 0.5 0.6 0.7 0.8 0.9

Persistence

20

40

60

80

100

120

140

160

180

200

220

240

260

No

of s

peci

es

2%

0%

3%

6% 6%

11%

16%

21%

32%

2%3%

5%

7%

12%

14%

16%

20%19%

Estonia the UK

Fig. 1 Frequency histograms to compare persistence in Estonia and the UK of 736 plant species that are common to both countries.



nor pollen vector had a significant impact on persis-

tence in any analysis. Neither the set of variables associ-

ated with plasticity in the ecological preferences

domain nor those reflecting the sensitivity of species to

human impact in the human influence domain were

associated with differences in persistence between the

two countries (Table 4). Of the individual variables

within these domains, plant height had a significant

impact on persistence in Estonia (Table 2), but not in

the UK. Hemeroby and urbanity had significant effects

on persistence both in Estonia (Table 2) and in the UK

(Table 3). The Ellenberg number for nitrogen was not

significantly associated with persistence in any analy-

sis.

Ten variables and trait states were selected by RDA

as predictors of species persistence (Fig. 4). A substan-

tial proportion of the variability remained unexplained

by the variables we selected for study. The first axis

explained 15.4% of the variation in the data (F = 131.6,

P = 0.002), while the second axis explained an addi-

tional 4.2% of the variation. The first axis was associ-

ated with variables related to persistence as reflected

by the complete set of data, whereas the second axis

revealed variables associated with differences in persis-

tence between the two countries. Persistence in both

countries was positively related to urbanity, with more

persistent species tending to be C-strategists, or to have

intermediate strategies involving competitiveness. Tal-

ler species also displayed greater persistence, especially

in Estonia. Differences in persistence between the coun-

tries were linked to hemeroby, ruderality, pollination

type and floristic origin. Species associated with dis-

turbed habitat persisted better in the UK than in Esto-

nia. Species with pollination syndromes not involving

other species (i.e. species with wind, water or self-polli-

nation) had higher persistence in the UK, whereas

insect-pollinated species had higher persistence in Esto-

nia. Species of European origin were more persistent in

the UK, whereas species of Eurasian origin were more

persistent in Estonia.

Most pairwise comparisons of persistence made at a

family level showed similar patterns in both countries

(Fig. 5). In most families with 15 or more species rep-

resentatives in both countries, persistence was close to

the overall mean persistence for all species. However,

species in the Orchidaceae and Potamogetonaceae had

0.0 0.2 0.4 0.6 0.8 1.0

EST persistence

0.0

0.2

0.4

0.6

0.8

1.0

UK

per

sist

ence

____ linear regression- - - 1:1 line

Fig. 2 Scatterplot of persistence in the UK vs. persistence in Estonia for 736 plant species that are common to both countries. In addi-

tion to the linear regression line through the data, a 1 : 1 line is drawn as a reference to indicate equal persistence in both countries.

Points above and below the line indicate greater persistence in the UK (258 species) and in Estonia (478 species), respectively.



significantly greater mean persistence in Estonia com-

pared to the persistence in the UK (Fig. 5).

Discussion

Our results confirm widespread net reduction in plant

species distribution ranges. Of the 736 species common

to both countries on which our analyses are based,

there was on average nearly 25% loss of national range

in Estonia and more than 30% loss in the UK during

approximately 30 years. The distribution range of all

other species declined in both countries (Figs 1 and 2).

The latter result is of particular concern because our

data set comprises the core of the native flora of both

countries. The species represented have wide regional

distributions (the distance between the UK and Estonia

is roughly 2000 km). While several previous studies

(e.g. Lavergne et al., 2005; Kull & Hutchings, 2006) have

shown that rare species, including many for which

conservation management plans have been developed,

have lost territorial range, this study demonstrates that

the ranges of more common, regionally widespread

plant species are suffering from the changes in climate

and human impact that have taken place during the

20th century.

On average, species with small national distribution

ranges had higher persistence in Estonia than in the UK

(Fig. 3). While a considerable number of species had

0 100 200 300 400 500

0.0

0.2

0.4

0.6

0.8

1.0

Per

sist

ence

P < 0.0001; r2 = 0.3901

P < 0.0001; r2 = 0.5324

Estonia

0 500 1000 1500 2000 2500Number of 100 sq km grids occupied at the end of the first survey period

0.0

0.2

0.4

0.6

0.8

1.0

Per

sist

ence

the UK

(a)

(b)

Fig. 3 Species persistence plotted against the number of

100 km2 grid squares occupied at the end of the first survey per-

iod. (a) Shows persistence of 736 species in Estonia (maximum

initial distribution range was 510 grid squares); (b) shows per-

sistence of the same 736 species in the UK (maximum initial dis-

tribution range was 2821 grid squares).

Table 2 Univariate test of significance of variables for deter-

mination of persistence of Estonian plant species. Significant

factors are in bold

SS MS F P

Evolutionary history domain

Age 0.055 0.055 1.311 0.253

Floristic element 0.533 0.107 2.563 0.026

Adaptation domain

Strategy 2.311 0.083 1.982 0.002

Diaspore type 0.014 0.014 0.343 0.558

Pollen vector 0.006 0.006 0.148 0.700

Ecological preferences domain

Type of reproduction 0.048 0.048 1.153 0.283

Height 0.174 0.174 4.168 0.042

Duration of flowering 0.055 0.055 1.333 0.249


N (Ellenberg) 0.029 0.0289 0.693 0.406

Hemeroby 0.309 0.309 7.428 0.007

Urbanity 0.549 0.549 13.17 0.000

Error 25.272 0.0416

Table 3 Univariate test of significance of variables for deter-

mination of persistence of United Kingdom plant species. Sig-

nificant factors are in bold

SS MS F P


Age 0.014 0.014 0.469 0.494


Adaptation domain

Strategy 7.793 0.278 9.647 0.000

Diaspore type 0.040 0.040 1.382 0.240

Pollen vector 0.069 0.069 2.387 0.123



Height 0.004 0.004 0.139 0.709



N (Ellenberg) 0.002 0.002 0.063 0.802

Hemeroby 0.336 0.336 11.650 0.001

Urbanity 0.719 0.719 24.936 0.000

Error 17.484 0.029



high persistence in Estonia, irrespective of the number

of grid cells they occupied at the end of the first survey

period (Fig. 3 upper graph), examples of such species

were virtually absent in the UK (Fig. 3 lower graph).

This difference between the countries suggests that

suitable habitats for populations of species to survive in

are more scarce in the UK and that human influence on

vegetation composition is more direct in the UK than in

Estonia.

Contrary to our expectations, one of the variables in

the Evolutionary history domain – namely species affil-

iation to floristic element – played a significant role in

explaining differences in persistence both within and

between the countries (Tables 2–4). Significance of the

effect of floristic affiliation confirms the importance of

large-scale processes in affecting vegetation (P€artel

et al., 2007). The often intangible processes operating

across large spatial and temporal scales can provide

valuable information for predicting the effects of global

change on vegetation and enable a distinction to be

made between anthropogenic and natural causes of

changes in local and regional distributions (Huston,

1999). In agreement with results obtained by Lesica &

McCune (2004), our data showed that species from the

arctic-alpine floristic element had lost the highest pro-

portion of their range in both countries (mean persis-

tence for both the UK and Estonia was only ~62%).

RDA analysis further indicated that whereas species

with Eurasian origin have greater persistence in Esto-

nia, species originating from Europe have greater per-

sistence in the UK (Fig. 4). This may be because the

range limits of Eurasian species are more commonly

reached in the UK, whereas those of European species

are more often reached in Estonia.

The effects of functional traits, both qualitative and

quantitative (i.e. the Adaptation and ecological prefer-

ences domain), on persistence agreed with our expecta-

tions (Tables 2–4; Fig. 4). In recent years, functional

Table 4 Univariate test of significance for pairwise differ-

ences in persistence of Estonian and United Kingdom plant

species. Significant factors are in bold

SS MS F P


Age 0.013 0.013 0.221 0.638


Adaptation domain

Strategy 4.151 0.148 2.616 0.000

Diaspore type 0.006 0.006 0.110 0.740

Pollen vector 0.125 0.125 2.204 0.138



Height 0.177 0.177 3.120 0.078



N (Ellenberg) 0.044 0.044 0.785 0.376

Hemeroby 0.001 0.001 0.012 0.916

Urbanity 0.012 0.012 0.220 0.640

Error 34.346 0.057

Fig. 4 Results of the redundancy analysis, showing the relation-

ships between variables (dotted lines), trait states (triangles)

and local and regional persistence patterns (solid lines). Explan-

atory variables were selected using forward selection (499 per-

mutations), and only the ten best predictors were selected. The

first and second ordination axes explained 14.1% and 4.3% of

the total variation, respectively (F = 119.6, P = 0.002).

Fig. 5 Variability plot of differences in persistence of species in

the UK (open symbols) and Estonia (filled symbols) categorized

by family. Only families with 15 or more representative species

both in the UK and in Estonia were included. Error bars denote

95% confidence interval. Significance of pairwise comparisons

of persistence between the two countries is indicated by aster-

isks (*P ≤ 0.1; ***P ≤ 0.01).



trait diversity has been strongly linked with plant dis-

tribution and is regarded as a key factor regulating

diversity dynamics (Suding et al., 2005; Flynn et al.,

2011). Decline in the abundance of specialist species

and expansion of the ranges of generalists has been

described as the most crucial reason for loss of plant

and animal diversity in terrestrial and marine ecosys-

tems (Helm et al., 2009; Clavel et al., 2011; Potts et al.,

2010; Ozinga et al., 2013). While plant height was posi-

tively related to persistence in Estonia (Table 2; Fig. 4),

species with longer flowering periods displayed greater

persistence in the UK (Fig. 4). Moreover, whereas

insect-pollinated species had higher persistence in Esto-

nia, species with abiotic pollination had higher persis-

tence in the UK (Fig. 4). These results reflect clear

differences in the impact of anthropogenic influences

between the two countries. In Estonia, taller species

have higher persistence. Such species are more success-

ful at attracting pollinators in fallow habitat (Jane�cek

et al., 2013; Otsus et al., 2014), the abundance of which

increased dramatically after the collapse of the USSR

(Peterson & Aunap, 1998). In the far more densely pop-

ulated UK (see Table 1), the relative scarcity of pollina-

tor species is reflected in species with abiotic

pollination, and those with long flowering periods that

maximize the probability of insect pollination, having

greater persistence (Pickering & Hill, 2002; Potts et al.,

2010).

There were significant effects of plant strategy on

persistence, both within and between countries

(Tables 2–4; Fig. 4). Although some studies have

shown that plant CSR strategy types (sensu Grime et al.,

2007) carry signals of niche conservatism (Knapp et al.,

2008), relationships between phylogeny and strategy

are rather weak (Pavoine & Ricotta, 2013). As predicted,

the RDA model showed that both stress tolerators and

ruderals had higher persistence in the UK, where

anthropogenic disturbance is higher, than in Estonia. In

addition, the ability to propagate vegetatively did not

influence persistence, even though clonal species are

known to have higher survival in disturbed and anthro-

pogenically influenced habitats (Fahrig et al., 1994; Saar

et al., 2012).

The variables hemeroby and urbanity in the human

influence domain significantly influenced persistence in

the UK and Estonia (Tables 2–4). The RDA model indi-

cates that hemeroby is linked with differences in persis-

tence between the two countries, while urbanity is

positively related to persistence in both the UK and

Estonia (Fig. 4). As predicted, species that are intolerant

of human influence are less persistent, especially in the

UK. The contrasting effects of hemeroby between the

countries are probably caused by differences in

persistence in two families, Orchidaceae and

Potamogetonaceae, with large numbers of species com-

mon to both countries (Fig. 4). Species in both families

have very low values of hemeroby.

Our data showed that the evolutionary age of lineage

of descent is not only negatively related to species

hemeroby (r = �0.148; P < 0.0001) and urbanity

(r = �0.165; P < 0.0001) indices, but also negatively

related to Ellenberg nitrogen number (r = �0.084;

P = 0.024). Despite this, evolutionary age itself was not

a significant explanatory variable in any analyses,

implying that while species belonging to younger lin-

eages tolerate human activities better than those of

older lineages, the relatively recent increase in anthro-

pogenic activities is yet to have a detectable influence

on plant species persistence at a phylogenetic level.

Both the UK and Estonia have ancient traditions of

herding domestic animals in pastures, and this type of

historical human influence, and the historical connec-

tivity between pastures, has had a lasting positive effect

on local plant species diversity (Helm et al., 2006). Even

when traditional land use has changed and the habitats

associated with these activities have declined in abun-

dance, local species extinctions are often not seen for a

considerable time. For example, the payment of extinc-

tion debts in temperate habitats can take anything from

50 to 100 years or more (Vellend et al., 2006; Kuussaari

et al., 2009). Despite current conservation measures and

subsidy of traditional land use methods, the general

trend of local population extinction might be irrevers-

ible, and communities may become more and more

species poor and similar (Clavel et al., 2011), both taxo-

nomically and functionally. Given that plant species

extinction debt has been clearly demonstrated in

regions where landscape changes are taking place

(K€orner & Jeltsch, 2008), the decline of species ranges

reported in this study may be only the start of a process

that has a long way to run before completion.

In addition to the roles of human and climatic

impacts, the fact that almost all the species examined in

this study are declining in range can be partly

explained by the fact that only native and archaeophyte

species were included in the sample. The influx of

many alien species into Europe has taken place only

since the 1970s (Py�sek et al., 2009). Consequently, we

were unable to analyse the effect of the arrival and

range expansion of alien species on the persistence of

common native species in the UK and Estonia. Other

studies have shown that the effect of invading alien

species on the range of native species can be highly

negative, especially because traits such as size, fitness

and growth rate of invasive species tend to be signifi-

cantly higher than those of native species (van Kleunen

et al., 2010). Nevertheless, even the distribution range

of Elodea canadensis, an invasive aquatic species that



qualified for inclusion in our data set, has declined by

13% in Estonia and by 20% in the UK over the studied

period.

For the Orchidaceae (cf. Kull & Hutchings, 2006) and

Potamogetonaceae, mean local extinction was signifi-

cantly higher in the UK than in Estonia (P ≤ 0.01); simi-

lar trends for Apiaceae, Asteraceae, Caryophyllaceae and

Fabaceae were significant at a greater P-level (P ≤ 0.1;

Fig. 5). Mean persistence in each of the other families

with 15 or more species in our study sample was very

similar in the UK and Estonia. Apart from the Orchida-

ceae, many species of which are rare and protected in

Europe, and the Potamogetonaceae, of which many spe-

cies have specific habitat requirements, persistence var-

ied considerably in these well-represented families.

Most of these families include species capable of occu-

pying a wide range of habitat types, but with the excep-

tion of Juncaceae, higher family level persistence was

found in Estonia (Fig. 5).

Our examination of persistence of plant species in

the UK and Estonia excluded rare, specialized and

narrowly distributed species that occur only in one

or the other country. Most of the species included in

our analysis are rather widely distributed throughout

Europe, and they are therefore probably more persis-

tent over ecological time scales. This may explain

why evolutionary adaptations and ecological prefer-

ences did not make a significant contribution to

explanation of differences in persistence between the

two countries.

Comparisons of plant distributions during the second

half of the 20th century demonstrated that widespread

and common European plant species are declining in

range by, on average, 30% in the UK and 25% in Esto-

nia. Species’ affinity for urban habitats (urbanity) and

tolerance of non-natural conditions (hemeroby) were

both significant factors explaining species persistence

within the countries. However, despite very large dif-

ferences in mean human population density and land

use between the two studied countries, differences in

persistence between the UK and Estonia were not pri-

marily affected by species’ tolerance towards human

activities, or by their mode of propagation or height.

Floristic element affiliation and CSR strategy type both

influenced persistence. A competitive strategy, a more

local floristic elemental origin (Eurasian in Estonia and

European in the UK) and tolerance of human influences

promote local persistence of plant populations. More

detailed studies are now needed to evaluate in greater

depth the way in which these key factors affect local

and regional persistence, their inter-relationships, and

the persistence of more narrowly distributed and rarer

species in the face of changes in vegetation, climate and

human activities.

Acknowledgements

The work of LL, TK and MS was supported by institutionalresearch funding IUT 21-1 of the Estonian Ministry of Educationand Research; we also acknowledge the support of the herbar-ium TAA. Additional funding for LL was provided by the Esto-nian Research Council grant PUT (607) and the EuropeanCommission through the European Regional Fund (the Centerof Excellence in Environmental Adaptation). PM was supportedby MSMT LM2010009 CzechPolar. We thank Meelis P€artel andtwo anonymous referees for valuable comments. The authorsdeclare no conflict of interest.

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Supporting Information

Additional Supporting Information may be found in theonline version of this article:

Appendix S1. List of plant species common to the UK andEstonia that were included in the analysis.Appendix S2. Variables used in the analysis. The numbersof species within each variable category and each quantita-tive class category are shown, together with the total num-ber of species included in the analysis.



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