Journal of Ecology

2001

89

, 451–463

© 2001 British Ecological Society

451

Blackwell Science, Ltd

Demographic variation and population viability in

Gentianella campestris

: effects of grassland management and environmental stochasticity

TOMMY LENNARTSSON and J. GERARD B. OOSTERMEIJER*

Swedish University of Agricultural Sciences, Department of Conservation Biology, Box 7002, S-750 07 Uppsala, Sweden; and

*

University of Amsterdam, Institute for Biodiversity and Ecosystem Dynamics Section Experimental Plant Systematics P.O. Box 94062, NL 1090GB Amsterdam, the Netherlands

Summary

1

Transition matrix models were used to evaluate the effects of environmentalstochasticity and four different methods of grassland management on dynamics andviability of a population of the biennial

Gentianella campestris

(Gentianaceae) inspecies-rich grassland. Data were collected between 1990 and 1995.

2

Continuous summer grazing, the prevailing management strategy in Scandinaviangrasslands, resulted in high recruitment of new plants, mainly because litter accumulationwas prevented and gaps were created by trampling. Trampling and repeated grazing,however, caused damage which reduced seed production. Lambda for the averagematrix was

c

.

0.77, and a stochastic matrix model yielded an extinction probability forthe total population of

c.

0.08 within 50 years.

3

Mowing in mid-July (used as a conservation tool) increased seed production, but litteraccumulation following re-growth of the vegetation prevented establishment. Lambdaand extinction risk were similar to continuous grazing.

4

Mowing in October (another conservation tool) promoted recruitment because oflow litter accumulation, but the seed output decreased because plant growth wasimpaired by tall vegetation. Lambda was 0.64, while the extinction probability was veryhigh (

c.

0.98 within 50 years).

5

Mid-July mowing followed by autumn grazing (the historical managementregime) yielded high values for both seed production and establishment of rosettes.Lambda was 0.94 and the probability of extinction within 50 years was below detec-tion level.

6

Log-linear analysis showed that the matrices differed significantly both betweentreatments and between years. The latter indicates environmental stochasticity, herecaused by summer drought that increased the extinction risk. Lambda may be slightlyunderestimated because drought occurred in one out of five summers during the studyperiod, which is high compared with the natural frequency.

7

We conclude that traditional grassland management is more favourable for

G. campestris

than the methods that prevail in Scandinavia today. This indicates a serious conservationproblem, because grazing has replaced traditional management in many of the remainingsemi-natural grasslands throughout Europe.

Key-words

:

environmental stochasticity,

Gentianella

, grassland management, popula-tion dynamics, population viability analysis, recruitment, seed production, transitionmatrix models

Journal of Ecology

(2001)

89

, 451–463

Correspondence: T. Lennartsson, Swedish University of Agricultural Sciences, Department of Conservation Biology, Box 7002,S-750 07 Uppsala, Sweden (fax: +46 (0)18 67 35 37; e-mail: tommy.lennartsson@nvb.slu.se).

JEC566.fm Page 451 Tuesday, May 15, 2001 2:30 PM

452

T. Lennartsson & J.G.B. Oostermeijer

© 2001 British Ecological Society,

Journal of Ecology

,

89

, 451–463

Introduction

A majority of the open grassland habitats in Europehas been created and maintained by human manage-ment. Without such disturbance of the vegetation, mostterrestrial areas in Europe would be covered by forestor shrubland (e.g. Påhlsson 1994). Management affectsalmost all aspects of grassland dynamics and studiesof species diversity and population dynamics shouldtherefore include management as one of the majorenvironmental factors (e.g. Morris 1991; Zobel 1992).Here, we conduct experimental tests of the hypothesisthat the different management methods in current usein Scandinavia, as well as the traditional form of man-agement, all exert different effects on the dynamics andviability of plant populations in semi-natural grasslands.

Most of the agricultural systems in temperate regionshave historically been based on unfertilized, fodder-producing habitats such as grassland pastures and mownhay-meadows. Such habitat types were necessary forlivestock farming, where the animals served both asfood, and supplied manure to sustain crop productionin permanent fields. In many regions pastures and hay-meadows were the dominant biotopes in the landscape(e.g. Emanuelsson

et al

. 1985). Unfertilized fodder-producing grasslands are often referred to as semi-natural (e.g. Duffey 1974; Rackham 1986; Berglund

et al

. 1991). A rapid decline of semi-natural grasslandsstarted in Scandinavia 50–100 years ago when artificialfertilizers began to replace manure, and intense animalhusbandry or farming without livestock replaced tradi-tional farming systems. Today, only fragments remain,and their area is still decreasing. Similar trends can beseen in all European countries. In Sweden the area ofmown semi-natural grasslands decreased from

c.

1.2

×

10

6

–2400 hectares between 1880 and 1990 (StatistiskaCentralbyrån 1990; Bernes 1994) and the United King-dom (UK) has lost more than 95% of its originallowland herb-rich meadows (Nature ConservancyCouncil 1984).

Well-managed, unfertilized semi-natural grasslandsare extremely rich in species, especially vascular plants,fungi and phytophagous insects, and the decline in thearea of the grasslands is associated with species loss. InSweden, for example,

c.

600 red-listed grassland plants,vertebrates and invertebrates are threatened by cessationor change in management (Bernes 1994), and abandon-ment of semi-natural grasslands has been identified asone of the major factors adversely affecting flora andfauna in Europe (Wolkinger & Plank 1981; Fuller1987; Tucker 1991; Stanners & Bourdeau 1995).

In addition to the extensive reduction in their habitatarea, the threats to grassland species also include changesin management from the traditional methods. Theconservation significance of this has attracted littleattention compared with the loss of habitats, cessationof management, and the genetic consequences of habitatfragmentation (e.g. Bacon 1990; Fry 1991; Schemske

et al

. 1994). Management methods, however, clearly

affect the diversity of, for example, plants (e.g. Wells1971; García & Navascues 1989; García 1992) andinvertebrates (e.g. Morris 1991), and the degree ofalteration of management may be correlated with thenumber of species lost (van Leeuwen 1966; Bakker 1979).In Scandinavia, many plant species show a decline evenin managed semi-natural grasslands, which may indi-cate the ecological significance of such alterations ingrassland management (Eriksson 1996; Lennartsson &Svensson 1996).

One especially widespread alteration in managementis the shift from mowing to grazing (e.g. García 1992;Bernes 1993; Beaufoy

et al.

1995). In Scandinavia,mowing for hay has been an essential component oflivestock farming for at least 2000 years, with grasslandsin the vicinity of villages traditionally mowed in thesummer and grazed in the autumn (Aronsson & Matzon1987; Ekstam

et al.

1988). Today, management of theremaining grasslands is usually restricted to grazing.When semi-natural grasslands are still mown (usuallyin nature reserves), autumn grazing is rarely included inthe management. Late mowing has been tried as a con-servation tool, in order to allow plants and insects toreproduce before disturbance (cf. Bakker

et al.

1980).The differences between traditional and present man-

agement may be small in terms of annual biomassremoved (e.g. Ekstam & Forshed 1996). However, largerdifferences may be found in timing, type and intensityof disturbance to the vegetation (Ekstam

et al.

1988),and these can be expected to strongly affect the survival,growth and recruitment of plants. The scope of thisstudy is restricted to analysis of the effects of livestockgrazing and mowing for hay, although non-humanfactors also influence biodiversity in semi-naturalgrasslands. For example, grasslands in other parts ofEurope may be strongly affected by grazing by wildanimals such as rabbits (Thomas 1960, 1963). Anotherimportant factor and conservation issue is soil nutrientstatus, including the increasing problem with atmosphericdeposition of nitrogen (e.g. Grubb 1976; Weiss 1999;Haddad, Haarstad & Tilman 2000).

In this field experiment, different management methodswere evaluated by studying the demography of thered-listed grassland biennial

Gentianella campestris

ssp.

campestris

(L.) Börner. The management types studiedwere:

1.

grazing from early June to late September (hereaftercalled grazing);

2.

mid-July mowing (July mowing);

3.

mid-July mowing followed by grazing from earlyAugust to late September (mowing/grazing); and

4.

mowing after the growing season, in early October(October mowing).

Type 1 has generally replaced the traditional type 3in Scandinavia today whereas 2 and 4 are currently usedin management for conservation. Type 4 also resemblesthe traditional management in grassland patchessurrounded by cereal fields, where the grazing took placein September–October, after cropping of the cereals

JEC566.fm Page 452 Tuesday, May 15, 2001 2:30 PM

453

Population viability of

Gentianella campestris

© 2001 British Ecological Society,

Journal of Ecology

,

89

, 451–463

(Ekstam

et al.

1988). A major aim of the study was touse demographic data to evaluate, from a conservationpoint of view, the differences between traditional andpresent management practices. Furthermore, the studyaimed to estimate the mechanisms behind the differencesby using demographic data. We used transition matricesin combination with a stochastic model to translatepopulation dynamics into estimates of populationviability.

Methods

Gentianella campestris

ssp.

campestris

is a Europeanbiennial herb (Pritchard & Tutin 1972; Hultén & Fries1986). It is spring-germinating, forms a rosette duringthe first summer, and overwinters as a taproot and abud. The plant flowers in the second summer, either inearly July (aestival type) or in August (autumnal type,von Wettstein 1895). Flowering individuals are normally5–20 cm high with 5–15 lilac flowers (Lennartsson1997). The life cycle is strictly biennial, as all plantsflower in their second summer and die after flowering(cf. Kelly 1985). The species has no form of vegetativepropagation (Fig. 1; Lennartsson 1997). The adultplant has a main stem with an apical meristem and 6–12 nodes, each of which bears two opposite, decussateleaves. Each node carries a number of meristems whichcan produce branches with secondary nodes and flowers.Re-growth of new branches and flowers after grazing orother damage is commonly observed in nature (Zopfi1991; Lennartsson

et al.

1998). However, bitter glycosidesubstances (e.g. Inouye

et al.

1967; Hostettmann-Kaldas& Jacot-Guillarmod 1978) give some protection againstherbivory.

The plants are frequently visited by bumblebees(Müller 1881; T. Lennartsson, personal observation),which are the only known pollinators, but the species ishighly self-fertile and normally achieves full seed setwithout insect pollination (Lennartsson

et al

. 2000).Each fruit contains 40–120 ovules. To our knowledge,

G. campestris

only has a transient seed bank (Milberg1994; cf. Kelly 1989a).

Gentianella campestris

occurs mainly in managedsemi-natural grasslands and very few natural habitatsare known (Pritchard & Tutin 1972; Zopfi 1991; Nilsson

et al.

1993). Populations rapidly become extinct followingcessation of management, mainly due to a short livedseed bank, and because litter accumulation increasesthe shading and mortality of seedlings (Milberg 1994;Lennartsson & Svensson 1996). The number of popu-lations in Sweden has decreased rapidly during the lastdecades. For example, in central Sweden it has dis-appeared from

c.

90% of its localities during the last50 years (Lennartsson & Svensson 1996). It is red-listedin Sweden (Aronsson

et al.

1995) as well as in severalother European countries (e.g. Ingelög

et al.

1993).

G. campestris

is associated with grasslands with highbiodiversity and conservation value (Naturvårdsverket1987) and is therefore likely to serve as an appropriateindicator species for evaluating the effects of differentmanagement regimes.

The experiment was performed in a large

G. campestris

population in Björnvad in the province of Södermanlandin central Sweden (59

°

20

′

21.29

″

N, 16

°

51

′

31.31

″

E).The population is located in an open, species-rich grass-land pasture of

c.

15 hectares. The soil is well-bufferedtill, mostly rich in fine fractions varying from clay to

Fig. 1 Life cycle diagram for Gentianella campestris. Arrows with broken line indicate fecundity transitions. Fij and aij are transitionsfrom stage j to stage i in 1-year time intervals (autumn–autumn). F denotes ‘fecundity’ transitions, a denotes ‘fate’ transitions.

JEC566.fm Page 453 Tuesday, May 15, 2001 2:30 PM

454

T. Lennartsson & J.G.B. Oostermeijer

© 2001 British Ecological Society,

Journal of Ecology

,

89

, 451–463

sand. The vegetation is a mosaic of

Avenula pratensis-Festuca rubra

meadow and herb-rich

Agrostis capillaris

meadow (Rydberg & Vik 1992; Påhlsson 1994). From1991 to 1996, the average population size was

c.

2500plants (adults and rosettes), with large fluctuationsbetween years. The peak of flowering is around 10 August.Normal management is horse grazing from late May toSeptember, and it has probably been grazed by cattleduring most of this century (information from presentland-owner). On historical maps, major parts of thearea are denoted ‘hay meadow’ and the managementhistory is thus representative of most other semi-naturalgrasslands in the region. In this experiment, the horseswere used for grazing in treatments 1 and 3.

We selected a 75

×

20 m plot with homogeneous vege-tation (including presence of

G. campestris

), homo-geneous mesic to dry soil and a horizontal, even surface.The plot was divided into five subplots, each 15

×

20 m,located in sequence along the main axis. The firstsubplot (hereafter denoted grazing) as well as the fifthremained a part of the normal pasture, open to grazingwhereas the others were fenced off to exclude or controlthe grazing. Subplot five was sampled only in 1993, toreveal any hidden environmental gradients over the75 m axis. Subplot two (mowing/grazing) was mownon 15 July, and was returned to pasture 21 days later byremoving the fence. Subplot three (July mowing) wasmown at the same date but not exposed to any subsequentgrazing. Subplot four (October mowing) was mownafter the growing season, on 5 October. Unfortunatelywe had no resources for arranging and sampling morethan one set of treatment plots. Mowing was performedby clipping with shears 5 cm above the ground and thecut material was removed. This treatment meant that

c.

90% of the

Gentianella

plants were affected by cutting,which can be compared with

c.

20% being damaged bygrazing.

We followed the fate of all individuals of

G. campestris

occurring in the central 9

×

14 m of each subplot (theouter 3 m zone of each of the subplots was omitted toavoid edge-effects and trampling damage along thefence). The number of plants (rosettes plus adults) persubplot was 214–240 at the beginning of the experiment.Each adult plant and rosette was individually identifiedby using a 50

×

50 cm grid, in turn subdivided into5

×

5 cm sections. Adults and rosettes were monitoredon three occasions per year, in early June, on 15 July, i.e.before July mowing and in late September (beforeOctober mowing). We counted the number of stem andbranch nodes, and the number of fruits on adult plants.The number of seeds per fruit was estimated annuallyby counting the seeds in one capsule per plant from 30randomly selected plants per treatment just before seedripening (cf. Kelly 1984). The adult plants were assignedto one of two size-classes (cf. Shea & Kelly 1998) accord-ing to the number of stem- and branch-nodes at fruiting.

This measurement is directly related to fecundity, becauseflowers and fruits are only produced at the nodes. Plantswith more than 10 nodes in total (approximately themedian number of nodes during the study) were con-sidered ‘large adults’, and plants with fewer than 10nodes considered ‘small adults’. The size of the rosetteswas calculated as an index given by the rosette diameter(in cm) times the number of leaf pairs. Rosettes weredivided into two size classes with a size index of 15 (themedian) as the break point.

Annual seedling emergence was counted in 25randomly chosen 50

×

50 cm quadrats per treatmentand the total number of seedlings in the populationwas extrapolated. In order to estimate the propor-tions of seedlings originating from last year’s plantsand from the seed bank, respectively, we performed acontrolled sowing experiment in outdoor cold framesin the Botanical Garden of Uppsala University. Wecollected 100 seeds per plant from 10 mother plantsper treatment (five large and five small plants) andsowed each seed-family in a separate pot with com-mercial plant soil in the autumn every year (

t

) from1991 to 1994. The germination of year

t

’s cohort wasmonitored annually in May–June and the numberof seedlings in the year

t

+ 2 and

t

+ 3 was taken asgermination from the first year’s seed bank (seedbank 1 in Fig. 1) and from the second year’s seed bank(seed bank 2), respectively.

In order to make seed production visible in theprojection matrices (see the following section), weassumed that all seeds produced in year

t

survived toyear

t

+ 1, either by producing rosettes or by enteringseed bank 1 (Fig. 1). Seedling mortality in the year

t

+ 1 was instead included in the seed mortality inseed bank 1, together with seed mortality in the soil(estimated by Milberg (1994), who performed a burialexperiment during the 3 years). As no seeds germinatedafter year

t

+ 3, we considered that all seeds that hadnot already died or germinated were dead after

t

+ 3,thus giving the estimate of the seed mortality in seedbank 2.

The recruitment of large and small rosettes from largeor small adults was estimated by combining the numberof seeds produced by adults of each size-class, the ger-mination of sown batches of seeds, the emergence ofseedlings in the field, and the proportion of seedlingsdeveloping into large and small rosettes, respectively.The recruitment from the seed bank was calculatedfrom the observed germination of dormant seeds andthe annual average seedling survival (to rosettes) foreach treatment.

In early June each year we measured the thicknessof the litter layer (distance between the soil surfaceand the top of the litter layer) at 100 random pointsper treatment. In mid-July (before mowing) and lateSeptember we estimated vegetation height in 25random 0.5

×

0.5 m quadrats per treatment, by using aplastic rising plate, which rested on top of the vegetation(e.g. Murphy

et al.

1995).

JEC566.fm Page 454 Tuesday, May 15, 2001 2:30 PM

455

Population viability of

Gentianella campestris

© 2001 British Ecological Society,

Journal of Ecology

,

89

, 451–463

-

The life-cycle graph for

Gentianella campestris

(Fig. 1)was translated into a projection matrix where thematrix elements

a

ij

and

F

ij

define transitions from stage

j

to stage

i

in 1-year time intervals. The ‘fate’ transitions(

a

ij

) vary between 0 and 1 whereas ‘fecundity’ transitions(

F

ij

) can have values higher than 1. Transition matricesare treated in detail by Caswell (1989). Analysis of aprojection matrix at asymptotic population growth givesthe population’s finite growth rate,

λ

= n

t+

1

/n

t

, where

n

is the population size. It also shows how populationgrowth rate changes due to a change in any matrixelement

a

ij

. To compensate for differences in absolutevalues of

a

ij

, we analysed elasticity, defined as the pro-portional change in

λ

caused by a proportional changein

a

ij

(de Kroon

et al

. 1986). Elasticities sum to one andreflect the relative importance of matrix elements forpopulation growth rate (Silvertown

et al.

1996).The experiment was started in 1990, which allowed

the construction of four autumn to autumn matrices pertreatment, for 1991–92, 1992–93, 1993–94, and 1994–95. To avoid a 1-year time lag in the transitions, the adultplants in year

t

were assumed to produce dormantseeds and rosettes directly in year

t

+ 1 (Caswell 1989).

Several studies in grassland ecosystems have shownthat summer precipitation influences performance ofvascular plants (e.g. Grime & Curtis 1976; Hopkins1978; Bengtsson 1993). In 1994 there was a very lowsummer rainfall (13 mm rain at the site in July, com-pared to a 30-year average of 72 mm at the nearbyUltuna Climate and Bio-climate Station, unpublisheddata). A stochastic model was used to evaluate howdifferent probabilities of summer drought affectedpopulation viability and how these effects differedbetween the treatments. The model sampled one of thefour possible matrices at each time step, with eachmatrix being sampled at a given probability. Theprobability of summer drought was introduced as theprobability of selecting the 1993–94 matrix.

A stochastic model yields a distribution of stablepopulation structures among the replicate populations.We simulated population growth over 50 years, with2000 replicates per simulation, using the computersoftware POPPROJ2, developed by Menges (1990,1992). Extinction probability was given by the propor-tion of iterations resulting in a population size equal tozero within 50 years. For such a relatively short period,extinction probability is preferred over the more com-monly used mean time to extinction (Quinn & Hastings1987; cf. Wilcox & Murphy 1985). We considered thatusing a longer time-span than 50 years in the modelwould be of limited value, because unknown effectsof changes in agricultural policies would overrule theeffects of natural stochasticity and demographic factors.

All statistical analyses were performed with the SPSS9.0 for Windows statistical package. Differences in seedproduction, germination and rosette recruitment betweentreatments (within years) were tested using non-parametrictests. Differences between matrices were analysed usinglog-linear analysis for transition frequency tables, wherethe goodness of fit of the models was assessed by the loglikelihood ratio

G

2

(Caswell 1989). The transition fre-quency table describes the fate of the individuals at eachstate, where the transition frequencies are the numbers(‘counts’) of individuals involved in each transition.Impossible transitions (Fig. 1) were excluded from thelog-linear analyses by assigning them structural zeros.

We tested both the differences between entire matrices,and between the matrices decomposed by state. Thelatter test provides information on the life stages thatmay account for observed differences between thematrices (Caswell 1989). The fate of seeds (seed banks1 and 2) from autumn year

t

to autumn year

t

+ 1 reflectsgermination and establishment of rosettes. The fate ofrosettes reflects the winter survival of rosettes and theperformance of adult plants in year

t

+ 1. Finally, thefate of adults reflects the production of seeds in year

t

,and germination and establishment of rosettes in year

t

+ 1.As the different management methods were tested

in different subplots, this design cannot exclude thepossibility that the observed differences were due toenvironmental gradients rather than management.However, such an explanation is highly unlikely becausethe study area was level, homogeneous in soil andvegetation, and the subplots were situated close to eachother.

Results

On average 26.4 (SD = 4.0)% of the seeds germinatedthe first spring after sowing, 6.3 (0.14)% 1 year later(germination of seed bank 1), and 0.67 (0.04)% 2 yearslater (germination of seed bank 2). No differenceswere detected between treatments in any of the years(

χ

2

≤

0.89, d.f. = 3,

P

≥

0.99).

-

Log-linear analysis revealed no differences between thetwo grazed subplots (1 and 5) for 1993, either for theentire matrices or for the matrices decomposed by state(

P

> 0.821). Hence we report only tests on subplots1–4.

Log-linear analysis showed significant differencesbetween transition matrices, and that the differenceswere explained both by management and year. Withexception of seed bank 2, this was also true for eachstage’s contribution to the total

G

2

(Table 1). Onlyfor seed bank 1 and rosettes were fates significantly

JEC566.fm Page 455 Tuesday, May 15, 2001 2:30 PM

456

T. Lennartsson & J.G.B. Oostermeijer

© 2001 British Ecological Society,

Journal of Ecology

,

89, 451–463

affected by the interaction between management andyear (Table 1).

A comparison of management methods within yearsshowed that their effects differed significantly in 1991–92 and 1992–93, but not in 1993–94 or 1994–95(Table 2). The contribution of the different stages tothe total G 2 was, however, not equal for all years. Ingeneral, the fates of adult plants were not significantlyaffected by management in 1993–94 or 1994–95 whereasthe fates of rosettes were affected by management in allyears except 1994–95 (Table 2).

Elasticities for the average transition matrix for eachtype of management generally showed high values forrosettes growing and developing into large adults (a64,F46, Table 3). The only exception was a low value of F46

for the October-mown population, where small rosettesforming small adults showed high elasticity, indicatingthat these stages were more important (a53, Table 3).The higher importance of small adults in Octobermowing than in any other treatment was also shown bythe elasticity for recruitment of seeds to the seed pool

(F15 > F16, Table 3). The average transition values showedthat establishment of large rosettes (F45, F46) was morecommon than establishment of small rosettes (F35, F36)in two treatments (grazing and mowing/grazing), butwith opposite effects following July or October mowing(Table 4). Small rosettes developing into large adultshad high elasticity in the July mown population, andto some extent in the mown/grazed population (a63,Table 3), clearly due to high transition values of a63 inthese treatments (Table 4). For the average transitionmatrices, λ was highest, and close to 1, for the mown/grazed population, lower for grazing and July mowing(being almost identical), and lowest for the October-mown population (Table 4).

All treatments except grazing allowed the vegetationto grow undisturbed until at least mid-July (Table 5).A considerable re-growth of vegetation was observedfollowing July mowing, although subsequent grazing,when present, removed most of this vegetation (Table 5).The depth of the litter layer did not differ between thetreatments at the beginning of the experiment. Allmanagement methods except for grazing successivelyincreased the litter layer and after 5 years July mowinghad created a significantly thicker litter layer thanmowing/grazing and October mowing (Table 5).

Table 1 Log-linear analysis of the between-year and between-management variation in transition matrices for Gentianella campestris, based on transitionfrequency tables. The log-likelihood ratios, G2, for the decomposed matrices show the relative contribution of each state to the G 2 for the entire matrices.S = state; F = fate; Y = year; M = management. SYM, SF, etc., in the model column shows which terms are included in the model, see text for explanation(Cf. Caswell 1989, chapter 7)

Model

Entire matrices

Matrices decomposed by state

Seed bank year 1

Seed bank year 2 Small rosette Large rosette Small adult Large adult

d.f. G 2 d.f. G 2 d.f. G 2 d.f. G 2 d.f. G 2 d.f. G 2 d.f. G 2

Management or Year 540 1682*** 45 266*** 30 10.6 NS 30 334*** 30 315*** 30 135*** 30 622***SYM, SFManagement 432 669*** 36 144*** 24 5.32 NS 24 220*** 24 86.1*** 24 41.4* 24 172***SYM, SFYYear 432 1259*** 36 201*** 24 6.20 NS 24 186*** 24 259*** 24 107*** 24 500***SYM, SFMManagement x Year 324 209 NS 27 81.3*** 18 1.19 NS 18 53.8*** 18 41.8** 18 7.03 NS 18 23.2 NSSYM, SFY, SFM

*P < 0.05, **P < 0.01, ***P < 0.001, NS = not significant.

Table 2 Log-linear analysis of the effect of management within year (SYM, SFY, Table 2) on the differences between transitionmatrices for Gentianella campestris, based on transition frequency tables. The log-likelihood ratios, G 2, for the entire anddecomposed matrices are as in Table 1

YearEntire matrices (G 2)a

Matrices decomposed by state (G 2)

Seed bank year. 1b

Seed bank year. 2c

Small rosettec

Large rosettec

Small adultc

Large adultc

1991–92 192*** 37.4*** 0.44 NS 72.8*** 4.35 NS 18.0** 56.2***1992–93 348*** 63.9*** 2.79 NS 123*** 40.8*** 19.3** 98.2***1993–94 76.1 NS 207* 0.28 NS 17.0** 30.4*** 0.45 NS 7.02 NS1994–95 52.4 NS 22.3** 1.82 NS 6.90 NS 10.5 NS 3.21 NS 7.77 NS

ad.f. = 108, bd.f. = 9, cd.f. = 6. * P < 0.05, ** P < 0.01, *** P < 0.001, NS = not significant.

JEC566.fm Page 456 Tuesday, May 15, 2001 2:30 PM

457Population viability of Gentianella campestris

© 2001 British Ecological Society, Journal of Ecology, 89, 451–463

Table 3 Elasticity values for the average transition matrix for each treatment. The average transition values are given in Table 4

Table 4 Average transition values and finite rate of population increase, λ, of four matrices for each treatment

Management method

Transition Grazing July mowing Mowing/grazing October mowing

a21 0.0184 0.0191 0.0163 0.0165a31 0.0181 0.0320 0.0226 0.0374a41 0.0731 0.0537 0.0630 0.0473a32 0.0042 0.0056 0.0049 0.0068a42 0.0142 0.0135 0.0113 0.0097a53 0.0767 0.0356 0.0433 0.1984a63 0.0022 0.1454 0.0540 0.0024a54 0.0671 0.0106 0.0192 0.0475a64 0.2901 0.2464 0.3245 0.1928F15 0.0393 0.0111 0.0186 0.0565F35 0.0191 0.0153 0.0106 0.0890F45 0.0854 0.0198 0.0332 0.1004F16 0.0702 0.0937 0.0832 0.0447F36 0.0376 0.1280 0.0591 0.0676F46 0.1845 0.1701 0.2362 0.0830

Management method

Grazing July mowing July mowing and grazing October mowing

Transition Mean SE Mean SE Mean SE Mean SE

a21 0.7123 0.0258 0.6922 0.0232 0.7144 0.0132 0.6622 0.0386a31 0.77 × 10 –4 0.23 × 10 –4 0.63 × 10 –4 0.25 × 10 –4 0.69 × 10 –4 0.18 × 10 –4 1.10 × 10 –4 0.34 × 10 –4

a41 1.84 × 10 –4 0.57 × 10 –4 0.55 × 10–4 0.20 × 10 –4 1.93 × 10 –4 0.77 × 10 –4 0.52 × 10 –4 0.14 × 10 –4

a32 0.12 × 10 –4 0.036 × 10 –4 0.088 × 10 –4 0.033 × 10 –4 0.089 × 10 –4 0.024 × 10 –4 0.17 × 10 –4 0.056 × 10 –4

a42 0.28 × 10 –4 0.090 × 10 –4 0.078 × 10 –4 0.026 × 10 –4 0.24 × 10 –4 0.094 × 10–4 0.080 × 10 –4 0.023 × 10 –4

a53 0.5205 0.0604 0.2731 0.0677 0.2609 0.0305 0.6055 0.1014a63 0.0100 0.0100 0.4619 0.1313 0.1899 0.1033 0.0045 0.0045a54 0.2127 0.0650 0.0769 0.0210 0.0709 0.0392 0.2261 0.1438a64 0.4776 0.0870 0.7025 0.1318 0.5492 0.1296 0.6432 0.1253F15 405 92.4 523 170 445 84.1 417 87.9F35 0.1539 0.0647 0.1949 0.0948 0.1553 0.0486 0.2607 0.0895F45 0.3632 0.1583 0.1748 0.0824 0.4845 0.2272 0.1305 0.0428F16 843 200 1297 319 1202 298 737 178F36 0.3320 0.1528 0.4482 0.2329 0.3779 0.1179 0.4606 0.1752F46 0.7847 0.3710 0.3975 0.1851 1.2519 0.6458 0.2326 0.0862λ 1991–92 1.002 0.991 1.411 0.837λ 1992–93 1.234 1.345 1.509 0.917λ 1993–94 0.275 0.257 0.289 0.315λ 1994–95 0.589 0.504 0.560 0.482λ average matrix 0.775 0.774 0.942 0.638

Table 5 Annual vegetation height and depth of litter layer under four different management regimes

Vegetation height, cm (SE) Depth of litter layer, mm (SE)a

Mid-July Late September Early June

Year GrazingJuly mowing

Mowing/grazing

October mowing Grazing

July mowing

Mowing/grazing

October mowing Grazing

July mowing

Mowing/grazing

October mowing

1990 1.7 (0.13) 7.3 (0.41) 7.4 (0.36) 7.1 (0.27) 1.4 (0.11) 7.4 (0.46) 2.5 (0.25) 9.0 (0.40) 12.4 (0.40)a 12.5 (0.42)a 12.8 (0.48)a 11.9 (0.41)a

1991 2.8 (0.16) 5.2 (0.29) 5.3 (0.22) 6.2 (0.41) 2.6 (0.15) 6.7 (0.38) 3.9 (0.22) 8.9 (0.35) 14.9 (0.55)a 14.7 (0.52)a 15.6 (0.58)a 13.0 (0.51)a

1992 2.4 (0.19) 6.8 (0.37) 6.5 (0.30) 7.0 (0.39) 1.9 (0.16) 5.9 (0.45) 2.0 (0.14) 7.9 (0.25) 14.1 (0.37)a 19.9 (0.61)b 15.3 (0.61)a 15.0 (0.49)a

1993 2.2 (0.11) 6.4 (0.20) 5.8 (0.31) 6.3 (0.26) 2.2 (0.19) 6.6 (0.30) 3.6 (0.20) 9.2 (0.30) 11.6 (0.52)a 16.6 (0.62)b 12.6 (0.50)a 13.8 (0.73)a

1994 2.0 (0.14) 4.8 (0.28) 4.7 (0.24) 5.1 (0.27) 2.1 (0.14) 5.1 (0.26) 2.5 (0.18) 6.4 (0.36) 13.3 (0.44)a 23.6 (0.63)c 16.0 (0.57)b 16.1 (0.65)b

1995 3.1 (0.20) 6.2 (0.30) 5.4 (0.26) 6.7 (0.33) 1.7 (0.12) 6.5 (0.33) 1.8 (0.21) 8.4 (0.41) 11.1 (0.29)a 21.1 (0.56)c 14.4 (0.44)b 15.3 (0.44)b

aOne-way for each year separately, with Tukey’s correction for multiple comparisons.Means denoted by the same letter within a row do not differ significantly at the 0.05 level.

JEC566.fm Page 457 Tuesday, May 15, 2001 2:30 PM

458T. Lennartsson & J.G.B. Oostermeijer

© 2001 British Ecological Society, Journal of Ecology, 89, 451–463

A higher probability of dry years (matrix 1993–94)strongly increased the probability of extinction inGentianella campestris (Fig. 2a). Populations in the

treatments grazing and July mowing were most sensitiveto drought, whereas October mowing showed veryhigh, and mowing/grazing very low, extinction prob-abilities, irrespective of drought frequency (Fig. 2a).

The same relative differences between the manage-ment methods were shown when extinction probabilitywas expressed as a function of initial population size(Fig. 2b). For the grazed and the July mown popula-tions, the extinction probability strongly increased inpopulations smaller than 200–300 plants (rosettes +adults). For the mown/grazed population, an extinctionprobability could be detected only when initial popu-lation size was smaller than c. 400 plants (Fig. 2b).

The number of plants in the treatment subplots variedover time, roughly following the λ alues in Table 4.During the first 2 years all treatments except Octobermowing resulted in growing populations, especiallymowing/grazing which increased from 214 to 307 plants.The dry summer in 1994 decreased the number of plantsin all subplots to 35–70.

Discussion

This study demonstrates that the choice of managementmethod has a strong effect on population viability inthe grassland biennial Gentianella campestris. In par-ticular, the currently prevailing management method inScandinavia (grazing throughout the growing season)resulted in considerably higher probability of popu-lation extinction than the traditional method of Julymowing followed by autumn grazing. The differentmethods affected population viability mainly throughtheir impact on seed production and establishment ofnew plants.

We consider it highly unlikely that the observeddifferences between treatment plots are due to environ-mental differences other than management method.The study area was homogeneous, the subplots locatedclose to each other, and the demographic differencesdid not follow the spatial order of the treatment plots.In addition, the two grazed subplots, located at eitherend of the study area differed from the adjacent plotswith different management, but not from each otherin 1993. Finally, the differences in overall populationdynamics between treatment plots could in all cases beexplained by observations of how the treatments affectedthe different life stages of G. campestris.

In Gentianella campestris, seed production can beassumed to be determined by adult size and damage toadults. Size is, in turn, correlated with rosette growth inthe previous year. Seed production was high in Julymowing and mowing/grazing, but lower in grazingand, especially, October mowing (Table 4). The lowseed production in grazing was probably an effect ofdamage to adult plants by trampling and repeatedgrazing, as indicated by the high proportion of large

Fig. 2 Probability of extinction within 50 years based on 2000iterations of a stochastic model with four autumn–autumnmatrices. The matrices are used in the model in random order,each matrix at a given probability. j = grazing; h = Julymowing; n = mowing and grazing; s = October mowing.(a) The effect of varying probability of summer drought(probability of occurrence of the matrix 1993–94), with allother matrices sampled with equal probabilities. The initialpopulation size equals the average size of the studypopulation, 2500 plants (rosettes plus adults). The initialnumber of seeds in the seed bank as in Fig. 2b. (b) The effectof varying initial population size, n. The number of smallrosettes, large rosettes, small adults and large adults is givenby n /4; the number of seeds in seed bank 1 by n × 500; thenumber of seeds in seed bank 2 by n × 225. The probability ofdry years equals the average probability for the region, 0.16.

JEC566.fm Page 458 Tuesday, May 15, 2001 2:30 PM

459Population viability of Gentianella campestris

© 2001 British Ecological Society, Journal of Ecology, 89, 451–463

rosettes developing into small adults (a54, Table 4). Thistransition value is high also in October mowing, buthere adult growth was inhibited by tall vegetationduring the entire growing period. Such an associationbetween fecundity and turf height has been establishedfor several grassland plants, especially short-lived species,for example the related G. amarella (Kelly 1989b).

Seed production was favoured (although not signi-ficantly, Mann–Whitney P ≥ 0.07) by mid-July mowingcompared to grazing throughout the growing season.This may be surprising as most of the plants were cut bythe mowing, which took place 4–5 weeks before the peakof flowering in the population. However, G. campestrishas been shown to have a high capacity for compensatoryseed production after damage and it may even over-compensate for damage caused by mowing (Lennartssonet al. 1997; Lennartsson et al. 1998) or grazing (Zopfi1991). Although this study does not control for initialsizes or comparable damage levels between plants,over-compensatory growth after damage is suggestedby high transition values for small rosettes developinginto large plants (a63, Table 4). This transition indicatesthat adults become larger than would be expected fromthe rosette sizes in the previous year. Normally, smallrosettes develop into small adults (a53). For the Julymown population, the elasticity value for a63 is also high.The plants responded to cutting by vigorous branchingand growth of new nodes, which increased the sizeof the adults. The compensatory growth was par-ticularly high in the July mown population, whilethe mown/grazed plants experienced some tramplingand secondary herbivory.

The probability that a seed will develop into a newrosette in G. campestris can be assumed to be a com-bined effect of conditions for germination/seedlingestablishment, and rosette mortality during the firstsummer. In the July mown population the recruitmentof rosettes was low (sum of F35–46 = 1.21, Table 4) inspite of a high seed production (sum of F15–16 = 1820).In the grazed population the opposite pattern wasindicated, with high recruitment of rosettes (sum ofF35–46 = 1.63, Table 4), but a lower seed production(sum of F15–16 = 1248). The number of rosettes perseed was thus almost twice as high under grazing(1.3 × 10 –3 rosettes per seed) as under July mowing(0.66 × 10 –3). The mown/grazed population had similarrosette recruitment to grazing (1.4 × 10–3 rosettesper seed), but in combination with high transitionvalues for seed production (Table 4). In a similarexperiment, Shea & Kelly (1998) demonstrated, usingmatrix models, an effect of management on estab-lishment of new plants of the invasive Carduus nutansin New Zealand.

Establishment of rosettes thus seems to be favoured byincluding grazing in the management. Enhanced recruit-

ment in a grazed turf can be expected if germinationand seedling establishment is facilitated by gaps inthe turf (Grubb 1976, 1988; Kelly 1989b; Hillier 1990).For G. campestris, depth of the litter layer may be themain factor limiting recruitment, as indicated by differ-ences in establishment between July mowing (0.66 × 10–3

rosettes per seed) and October mowing (1.2 × 10–3).Neither July nor October mowing creates any gaps bydamaging the turf, but after mowing in mid-July, thevegetation had time for substantial re-growth, and thisbiomass appeared as litter in the following spring.October mowing, in contrast, removed the hay too latefor re-growth, thus producing less litter. Grazing afterJuly mowing may facilitate establishment by reducinglitter production, in addition to creating gaps by tramp-ling. Rosette production per seed among treatmentswas negatively correlated with thickness of litter layerin 2 years out of 4 (Spearman rank correlation, P < 0.001).In 2 of the years, trampling caused an increasedmortality of rosettes in the grazed and mown/grazedpopulations.

Tall vegetation during the summer clearly also affectedgrowth of the rosettes. When grazing was included inthe management, large rosettes were more commonthan small rosettes (F45, F46 > F35, F36, Table 4), whilewithout grazing the situation was reversed. UnderOctober mowing, the rosettes encountered tall vegeta-tion throughout the summer, and thus only c. 30% ofthe rosettes became large (cf. 70% under grazing, 48%under July mowing, and 76% under mowing/grazing).These figures, together with a high elasticity for F36 inJuly mowing, show that rosettes may also suffer tosome extent from shade in the July-mown population,probably due to tall vegetation in the late summer.Similar results were obtained for the perennial Gentianapneumonanthe L. (Oostermeijer et al. 1994; Oostermeijeret al. 1996) and for the biennial Gentianella amarella inchalk grassland (Kelly 1989b). For both species thesurvival of seedlings and juvenile plants was negativelyaffected by turf height or cover.

The 1994 drought had a strong negative impact on theestablishment and survival of rosettes, and thus reducedboth the differences between treatments (non-significanteffects of management on the fate of adults in 1993–94,Table 2) and in vegetation height between treatments(non-significant differences between treatments forthe fate of rosettes in 1994–95, Table 5). However,the low G 2 in 1994–95 is also explained by low numberof plants the year after the drought (35–70 plants pertreatment plot). Grubb (1976) concluded from theEuropean distributions of shallow-rooted hapaxanthsthat this group of plants in general has limited drought-tolerance.

The effects of different management methods onseed production and rosette establishment resulted indifferential population viability of Gentianella campestris,

JEC566.fm Page 459 Tuesday, May 15, 2001 2:30 PM

460T. Lennartsson & J.G.B. Oostermeijer

© 2001 British Ecological Society, Journal of Ecology, 89, 451–463

measured as population growth rate and probability ofpopulation extinction. It is clear that the traditionalmanagement regime – July mowing followed by autumngrazing – was considerably more favourable than thecurrent practice of grazing only. July mowing withoutsubsequent grazing was even slightly more deleterious,nor can October mowing be recommended as a tool forconservation of this species.

The significance of summer drought for the popu-lation viability of G. campestris was demonstrated bothby the lower population growth rate, λ (Table 4), andfor grazed and July-mown populations by the higherextinction risk (Fig. 2). This suggests that G. campestrisis sensitive to climatic changes that influence precipita-tion and that it can be expected to have a higher extinc-tion risk in the drier habitats into which lowlandGentianella species in Sweden have been forced toretreat during the last 50 years (Lennartsson & Svensson1996). In the absence of management, populationsbecome extinct faster in mesic than in dry habitats,because mesic vegetation types have taller vegetationand more rapid litter accumulation. The potentiallyincreased extinction risk for populations in dry hab-itats has to be taken into account when analysing thenational or regional status of G. campestris.

Averaging the λ estimates in Table 4 always gavevalues below 1, both in terms of geometric mean of theseparate years (0.77 ≥ λ ≥ 0.58), and for the averagetransition matrices (arithmetic mean; 0.94 ≥ λ ≥ 0.64),indicating that the population will become extinctregardless of management method. However, the averagematrices include the effect of drought occurring in onesummer out of five, compared with an average frequencyof 0.16 for the region (Andersson 1970; Eriksson 1980).Longer term estimates of λ will be slightly higher, butthe geometric means will remain below 1. We concludethat many populations of Gentianella campestris can beexpected to decline, especially when managed by con-tinuous grazing. This is supported by field inventoriesshowing that Gentianella populations are often very small,and that many grazed populations have become extinctduring the last 50 years (Lennartsson & Svensson 1996).

Average values of λ (arithmetic or geometric) providea poor estimate of population growth rate in a stochasticenvironment, because both growth rate and populationsize become random variables. Instead, lnλ can be used(Caswell 1989), or – as in this analysis – extinctionprobability within a certain time period, based onstochastic simulation of population growth. The presentpopulation of G. campestris in Björnvad (on averagec. 2500 plants) with the current management regime(grazing) has an extinction probability within 50 yearsof c. 0.08 if the probability of dry years = 0.13. Thisindicates that, in spite of negative λ, the decline israther slow, but nevertheless, it is obvious that tradi-tional management would provide considerably betterconditions for persistence.

A more alarming conservation problem may beindicated by the relationship between population size

and probability of extinction (Fig. 2b). In the grazedand July-mown populations, the shape of the curvesindicates that there is a threshold population size ofc. 200–300 plants for viability. Below this, a smalldecrease in population size produced a large increasein extinction risk (steep slope of the curves), which, ifgeneral, is alarming for the overall status of the species.An inventory in two regions in Sweden showed that lessthan 10% of the populations contained more than 200plants (Lennartsson & Svensson 1996). Increased useof traditional management (mowing/grazing) could,however, decrease extinction probability within 50 yearsin small populations from 30 to 50% to 1–5% (Fig. 2b).Kéry et al. (2000) found strongly reduced reproductionin populations of the grassland plants Primula verisand Gentiana lutea consisting of less than 200 and 500plants, respectively, as an effect of inbreeding depres-sion. Hence, the extinction thresholds demonstrated inour study may be even more pronounced if inbreedingeffects are included.

Our model assumes that no more than a 3-year-oldseed pool exists. This assumption is supported by ourgermination data as well as by Milberg (1994), althoughsome field observations have suggested a seed bank ofat least 6 years longevity (O. Eriksson, personal com-munication) or even longer (H. Öhrling, personal com-munication), which would prolong the average time toextinction (e.g. Bakker et al. 1991; Kalisz & McPeek1993; Dutoit & Alard 1995). This should not, however,alter the relative effects of the different managementmethods or the overall tendencies. The model estimatespopulation growth in a continuously and rapidlyvarying environment (cf. Leigh 1981). Other possibleapproaches are stochastic extinction models determinedby demographic stochasticity only, and models deter-mined by environmental disasters (Shaffer 1981; Quinn& Hastings 1987; Menges 1990). Pure demographicstochasticity obviously does not apply here, becausethe effect of year is significant in all treatments.Environmental stochasticity is included in the model asdrought, and we are aware of no other environmentalcatastrophes in this type of grassland as long asmanagement is continued.

The capacity for compensatory growth following damageto Gentianella campestris may be very high comparedwith that of other monocarpic plants (e.g. Lehtilä &Syrjänen 1995; Bergelson et al. 1996). Although theeffects of July mowing and mowing/grazing may thusbe less representative for other grassland species, highcompensation capacity is not the only reason for a plantpotentially being favoured by mowing rather thangrazing. Mowing can be assumed to favour, for example,species which are especially palatable to grazers (Ekstam& Forshed 1996), early flowering species or ecotypes

JEC566.fm Page 460 Tuesday, May 15, 2001 2:30 PM

461Population viability of Gentianella campestris

© 2001 British Ecological Society, Journal of Ecology, 89, 451–463

(Karlsson 1984; Lennartsson 1997; Lennartsson et al.1997), or species which are sensitive to herbivory orother damage early in the season (Paige 1994). Theimportance of the regeneration niche (Grubb 1977), asemphasized by this study, is likely to be widespread ingrassland species, especially monocarpic species andperennials without vegetative reproduction.

To conclude, this study strongly advises an increaseduse of traditional management in semi-natural grass-lands where Gentianella campestris occurs. We believethat this is likely to be beneficial for other vascularplant species, as G. campestris usually occurs in the mostspecies-rich grasslands. If traditional management cannotbe ensured, it is important to design other methods thatare functionally equivalent. For example, late onset ofgrazing might be an equivalent to mowing/grazing. Inparticular, management must provide suitable condi-tions both for seed production and for establishment ofnew plants. Grazing and July mowing had very differenteffects on these two life cycle events, but resulted inalmost identical population growth rates. This indi-cates that we need to know the mechanisms behindthe observed effects, to be able to generalize from theresults of one management experiment, and to makethe results useful for recovery planning. Finally, thisstudy indicates that the alteration of management frommowing/grazing to grazing only, that has taken placeon a large proportion of the European grasslands, mayrepresent a serious conservation problem.

Acknowledgements

We are grateful to the Anduri family in Björnvad forplacing their land at our disposal for the experiments.Jiska van Dijk, Lotta Johansson, Maria Johansson,Robert Lager, Sanna Simán and Ingrid Wänstrand havedone much of the field work. Valuable comments on themanuscript have been given by Roger Svensson, Hansden Nijs, Bengt Jonsell, Lena Jonsell, Karin Bengtsson,three referees and the editor of the Journal of Ecology.Örjan Nilsson and the Uppsala University BotanicalGarden enabled the sowing experiments. Eric Mengeskindly provided computer software and advice for itsuse. The work was financially supported by grants fromthe National Swedish Environmental Protection Agency,the Swedish Association for Conservation of Nature,and the Swedish Biodiversity Centre.

References

Andersson, T. (1970) Swedish Temperature and PrecipitationRecords since the Middle of the 19th Century. NationalSwedish Building Research Document, 4. Svensk Byggtj-dnst, Stockholm.

Aronsson, M., Hallingbäck, T. & Mattsson, J. (eds) (1995)Rödlistade växter i Sverige, 1995. Swedish ThreatenedSpecies Unit, Uppsala.

Aronsson, M. & Matzon, C. (1987) Odlingslandskapet. LT:s,Stockholm.

Bacon, J.C. (1990) The use of livestock in calcareous grass-land management. Calcareous Grasslands: Ecology and

Management (eds S.H. Hillier, D.W.H. Walton & D.A. Wells),pp. 121–127. Proceedings of a joint British Ecological Society/Nature Conservancy Council symposium. BluntishamBooks, Huntingdon.

Bakker, P.A. (1979) Vegetation science and nature conservation.The Study of Vegetation (ed. M.J.A. Werger), pp. 247–288.Junk, den Haag.

Bakker, J.P., Bos, A.F., Hoogveld, J. & Muller, H.J. (1991)The role of the seed bank in restoration management ofsemi-natural grasslands. Terrestrial and Aquatic Ecosystems.Perturbation and Recovery (ed. R. Ravera), pp. 449–455.Ellis Horwood, London.

Bakker, J.P., Dekker, M. & de Vries, Y. (1980) The effect ofdifferent management practices on a grassland communityand the resulting fate of seedlings. Acta Botanica Neerlandica,29, 469–482.

Beaufoy, G., Baldock, D. & Clark, J. (1995) The Nature ofFarming. Low Intensity Farming Systems in Nine EuropeanCountries. Institute for European Environmental Policy,London.

Bengtsson, K. (1993) Fumana procumbens on Öland – popu-lation dynamics of a disjunct species at the northern limit ofits range. Journal of Ecology, 81, 745–758.

Bergelson, J., Juenger, T. & Crawley, M.J. (1996) Regrowthfollowing herbivory in Ipomopsis aggregata: compensationbut not overcompensation. American Naturalist, 148,744–755.

Berglund, B.E., Larsson, L., Lewan, N., Olsson, E.G.A.,Riddersporre, M. & Skansjö, S. (1991) Terms and concepts.The Cultural Landscape During 6000 Years in Southern Sweden(ed. B.E. Berglund), pp. 54–60. Ecological Bulletins 41,special issue. Ecologicall Bulletins, Copenhagen.

Bernes, C. (1993) The Nordic Environment – Present State,Trends and Threats. NORD-reports. The Nordic Councilof Ministers, Stockholm.

Bernes, C. (1994) Biological Diversity in Sweden – a Land Survey.Monitor 14, Naturvårdsverket, Stockholm.

Caswell, H. (1989) Matrix Population Models. Construction,Analysis, and Interpretation. Sinauer Associates, Sunderland.

Duffey, E. (1974) Grassland Ecology and Wildlife Management.Institute of Terrestrial Ecology, London.

Dutoit, T. & Alard, D. (1995) Permanent seed banks in chalkgrassland under various management regimes: their role inthe restoration of species-rich plant communities. BiodiversityConservation, 4, 939–950.

Ekstam, U., Aronsson, M. & Forshed, N. (1988) Ängar. OmNaturliga Slåttermarker I Odlingslandskapet. LT:s,Helsingborg.

Ekstam, U. & Forshed, N. (1996) Äldre fodermarker. Betydelsenav hävdregimen i det förgångna. Målstyrning, Mätning OchUppföljning. Naturvårdsverket, Stockholm.

Emanuelsson, U., Bergendorff, C., Carlsson, B., Lewan, N. &Nordell, O. (1985) Det Skånska Kulturlandskapet. Signum,Lund.

Eriksson, B. (1980) Statistical analysis of precipitation data.Part III: Precipitation data from the last two centuries.Swedish Meteorological and Hydrological Institute, RMK,27, 1–54.

Eriksson, O. (1996) Population ecology and conservation –some theoretical considerations with examples from theNordic flora. Symxbolae Botanicae Upsaliensis, 31, 159–167.

Fry, G.L.A. (1991) Conservation in agricultural ecosystems.The Scientific Management of Temperate Communitiesfor Conservation (eds I.F. Spellerberg, F.B. Goldsmith &M.G. Morris), pp. 415–443. Blackwell, London.

Fuller, R.M. (1987) The changing extent and conservationinterest of lowland grasslands in England and Wales: a reviewof grassland surveys. 1930–84. Biological Conservation, 40,281–300.

García, A. (1992) Conserving the species-rich meadows of Europe.Agricultural Ecosystems and Environments, 40, 219–232.

JEC566.fm Page 461 Tuesday, May 15, 2001 2:30 PM

462T. Lennartsson & J.G.B. Oostermeijer

© 2001 British Ecological Society, Journal of Ecology, 89, 451–463

García, A. & Navascues, I. (1989) Effect of management onfloristic composition and production on species-rich haymeadows. Proceedings of the XVI International GrasslandCongress, pp. 1479–1480. Daver, Montrouge.

Grime, J.P. & Curtis, A.V. (1976) The interaction of droughtand mineral nutrient stress in calcareous grassland. Journalof Ecology, 64, 975–988.

Grubb, P.J. (1976) A theoretical background to the conserva-tion of ecologically distinct groups of annuals and biennialsin the chalk grassland ecosystem. Biological Conservation,10, 53–76.

Grubb, P.J. (1977) The maintenance of species-richness inplant communities: the importance of the regenerationniche. Biological Reviews of the Cambridge PhilosophicalSociety, 52, 107–145.

Grubb, P.J. (1988) The uncoupling of disturbance and recruit-ment, two kinds of seed bank, and persistence of plantpopulations at the regional and local scales. AnnalesZoologici Fennici, 25, 23–36.

Haddad, N.M., Haarstad, J. & Tilman, D. (2000) The effectsof long-term nitrogen loading on grassland insect communities.Oecologia, 124, 73–84.

Hillier, S.H. (1990) Gaps, seed banks and plant species diversityin calcareous grasslands. Calcareous Grasslands: Ecologyand Management (ed. S.H. Hillier), pp. 57–66. BluntishamBooks, Oxford.

Hopkins, B. (1978) The effects of the 1976 drought on chalkgrassland in Sussex, England. Biological Conservation, 14,1–12.

Hostettmann-Kaldas, M. & Jacot-Guillarmod, A. (1978)Xanthones et C-glucosides flavoniques du genre Gentiana(sous-genre Gentianella). Phytochemistry, 17, 2083–2086.

Hultén, E. & Fries, M. (1986) Atlas of North European VascularPlants, North of the Tropic of Cancer. II. Koeltz Scientificbooks, Königstein.

Ingelög, T., Andersson, R. & Tjernberg, M. (eds) (1993) RedData Book of the Baltic Region. Part 1. List of ThreatenedVascular Plants and Vertebrates. Swedish Threatened SpeciesUnit, Uppsala.

Inouye, H., Ueda, S. & Nakamura, Y. (1967) Biosynthesis ofthe bitter glycosides of Gentianaceae, gentiopicroside,swertiamarin, and sweroside. Tetrahedron Letters, 21,3221–3226.

Kalisz, S. & McPeek, M.A. (1993) Extinction dynamics,population growth and seed banks. An example using anage-structured annual. Oecologia, 95, 314–320.

Karlsson, T. (1984) Early-flowering taxa of Euphrasia (Scro-phulariaceae) on Gotland, Sweden. Nordic Journal ofBotany, 4, 303–326.

Kelly, D. (1984) Seeds per fruit as a function of fruits per plantin ‘depauperate’ annuals and biennials. New Phytologist,96, 103–114.

Kelly, D. (1985) On strict and facultative biennials. Oecologia,67, 292–294.

Kelly, D. (1989a) Demography of short-lived plants in chalkgrassland. I. Life-cycle variation in annuals and strictbiennials. Journal of Ecology, 77, 747–769.

Kelly, D. (1989b) Demography of short-lived plants in chalkgrassland. II. Control of mortality and fecundity. Journalof Ecology, 77, 770–784.

Kéry, M., Matthies, D. & Spillmann, H.-H. (2000) Reducedfecundity and offspring performance in small populationsof the declining grassland plants Primula veris and Gentianalutea. Journal of Ecology, 88, 17–30.

de Kroon, H.A., Plaisier, J.M., van Groenendael, J.M. &Caswell, H. (1986) Elasticity; the relative contribution ofdemographic parameters to population growth rate. Ecology,67, 1427–1431.

van Leeuwen, C.G. (1966) A relation theoretical approach topattern and process in vegetation. Wentia, 15, 25–46.

Lehtilä, K. & Syrjänen, K. (1995) Compensatory responses oftwo Melampyrum species after damage. Functional Ecology,9, 511–517.

Leigh, E.G.J. (1981) The average lifetime of a population in avarying environment. Journal of Theoretical Biology, 90,213–239.

Lennartsson, T. (1997) Seasonal differentiation – a conserva-tive reproductive barrier in two grassland Gentianella(Gentianaceae) species. Plant Systematics and Evolution,208, 45–69.

Lennartsson, T., Nilsson, P. & Tuomi, J. (1998) Induction ofovercompensation in the field gentian, Gentianella campes-tris. Ecology, 79, 1061–1072.

Lennartsson, T., Oostermeijer, J.G.B., van Dijk, J. &den Nijs, H.C.M. (2000) Ecological significance and herit-ability of floral reproductive traits in Gentianella campestris(Gentianaceae). Basic and Applied Ecology, 1, 69–81.

Lennartsson, T. & Svensson, R. (1996) Patterns in the declineof three species of Gentianella in Sweden illustrating thedeterioration of semi-natural grasslands. Symxbolae Botan-icae Upsaliensis, 31, 169–184.

Lennartsson, T., Tuomi, J. & Nilsson, P. (1997) Evidence foran evolutionary history of overcompensation in the grasslandbiennial Gentianella campestris (Gentianaceae). AmericanNaturalist, 149, 1147–1155.

Menges, E. (1990) Population viability analysis for an endan-gered plant. Conservation Biology, 4, 52–62.

Menges, E. (1992) Stochastic modeling of extinction in plantpopulations. Conservation Biology: the Theory and Practiceof Nature Conservation, Preservation, and Management (edsP.L. Fiedler & S.K. Jain), pp. 253–276. Chapman & Hall,New York.

Milberg, P. (1994) Germination ecology of the endangeredgrassland biennial Gentianella campestris. BiologicalConservation, 70, 287–290.

Morris, M.G. (1991) The management of reserves and protectedareas. The Scientific Management of Temperate Communitiesfor Conservation (eds I.F. Spellerberg, F.B. Goldsmith &M.G. Morris), pp. 323–347. Blackwell, London.

Müller, H. (1881) Alpenblumen, Ihre Befruchtung Durch Insek-ten und Ihre Anpassungen an Diefelben. W. Engelmann,Leipzig.

Murphy, W.M., Silman, J.P. & Barreto, A.D.M. (1995) Acomparison of quadrat, capacitance meter, HFRO swardstick, and rising plate for estimating herbage mass in asmooth-stalked, meadowgrass-dominant with white cloversward. Grass and Forage Science, 50, 452–455.

Nature Conservancy Council (1984) Nature Conservation inGreat Britain. NCC, Peterborough.

Naturvårdsverket (1987) Inventering Av Ängs-Och Hagmarker.Handbok. Naturvårdsverket, Stockholm.

Nilsson, Ö., Jonsell, B. & Lennartsson, T. (1993) Gentianellacampestris, Fältgentiana och Kustgentiana. Faktablad.Floravård I Jordbrukslandskapet – Skyddsvärda Växter(eds T. Ingelög, G. Thor, T. Hallingbäck, R. Andersson &M. Aronsson), pp. 168–169. SBT Books, Lund.

Oostermeijer, J.G.B., Brugman, M.L., de Boer, E.R. & denNijs, H.C.M. (1996) Temporal and spatial variation in therare perennial herb Gentiana pneumonanthe. Journal ofEcology, 84, 158–162.

Oostermeijer, J.G.B., Van’t Veer, R. & den Nijs, J.C.M. (1994)Population structure of rare, long-lived perennial Gentianapneumonanthe in relation to vegetation and managementin the Netherlands. Journal of Applied Ecology, 31, 428–438.

Påhlsson, L. (1994) Vegetationstyper i Norden. Tema Nordfrom The Nordic Council of Ministers, Copenhagen.

Paige, K.N. (1994) Herbivory and Ipomopsis aggregata:difference in response, difference in experimental protocol:a reply to Bergelson and Crawley. American Naturalist,143, 739–749.

JEC566.fm Page 462 Tuesday, May 15, 2001 2:30 PM

463Population viability of Gentianella campestris

© 2001 British Ecological Society, Journal of Ecology, 89, 451–463

Pritchard, N.M. & Tutin, T.G. (1972) Gentianella Moench.Flora Europaea Volume 3 (eds T.G. Tutin, V.H. Heywood,N.A. Burges, D.M. Moore, D.H. Valentine, S.M. Walters& D.A. Webb), pp. 63–67. Cambridge University Press,Cambridge.

Quinn, J.F. & Hastings, A. (1987) Extinction in subdividedhabitats. Conservation Biology, 1, 198–208.

Rackham, O. (1986) The History of the Countryside. Weidenfeld& Nicolson, London.

Rydberg, H. & Vik, P. (1992) Ängs-Och Hagmarker ISödermanlands Län. Länsstyrelsen Södermanlands län,Nyköping.

Schemske, D.W., Husband, B.C., Ruckelshaus, M.H.,Goodwillie, C., Parker, I.M. & Bishop, J.G. (1994)Evaluating approaches to the conservation of rare andendangered plants. Ecology, 75, 584–606.

Shaffer, M.L. (1981) Minimum population sizes for speciesconservation. Bioscience, 31, 131–134.

Shea, K. & Kelly, D. (1998) Estimating biocontrol agentimpact with matrix models: Carduus nutans in New Zealand.Ecological Applications, 8, 824–832.

Silvertown, J., Franco, M. & Menges, E. (1996) Interpretationof elasticity matrices as an aid to the management of plantpopulations for conservation. Conservation Biology, 10,591–597.

Stanners, D. & Bourdeau, P. (1995) Europe’s Environment.The Dobris Assessment. European Environment Agency,Copenhagen.

Statistiska Centralbyrån (1990) Betesmarker – historiskadata. Statistiska Meddelanden, 36,

Thomas, A.S. (1960) Changes in vegetation since the adventof myxomatosis. Journal of Ecology, 48, 287–306.

Thomas, A.S. (1963) Further changes in vegetation sincethe advent of myxomatosis. Journal of Ecology, 51, 151–186.

Tucker, G. (1991) The status of lowland dry grassland birdsin Europe. The Conservation of Lowland Dry GrasslandBirds in Europe (eds P.D. Goriup, L.A. Batten & J.A. Norton),pp. 37–48. Joint Nature Conservation Committee,Peterborough.

Weiss, S.B. (1999) Cars, cows, and checkerspot butterflies:nitrogen deposition and management of nutrient-poor grass-lands for a threatened species. Conservation Biology, 13,1476–1486.

Wells, T.C.E. (1971) A comparison of the effects of sheepgrazing and mechanical cutting on the structure and botanicalcomposition of chalk grassland. The Scientific Managementof Animal and Plant Communities for Conservation (edsE. Duffey & A.S. Watt), pp. 497–515. Blackwell Scientific,Oxford.

von Wettstein, R. (1895) Der Saison-Dimorphismus alsAusgangspunkt für die Bildung neuer Arten im Pflanzen-reiche. Berichte der Deutsche Botanischen Gesellschaft, 13,303–313.

Wilcox, B.A. & Murphy, D.D. (1985) Conservation strategy:the effects of fragmentation on extinction. American Nat-uralist, 125, 879–887.

Wolkinger, F. & Plank, S. (1981) Dry Grasslands of Europe.European Committee for the Conservation of Nature andNatural Resources, Strasbourg.

Zobel, M. (1992) Plant species coexistence – the role of his-torical, evolutionary and ecological factors. Oikos, 65,314–320.

Zopfi, H.J. (1991) Aestival and autumnal vicariads of Gen-tianella (Gentianaceae): a myth? Plant Systematics andEvolution, 174, 139–158.

Received 28 March 2000 revision accepted 12 December 2000

JEC566.fm Page 463 Tuesday, May 15, 2001 2:30 PM