Science Facing Aliens - Invasive Alien Species in Belgium - Belgian ...

Science Facing Aliens 

Proceedings of a Scientific Meeting on 

Invasive Alien Species


Proceedings of a scientific meeting 

on Invasive Alien Species 

held in Brussels, May 11 th , 2009 

Edited by H. Segers & E. Branquart 

Belgian Biodiversity Platform

Segers, H & E. Branquart (Eds). Science Facing Aliens. Proceedings of a scientific meeting 

on Invasive Alien Species, Brussels, May 11 th 2009. Belgian Biodiversity Platform, 2010. 

ISBN 9789081514903 

D/2010/0339/1 

NUR 910/943 

© 2010 Belgian Biodiversity Platform 

The scientific conference “Science Facing Aliens” was organized on occasion of the International Day 

for Biological Diversity 2009, Theme: Invasive Alien Species, as designated by the Parties to the 

Convention on Biological Diversity 

Organizing Committee 

This conference is an initiative of the Belgian Biodiversity Platform’s Forum on Invasive 

Alien Species 

Scientific Committee 

Tim Adriaens (Instituut voor Natuur- en Bos Onderzoek) 

Etienne Branquart (Belgian Biodiversity Platform) 

Jean-Claude Grégoire (Université Libre de Bruxelles) 

Francis Kerckhof (Royal Belgian Institute of Natural Sciences) 

Erik Matthysen (Universiteit Antwerpen), 

Ivan Nijs (Universiteit Antwerpen) 

Hendrik Segers (Belgian Biodiversity Platform) 

Wim van Bortel (Institute of Tropical Medecine, Antwerpen) 

Sonia Vanderhoeven (Faculté universitaire des Sciences agronomiques de Gembloux) 

Fabienne Van Rossum (National Botanic Garden of Belgium) 

Cover Photographs 

Harmonia axiridis © Gilles San Martin 

Heracleum mantegazzianum © Benoît Bedin 

Hydrocotyle ranunculoides background © Vilda - Yves Adams 

Hydrocotyle ranunculoides © Paul Busselen 

Impatiens glandulifera © Paul Busselen 

Myriophyllum aquaticum © Vilda - Yves Adams 

Pacifastacus leniusculus © Guillaume Doucet 

Procyon lotor © Shutterstock.com 

Psittacula krameri © Frank Adriensen 

Rana catesbeiana © Vilda – Rollin Verlinde

Content 

Welcome to the World? Non-native breeding birds in Flanders (Belgium) 

A. ANSELIN, G. VERMEERSCH & G. SPANOGHE........................................................................ 1 

Alien macro-crustaceans in freshwater ecosystems in Flanders 

P. BOETS, K. LOCK & P L.M. GOETHALS................................................................................ 7 

ISEIA, a Belgian non-native species assessment protocol 

E. BRANQUART et al .............................................................................................................. 11 

Impact of Fallopia spp. on ecosystem functioning: Nitrogen and organic matter cycling and 

implicated soil biota 

N. DASSONVILLE et al. .......................................................................................................... 19 

Soil-dependent growth strategy of invasive plants: empirical evidence and model predictions 

using Carpobrotus edulis as target species 

E. DE LA PEÑA & D. BONTE ................................................................................................ 23 

Detection of intraguild predation by Harmonia axyridis on native cocccinellids by alkaloids 

L. HAUTIER et al. .................................................................................................................. 27 

Measuring the impact of Harmonia axyridis intraguild predation on native coccinellids in the 

field 

L. HAUTIER et al. ................................................................................................................. 31 

Comparing Fallopia japonica, F. sachalinensis and their hybrid F. xbohemica in Belgium: 

population ecology, functional traits and invasiveness 

B. HERPIGNY, G. MAHY & P. MEERTS .................................................................................. 35 

Mediterranean container plants and their stowaways: A potential source of invasive plant 

species 

I. HOSTE & F. VERLOOVE ..................................................................................................... 39 

Water frogs in Wallonia: genetic identification of the introduced taxa (Pelophylax ssp.) and 

impact on indigenous water frogs (Pelophylax lessonae and P. kl. esculentus) 

C. PERCSY & N. PERCSY ....................................................................................................... 45 

Trends in the distribution of the Chinese mitten crab in the Scheldt estuary 

M. STEVENS et al. ................................................................................................................. 51 

The invasion of ring-necked parakeet (Psittacula krameri) in Europe and Belgium: 

mechanisms and consequences for native biota 

D. STRUBBE & E. MATTHYSEN .............................................................................................. 53 

Patterns of Prunus serotina invasion in two contrasting forests 

M. VANHELLEMONT et al. ..................................................................................................... 59 

Soil arthropods associated to the invasive Senecio inaequidens compared to the native S. 

jacobaea (Asteraceae) 

V. VANPARYS ET AL. .............................................................................................................. 65 

Non-indigenous freshwater fishes in Flanders: status, trends and risk assessment 

H. VERREYCKEN, G. VAN THUYNE & C. BELPAIRE ................................................................ 71 

Non-indigenous species of the Belgian part of the North Sea and adjacent estuaries 

VLIZ ALIEN SPECIES CONSORTIUM ....................................................................................... 77 

Brussels Psittacidae: impacts, risk assessment and action range 

A. WEISERBS ......................................................................................................................... 81 

Research on biological invasions: a Belgian perspective 

E. Branquart et al. ............................................................................................................... 85

Welcome to the World? Non-native breeding birds in Flanders 

(Belgium) 

Anny ANSELIN*, Glenn VERMEERSCH & Geert SPANOGHE 

Instituut voor Natuur- en Bosonderzoek, Kievitstraat 25, 1070 Brussel. *Corresponding 

author: Anny.Anselin@Inbo.be 

Introduction 

Non-native bird species with their natural breeding range in the Americas, Asia, Africa and 

Australia have become more common in large areas of Europe, and the situation in Flanders 

(Belgium) is no exception to this. Waterbirds constitute an important part of these non-native 

species. In Flanders, the keeping of waterbirds in captivity, in zoos and private collections is 

commonplace and the deliberate and accidental release of full-winged birds has led to the 

development of most of our present feral populations. It is widely accepted that non-native 

species could be a threat to biodiversity and several international legislative instruments, 

including the Convention on Biological Diversity (1992), the Bern Convention (1979), the 

AEWA Agreements (1999) under the Bonn Convention (1983) and the European Birds 

Directive (1979) strengthen the need to control invasive and non-native species that threaten 

that biodiversity. In this contribution, we give a short overview of the status of the most 

important non-native bird species in Flanders, with the exception of the Rose-ringed Parakeet, 

a species mostly concentrated in and around Brussels and for which we refer to Weiserbs 

(2010, present volume). We first present information on population numbers, distribution, 

trends, dispersion potential and known impact, to determine the Invasive Species 

Environmental Impact Assessment scores (ISEIA scores) of the various species, but to show 

also possible economic and public health hazards. We give a short summary of the existing 

legislation and management measures, and end with some guidelines for future actions and 

needs. 

Assessment of environmental and other hazards 

Population numbers, trends and distribution 

For most of the species good information is available on historical presence, long-time 

breeding population numbers (Vermeersch & Anselin, 2009), wintering numbers (Devos 

2009, in prep) and distribution (Vermeersch et al, 2004). At present, in 70% of all 5×5 km 

UTM squares in Flanders at least one non-native species is breeding. Three species, Canada, 

Egyptian and Barnacle goose occupy 93% of all squares and count for 92% of the total 

breeding population of all non-native species. Wintering trends are in general very similar to 

breeding trends. 

The Canada Goose, Branta canadensis is the most successful non-native waterbird 

species in Flanders. Numbers increased from 1 breeding pair in 1973 to the present population 

of 1600-2000 breeding pairs! The species is now widely distributed and is particularly 

abundant in the central part and along the Scheldt and Leie river valleys. Areas with high 

numbers coincide with the (former) presence of wildfowl collections. Numbers of the 

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Egyptian Goose, Alopochen aegyptiacus have increased markedly from a few pairs in 1973- 

1977 (Brussels) to the present population of 800-1500 pairs. The species is now widely 

distributed but is more abundant in the central and eastern part of Flanders. Squares with more 

than 10 breeding pairs are frequent and local densities can reach 35 pairs/square. Breeding 

birds in eastern Limburg most probably originate from adjacent Dutch populations. In certain 

areas growth rates are very high. Although the feral Barnacle Goose, Branta leucopsis has a 

smaller breeding population (120-150 bp) and a less extensive range than the two former 

species, numbers are increasing rapidly. The first free-living birds were observed in the 

1960s, and the first free-breeding pair was only recorded in 1992 in the Campine region, but 

the species is now distributed in the whole of Flanders and occupies 10% of all atlas squares. 

The Bar-headed Goose, Anser indicus is a scarce breeding species (20-25 bp) over the whole 

of Flanders and shows a very slow increase. It has been observed in the wild since the 1960s 

and the first breeding record dates from 1989. The Magellan Goose, Cloephaga picta has 50- 

60 breeding pairs of which the majority is present in one area in the province of East- 

Flanders. These birds originate from a nearby waterfowl collection and started breeding in the 

wild in 1993. The population has slowly increased. Free-living birds have been reported since 

the seventies, but numbers have always remained low. The feral Black Swan, Cygnus atratus 

is a rather scarce breeding species with 40-45 pairs (5% of the atlas squares) but shows a 

steady increase. In most squares only one breeding pair is recorded. They have been reported 

since the 1970s. Observations increased in the 1990s and the first free-breeding pair was 

noted in 1998 north of Brussels. The population of the Mandarin Duck, Aix galericulata is 

estimated at 80-95 pairs. It occurs over the whole of Flanders but is scarce in the western part. 

Free-living birds have been reported since the 1950s, but observations have become more 

frequent since the 1980s. The first pair was noted near Brussels in 1987. The species seems to 

increase slowly. The Wood Duck, Aix sponsa occupies only 3.8 % of the atlas squares and has 

a population of 25-30 breeding pairs, which remains very low. The species occurs less in the 

western part. The first free-living birds were observed in 1957, and the first breeding record 

was noted in 1982. The Ruddy Shelduck, Tadorna ferruginea is a very scarce and irregular 

breeding species (5-10 bp). Breeding pairs are probably all escaped birds although wild birds 

could arrive by natural dispersion. The first breeding record dates from 1981 and since, only a 

few breeding attempts have been reported. In 2008 two pairs of the Ruddy Duck, Oxyura 

jamaicensis bred for the first time in Flanders and several young were observed. In 2009 at 

least 4 pairs (maybe 5) pairs are present and the chances are high that the population will 

quickly extend if no measures are taken (obs. Geert Spanoghe). 

Dispersal 

To gather more information on the spatial and seasonal distribution and dispersion of the 

Canada Goose in Flanders a neck-banding project was started in 1995 in cooperation with the 

Belgian ringing service (Anselin et al, 1996) and later also in the southern part of the country. 

Between 1994 and 2003 a 200 neck bands and 69 leg rings were fitted on adult and immature 

birds in 17 locations. Analysis of the 4600 controls show that there is a lot of exchange 

between locations (and regions) but birds move mostly no further than 20-25 km and usually 

a radius of 50 km is not exceeded. Movements over several hundreds of kilometres seem to be 

an exception. With a species distributed over large parts of Belgium and neighbouring zones 

and showing a high level of exchange this means that to control numbers in an efficient way, 

the actions need to be organized on a wide scale (Cooleman et al, 2005). Colour ring 

programs for the Barnacle Goose in the Netherlands (Van der Jeugd, 2005) and a similar 

program in Brussels for the Egyptian Goose (Vangeluwe & Roggeman, 2000) show that in

these geese species exchange between populations is also frequent, which demonstrates again 

the need for wide scale, and preferably, international actions. 

Impact 

Most likely environmental problems with non-native waterbirds arise from hybridisation with 

closely related species, previously separated by geographical barriers. Other environmental 

hazards are aggression towards other species, damage to vulnerable habitats by grazing and 

trampling and deposition of nutrients by roosting birds leading to eutrophication (geese). 

Apart from this, economic damage can be caused by large flocks of geese species to crops and 

grasslands by year-round grazing and trampling. Other impacts on man are damage to 

amenity areas, threats to public health in parks and water areas and threats to air safety 

(collisions with aircraft)(Owen et al, 2006). In Flanders, damage on vulnerable habitats in 

nature reserves by overgrazing, trampling and water eutrophication by the three most 

common non-native geese species is frequently reported and is most certainly a major 

problem, but no studies have been carried out on the effects. Damage in agricultural habitats 

occurs but has not been quantified. Aggressive behaviour of Canada and Egyptian geese 

towards other species has been reported, but again very few is known about magnitude and 

effects (Beck & Anselin, 2005). On the other hand, the problem of interbreeding of the North 

American Ruddy Duck and the rare European White-headed Duck, Oxyura leucocephala (an 

Annex I species of the EU Birds Directive) is well documented and guidelines for actions are 

presented in an international action plan (Hughes et al. 2006). Now that the Ruddy Duck has 

recently started breeding in Flanders, special attention must be paid to the species and actions 

undertaken to halt this development. 

Species Breeding 

pop/trend 

Wintering 

pop/trend 

Dispers Impact 

habitat 

Impact 

species 

ISEIA score 

category 

Branta canadensis ? Black 

Alopochen aegyptiacus ? Watch>Black? 

Branta leucopsis ? Not>Black? 

Anser indicus ? Watch 

Cloephaga picta ? 

Cygnus atratus ? 

Aix galericulata ? Watch 

Aix sponsa ? 

Tadorna ferruginea ? 

Oxyura jamaicensis Alert>Black 

Table 1: The various risk factors in three broad categories from light grey (low or medium/less 

important) to dark (high/more important), together with the proposed changes in ISEIA category for 

some species (in bold) 

3 

Impact 

human 

In Table 1 we can distinguish two main groups. First, the three most common geese, (now all 

proposed for Black) and the Ruddy Duck (Alert), now proposed for Black taken into account 

the recently changed status of the species in Flanders. The second group consist on two 

‘Watch’ species, the Bar-headed Goose and the Mandarin Duck and several not categorised 

species, thus generally spoken, the species with (at present) a lower environmental impact.

4 

Legislation and Management 

Apart from the international actions in the framework of the Ruddy Duck interbreeding 

problem, only the management of the Canada Goose is legally regulated. Ruddy Shelduck and 

Barnacle Goose are special cases. The first has wild breeding populations within the EU (in 

the Black Sea region) and is as such not considered as a ‘non-native’ species, although feral 

populations mostly origin from escapees. The Barnacle Goose figures on the Annex I of the 

EU Birds’ Directive and the unclear mix of populations with birds from escaped and wild 

origin outside their traditional Scandinavian breeding grounds complicates actions to prevent 

further growth of these populations (in certain regions in major part originating from 

escapees). All other non-native species are ‘outlawed’ and can be ‘destroyed’ by all means. 

The management of Canada Goose populations consists of hunting, destroying of eggs and 

rounding up and killing of flightless birds during moult. Bag statistics are gradually better 

documented (Scheppers & Casaer, 2008). Statistics on other measures exist but are still less 

accessible. 

Future actions and needs 

To be able to evaluate management actions in an efficient way, the existing monitoring 

systems have to be maintained and developed or adapted where necessary. Non-native species 

form part of a regional environmental indicator and therefore it is important to have sound 

data. There is a need to start thinking about developing structural and legally regulated 

management actions for the Egyptian Goose and to find a solution for the Barnacle Goose 

‘problem’. Intervention actions on these species should be coordinated and organised on a 

much wider (international) scale. Special attention must be paid to the Ruddy Duck and 

actions undertaken to halt its present population development. We strongly advice to support 

or start studies on damage impact (crops, vulnerable vegetations, species interactions) as 

sound quantitative information is largely lacking, and bring together all non- published 

knowledge in a comprehensive review. And last but not least we simply propose to apply 

more extensively the AEWA Conservation Guidelines on Avoidance of Introductions of Nonnative 

Waterbird Species (Owen et al, 2006).

References 

Anselin A., Geers V. & Kuijken E., 1996. Population trends and ecology of the Canada Goose 

Branta canadensis in Flanders. In: Holmes J. & Simons J. The introduction and 

naturalisation of birds. HSMO, London 71-72. 

Beck O. & Anselin A., 2005. Management of feral geese populations in Flanders (in Dutch). 

Beheer van verwilderde ganzenpopulaties in Vlaanderen. Natuur.oriolus 71:166-169. 

Cooleman S., Anselin A., Beck O., Kuijken E. & Lens L., 2005. Movements and mortality of 

Canada Goose Branta canadensis in Flanders (in Dutch). Verplaatsingen en mortaliteit 

van de Canadese Gans Branta canadensis in Vlaanderen. Natuur.oriolus 71:152-160. 

Devos K., 2009. Wintering waterbirds in Flanders. Numbers, distribution and trends in the 

period 1991/92 – 2008/09 (in Dutch). Overwinterende watervogels in Vlaanderen. 

Aantallen, verspreiding en trends in de periode 1991/92 – 2008/09. Mededeling INBO, 

Brussel (in prep). 

Hughes B., Robinson J., Green A., Li D. & Munkur T., 2006. International Single Species 

Action Plan for the Conservation of the White-headed Duck Oxyura leucocephala. CMS 

Technical Series 13, AEWA Technical Series 8. Bonn. 

Owen M., Callaghan D. & Kirby J., 2006. Guidelines on Avoidance of Introductions of Nonnative 

Waterbird Species. AEWA Technical Series N° 12. Bonn. 

Scheppers T. & Casaer J., 2008. Hunting Management Units: statistics, report and analysis of 

the period 1998-2007.(in Dutch). Mededelingen INBO nr 9, Brussel. 

Van der Jeugd H., 2005. Barnacle Geese Branta leucopsis on the move (in Ditch). 

Brandganzen Branta leucopsis volop in beweging! Natuur.oriolus 71:161-164. 

Vangeluwe, D. & Roggeman W., 2000. Evolution, structure et gestion des rassemblements 

d’Ouettes d’Egypte Férales en Région Bruxelles-Capitale. Rapport IRSNB, Brussel. 

Vermeersch G. & Anselin A., 2009. Breeding birds in Flanders 2006-2007 (in Dutch). 

Broedvogels in Vlaanderen 2006-2007. Mededelingen INBO nr. 3, Brussel). 

Vermeersch G., Anselin A., Devos K., Herremans M., Stevens J., Gabriëls J. & Van Der 

Krieken B., 2004. Atlas of the breeding birds in Flanders (in Dutch). Atlas van de 

Vlaamse Broedvogels 2000-2002. Mededelingen IN nr. 23, Brussel. 

Weiserbs, A., 2010. Brussels Psittacidae: impacts, risk assessment and action range. Present 

volume, 77-79. 

5

Alien macro-crustaceans in freshwater ecosystems in Flanders 

Pieter BOETS * , Koen LOCK & Peter L. M. GOETHALS 

Laboratory of Environmental Toxicology and Aquatic Ecology, Ghent University, J. 

Plateaustraat 22, B-9000 Ghent, Belgium. *Corresponding author: pieter.boets@ugent.be 

Introduction 

The introduction of invasive species has increased enormously during the past decades. 

Currently, eighteen alien macro-crustaceans have been found in freshwater ecosystems in 

Flanders. One of these invasive species, which is already widely distributed throughout 

Europe, is Dikerogammarus villosus (Sowinsky, 1894)(Bollache et al., 2004). Since 1997, D. 

villosus has been found in Flemish watercourses (Messiaen et al., unpublished data). D. 

villosus was first observed in the Albert canal, while nowadays, it also occurs in other canals 

and other stagnant and running watercourses in Flanders. Lab experiments have proven useful 

to determine the impact of D. villosus on a microcosm scale (Dick and Platvoet, 2000). A 

problem with these experiments is to translate the obtained results to field situations. 

Therefore, it can be useful to combine lab experiments with field observations and data-driven 

models. This study aims to identify the most important variables determining the habitat 

suitability of D. villosus using decision trees. In addition, lab experiments were conducted to 

gain insight in the behaviour of D. villosus. In this way, models based on field observations 

can be used in combination with lab experiments to make useful predictions about the impact 

of the invasive species D. villosus on native macroinvertebrate communities. 

Material and methods 

Multiple as well as single prey experiments were conducted in glass aquaria filled with five 

litres of carbon filtered water. In the multiple prey experiments, five individuals of the 

predator Dikerogammarus villosus were released in the aquaria containing four individuals of 

five different prey (Asellus aquaticus, Crangonyx pseudogracilis, Gammarus pulex, Cloeon 

dipterum and Chironomus species). Single species experiments with A. aquaticus, C. 

pseudogracilis, G. pulex, G. tigrinus or C. dipterum as prey were also conducted to study the 

interaction between prey and predator in the absence of other macroinvertebrates. In these 

experiments, five individuals of one prey species were exposed to five individuals of D. 

villosus. All experiments lasted 24 h after which the survival of the macroinvertebrates was 

checked. To check the influence of the substrate on predation, all experiments were conducted 

three times: once on gravel, once on sand and once without substrate. All predator-prey 

experiments were replicated five times. 

To determine the substrate preference of the invasive D. villosus and the native G. 

pulex, 10 individuals of each one species were released in an aquarium filled with 10 litres of 

carbon filtered water. To assess if substrate preference changed if prey and predator occurred 

together, an additional experiment was conducted, where 10 individuals of both species were 

put together in one aquarium. The substrate preference was checked after 24 hours. 

The dataset used to model the habitat suitability is based on the samples collected by 

the Flemish Environment Agency, which monitors a large number of sampling points 

scattered over the different stagnant and running water systems in Flanders. Environmental 

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variables, hydro-morphological characteristics and data of other macroinvertebrates were 

available. In total, 232 presence or absence data were available with information on the 

hydro-morphological characteristics and physical-chemical variables. In 145 samples, D. 

villosus was absent, while in 87 other samples, the species was present. For the decision tree 

construction, the machine learning package WEKA – J 48 algorithm (Witten and Frank, 

2000) was used. 

Results 

The multiple prey experiment showed a strong predation on all macroinvertebrates in the 

presence of D. villosus (Figure 1), while all individuals of D. villosus survived. The highest 

survival of prey was usually found with gravel as substrate, except for G. pulex, which had 

the highest survival with sand as substrate. There was a significant difference in survival of 

the different prey (p

Discussion 

Results of the predator-prey experiments showed a similar predatory behaviour of D. villosus 

compared to the results of previous studies (Dick and Platvoet, 2000; Krisp and Maier, 2005). 

Not only native species, but also the exotic species G. tigrinus and C. pseudogracilis 

originating from North America were predated. In this way, D. villosus not only has an 

influence on native fauna, but also on exotic species, as was already observed in the river 

Rhine and the river Meuse (Josens et al., 2005; Van Riel et al., 2006). In the presence of D. 

villosus, a general decline in macroinvertebrate diversity and abundance in natural systems 

was observed (Van Riel et al., 2003). Despite its predatory behaviour, D. villosus should, 

however, not be seen as a strict carnivore: studies conducted by Platvoet et al. (2005), 

Maazouzi et al. (2007) and Mayer et al. (2008) showed that D. villosus is an omnivorous 

species able to eat plant as well as animal material. This diverse food spectrum probably 

contributes to the successful spread of this species. 

The substrate preference experiment pointed out that D. villosus preferred gravel 

substrate, as was also found in previous laboratory studies (Van Riel et al., 2003; Kley and 

Maier, 2006). Studies conducted in the Moselle river (France) indicated that D. villosus was 

present on different types of substrates (Devin et al., 2003), however, there was a difference in 

preference based on the age and the size of the species. Juveniles were more often present 

between roots and macrophytes whereas adults had a preference for boulders and stones. 

The developed habitat suitability model indicated that watercourses with an artificial 

bank structure, a high oxygen saturation and a low conductivity were preferred by D. villosus. 

D. villosus can thus invade artificial watercourses, however, the water quality regarding 

oxygen content and conductivity has to be good. The species avoided watercourses with a 

good biological water quality, which possibly means that natural systems with a high diversity 

of macroinvertebrates are more resistant to invasions than watercourses with a low diversity. 

However, according to Bollache et al. (2004), it is possible that in the near future, whole 

drainages of natural and semi-natural rivers, can be invaded by this species. Therefore, 

continuous monitoring of invasive species remains necessary. 

Acknowledgements 

Dr. Koen Lock is currently supported by a post-doctoral fellowship from the Fund for Scientific 

Research (FWO-Vlaanderen, Belgium). 

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References 

Bollache, L., Devin, S., Wattier, R., Chovet, M., Beisel, J-N., Moreteau, J-C., Rigaud, T., 

2004. Rapid range extension of the Ponto-Caspian amphipod Dikerogammarus villosus in 

France: potential consequences. Arch. Hydrobiol. 160, 57-66. 

Devin, S., Piscart, C., Beisel, J.N., Moreteau, J-C., 2003. Ecological traits of the amphipod 

invader Dikerogammarus villosus on a mesohabitat scale. Arch. Hydrobiol. 158, 43-56. 

Dick, J.T.A., Platvoet, D., 2000. Invading predatory crustacean Dikerogammarus villosus 

eliminates both native and exotic species. Proc. R. Soc. Lond. B. 267, 977–983. 

Josens, G., Bij de Vaate, A., Usseglio-Polatera, P., Cammaerts, R., Cherot, F., Grisez, F., 

Verboonen, P., Vanden Bossche, J-P., 2005. Native and exotic Amphipoda and other 

Peracarida in the River Meuse: new assemblages emerge from a fast changing fauna. 

Hydrobiologia 542, 203-220. 

Kley, A., Maier, G., 2006. Reproductive characteristics of invasive gammarids in the Rhine- 

Maine-Danube catchment South Germany. Limnologica 36, 79-90. 

Krisp, H., Maier, G., 2005. Consumption of macroinvertebrates by invasive and native 

gammarids: a comparison. J. Limnology 64, 55-59. 

Van Riel, M.C., Van der Velde, G., Bij de Vaate, A., 2003. Alien amphipod invasions in the 

river Rhine due to river connectivity: a case of competition and mutual predation. In 

Douben, N., Van Os, A.G., (eds). Proceedings NCR-days (2003); Dealing With Floods 

Within Constraints. NCR publication 24. Netherlands Centre for River Studies, Delft, 51– 

53.

ISEIA, a Belgian non-native species assessment protocol 

Etienne BRANQUART 1 , Hugo VERREYCKEN 2 , Sonia VANDERHOEVEN 3 , Fabienne VAN ROSSUM 4 

and other members of the Belgian Forum on Invasive Species. 

1 Belgian Biodiversity Platform, 

2 Instituut voor Natuur- en Bosonderzoek, 

3 Laboratory of Ecology, Gembloux Agro-Bio Tech, Université de Liege, 

4 National Botanic Garden of Belgium 

Introduction 

Belgian land managers and policy makers have to face up to an increasing number of nonnative 

species with contrasted impacts on the environment. To help them in the identification 

of species of most concern for preventive or mitigation actions, Harmonia, an information 

system on invasive species in Belgium, has recently been developed at the initiative of 

scientists gathered within the Belgian Forum on Invasive Species (http://ias.biodiversity.be). 

Harmonia is based on a standardised assessment protocol (ISEIA) which allows 

assessing, categorising and listing of non-native species from any taxonomic group according 

to their invasion stage in Belgium and to their impact on native species and ecosystem 

functions (Branquart 2007). The ISEIA protocol is one of the first national standardised risk 

assessment tools developed for non-native species in Europe (Essl et al., submitted). 

Here we present the ISEIA protocol, the assessment procedure and the results of the 

first assessments performed on vascular plant and vertebrate species in Belgium. The way 

those results may be used to develop regulatory instruments and management guidelines are 

also briefly discussed. 

The ISEIA protocol and the Belgian list system 

The ISEIA protocol aims at categorising non-native species on the basis of a standardised 

methodology designed to minimise the use of subjective opinions and to make the process of 

assessing and listing invasive species transparent and repeatable. 

Contrary to predictive pest risk assessment protocols mainly based on species' intrinsic 

attributes for evaluating invasion likelihood and potential to cause adverse ecological effects, 

the ISEIA approach favours the use of invasion histories documented in peer-reviewed 

publications and in scientific reports from Belgium and neighbouring areas. It is considered 

that non-native species are likely to cause significant impacts on native species and 

ecosystems in Belgium if they have already done so in neighbouring countries. The reference 

area taken into consideration for the assessment includes the European regions with ecoclimatic 

conditions comparable to Belgium, i.e. hardiness zones 7 and 8 characterised by an 

average annual minimum temperature between –7 and –17°C (Cathey 1990). It covers 

Denmark, the Netherlands and large parts of Germany, France, Ireland, Switzerland and the 

United Kingdom (figure 1). 

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Figure 1 – USDA hardiness zones in 

western Europe based on the ability of a 

species to withstand the minimum 

temperatures of the zone (Cathey 1990). 

The reference area used in the ISEIA 

protocol covers hardiness zones 7 and 8. 

The Belgian list system is based on three different list categories as proposed in the European 

strategy on Invasive Alien Species (Genovesi & Shine 2003). Those categories are defined 

according to the severity of impacts on the environment: no negative impact (white list), 

negative impact suspected (grey list) and negative impact confirmed (black list). The 

assignment of a non-native species to one of those categories is assessed by four main criteria 

matching the last steps of the invasion process, i.e. potential for spread, colonisation of natural 

habitats and adverse ecological impacts on native species and ecosystems. Consistent with 

other risk assessment standards, equal weight is assigned to each of the four criteria and a 

three-point scale is used for criteria scoring: low (or unlikely), medium (or likely) and high. 

The global ISEIA score is calculated as the sum of risk rating scores of the four criteria (see 

Branquart 2007 for additional explanation). 

Non-native species are allocated to the different categories of the Belgian list system 

combining information from the ISEIA scoring and data on their invasion stage in the country 

(figure 2). Detrimental species not yet established in Belgium (A0) or occurring only in a few 

localities (A1) are to be considered as a high priority for prevention and eradication actions 

(Genovesi & Shine 2003, Genovesi et al. 2009). 

Figure 2 - List system proposed to identify 

non-native species of most concern for 

preventive and mitigation actions in Belgium 

(Branquart 2007).

The assessment procedure 

Five different expert working groups were established to deal with vascular plants, fishes, 

amphibians, birds and mammals, each of them including three to six scientists from different 

research institutes and universities in Belgium. In a first step, the environmental impact of 

each species is assessed independently by the different experts, sometimes leading to 

diverging results. At last, these results are discussed during a working group meeting in order 

to make sure that experts share a common understanding of criteria and definitions and to 

search for a robust consensus for each species. 

The ISEIA protocol has been improved several times based on those discussions. It 

has proven to be flexible enough to be used to assess the environmental impact of non-native 

species from very different taxonomic groups. The major difficulty encountered during the 

assessment process was that environmental impacts of non-native species in the reference area 

are often poorly documented in the literature. This is typically the case for most species 

included in the watch list (e.g. Acer rufinerve, Alopochen aegyptiacus, Ameiurus spp., 

Cyperus eragrostis, Lepomis gibbosus and Tamias sibiricus). 

Up to now, assessments were focused on species considered as detrimental in at least 

one country of the reference area (e.g. according to Pascal et al. 2003, Muller 2004, Weber et 

al. 2005, Wittenberg et al. 2006, Copp et al. 2008). 

Results and trends 

Nearly 400 non-native species of vascular plants and vertebrates can be considered as 

currently established in Belgium. So far, 57 non-native vascular plant (neophytes) and 32 

vertebrate species have been evaluated by Belgian experts using the ISEIA protocol, 

comprising 72 species considered as naturalised in Belgium. Thirty-nine out of these 72 

species were assessed as organisms with a strong detrimental impact on native biodiversity 

(black list species, A1-A3), for which preventive and mitigation actions are strongly 

recommended (see the list in appendix). Most of the remaining species were recorded on the 

watch list (B1-B3), which means either that their impact on native biodiversity is moderate or 

that their impact is still unclear due to a deficiency in scientific studies. 

Compared to vascular plant species, a higher proportion of the vertebrate species has 

been shown to be detrimental to native species and ecosystem functions. Invasive neophytes 

typically affect biodiversity by growing in very dense populations in semi-natural habitats, 

outcompeting native plant species and modifying the vegetation structure. Invasive vertebrate 

species rather adversely impact biodiversity through a wide range of interspecific interactions 

(competition, predation, disease transmission and hybridisation) that may act separately or 

synergistically (Harmonia database 2009). 

Besides the evaluation of 72 non-native species naturalised in Belgium, 17 taxa 

established in neighbouring countries but not (yet) in Belgium were assessed and included in 

the alert list (A0-B0) (see appendix). All of them are likely to become established in the 

coming years if no preventive action is undertaken. 

13

14 

40 

35 

30 

25 

20 

15 

10 

5 

0 

# of black list species 

< 1825 < 1875 < 1925 < 1975 < 2009 

Time 

Figure 3A – Number of black list species 

established in the wild in Belgium at different 

periods of time. Data: Harmonia database 2009. 

29 

31 

# of black list species 

20 25 30 35 

34 

34 

Figure 3B – Number of black list species within 

the main biogeographic regions of Belgium Data: 

Harmonia database 2009. 

When examining the change in number of invasive species for 200 years (Harmonia database 

2009), we can see that there is an increasing number of established black-list species in 

Belgium causing major ecological damages and growing management costs (figure 3A). 

Those species tend to concentrate in areas where human density and activities are at the 

highest (figure 3B) and where habitats incur frequent alteration, eutrophication and pollution, 

and therefore may be prone to invasion. Many of those black-list species thrive along river 

banks and in freshwater environments. 

From science to management 

Eighteen non-native species out of the group of organisms responsible for high environmental 

impacts are either not yet established in Belgium (A0) or only known from a limited number 

of localities (A1) (see appendix). Prevention actions and early eradication of these species 

deserve to be conducted in priority. Indeed, the ecological damages they may cause can still 

be restricted to a minimum at a low cost if actions are undertaken without delay. This is the 

reason why these 18 species have been proposed to be considered in a new Royal Decree 

(expected in early 2010) aiming at restricting their importation, exportation and rearing. 

Another 30 detrimental species are already widely distributed in Belgium (A2 & A3) 

and cannot be eradicated anymore. However, it is still worthwhile avoiding further secondary 

releases to slow down the invasion process. Voluntary codes of conduct developed in 

partnership with key sectors of activity (horticulture, pet industry, etc.) may help to reduce the 

introduction of these species in the wild (Branquart & Halford 2009). In addition to 

preventive actions, the monitoring and the management of those species is strongly 

recommended in areas of high conservation value in order to preserve native red-listed species 

and threatened habitats (Tu 2009). 

At last, it is now widely acknowledged that early warning and rapid response are 

crucial for mitigating the impacts caused by biological invasions in Europe. Early warning 

tools should ideally be developed through information exchange at a regional scale and need a 

strong international scientific collaboration, a common understanding of invasiveness issues 

and shared risk assessment schemes (Genovesi et al. 2009, Hulme et al. 2009). A standardised 

ISEIA-like protocol deserves to be developed at a European scale to reach that goal. 

29 

23 

21


The following scientists (listed by taxonomic group and in alphabetical order) have participated to the 

impact assessment of non-native vascular plant and vertebrate species: Iris Stiers, Ludwig Triest, 

Sonia Vanderhoeven (FNRS), Wouter Van Landuyt, Fabienne Van Rossum, Filip Verloove (vascular 

plants); Dieter Anseeuw, François Lieffrig, Jean-Claude Micha, Denis Parkinson, Hugo Verreycken 

(fishes); Arnaud Laudelout, Gérald Louette, Youri Martin, Joachim Mergeay, Chistiane Percsy 

(amphibians); Anny Anselin, Diederik Strubbe, Anne Weiserbs (birds); Margo D’aes, Alain Licoppe, 

Grégory Motte, Vinciane Schockert, Jan Stuyck (mammals); Etienne Branquart (coordination). 

References 

Branquart E. (Ed.), 2007. Guidelines for environmental impact assessment and list 

classification of non-native organisms in Belgium (available from 

http://ias.biodiversity.be). 

Branquart E. & Halford M., 2009. Increase awareness to curb horticultural introductions of 

invasive plants in Belgium (InvHorti). EPPO/Council of Europe Workshop: 'Code of 

conduct on horticulture and invasive alien plants', Ski, 4-5 June 2009. 

Cathey H.M., 1990. USDA plant hardiness zone map. USDA, Washington D.C. 

Copp G.H., Vilizzi L., Mumford J., Fenwick G.V., Godard M.J. & Gozlan R.E., 2008. 

Calibration of FISK, an invasiveness screening tool for non-native freshwater fishes. Risk 

Analysis 29: 457-467. 

Essl F., Klingenstein F., Milasowszky N., Nehring S., Otto C. & Rabitsch W., submitted. The 

German-Austrian black list information system (GABLIS) : a tool for assessing 

biodiversity risks of invasive alien species in Europe. Journal of Nature Conservation. 

Genovesi P. & Shine C., 2003. European strategy on invasive alien species. Convention on 

the Conservation of European Wildlife and Natural Habitats. Council of Europe 

Strasbourg, T-PVS, 60 pp. 

Genovesi P., Scalera R., Solarz W. & Roy D., 2009. Towards an early warning and 

information system for invasive alien species threatening biodiversity in Europe. Extensive 

executive summary, contract n° 3606/B2008/EEA.53386. 

Harmonia database, 2009. Belgian Forum on Invasive Species, last accessed on 05 May 2009 

from: http://ias.biodiversity.be . 

Hulme P.E., Pysek P., Nentwig W. & Vila M., 2009. Will threat of biological invasions unite 

the European Union? Science 324: 40-41. 

Muller S., 2004. Plantes invasives en France: état des connaissances et propositions d'actions. 

Publication scientifique du Museum d'Histoire naturelle, Patrimoines naturels n°62, 168 p. 

Pascal M., Lorvelec O.L., Vigne J.D., Keith P. & Clergeau P., 2003. Evolution holocène de la 

faune de vertébrés de France: invasions et extinctions. INRA, CNRS, MNHN & MEDD. 

Tu M., 2009. Assessing and Managing Invasive Species within Protected Areas. Protected 

Area Quick Guide Series. Editor, J. Ervin. Arlington, VA. The Nature Conservancy, 40 pp. 

Weber E., Köhler B., Gelpke G., Perrenoud A. & Gigon A., 2005. Schlüssel zur Enteilung 

von Neophyten in der Schweiz in die Schwarze Liste oder die Watch-Liste. Botanica 

Helvetica 115: 169-194. 

Wittenberg R., 2005. An inventory of alien species and their threat to biodiversity and 

economy in Switzerland. CABI Bioscience Switzerland Centre report to the Swiss Agency 

for Environment, Forests and Landscape. The environment in practice no. 0629: 155 pp. 

15

16 

Appendix – List of non-native species with a high detrimental impact on the environment 

Scientific name English name Taxonomic 

group* 

Species not established in Belgium 

Callosciurus finlaysonii Finlayson's squirrel M A0 

Carpobrotus spp. Hottentot fig P A0 

Cervus nippon Sika deer M A0 

Muntiacus reevesi Reeves' muntjac M A0 

Mustela vison American mink M A0 

Neogobius melanostomus Round goby F A0 

Nyctereutes procyonoides Raccoon dog M A0 

Perccottus glenii Rotan, Amur sleeper F A0 

Sciurus carolinensis Grey squirrel M A0 

Threskiornis aethiopica Sacred ibis B A0 

Species with isolated populations in Belgium 

Callosciurus erythraeus Pallas's squirrel M A1 

Crassula helmsii Australian swamp stonecrop P A1 

Egeria densa Brazilian waterweed P A1 

Lagarosiphon major Curly waterweed P A1 

Ludwigia peploides Water primrose P A1 

Myocastor coypus Coypu, Nutria M A1 

Myriophyllum heterophyllum Variable watermilfoil P A1 

Rana catesbeiana American bullfrog A A1 

Species with a restricted distribution in Belgium 

Acer negundo Box-elder, Ash-leaved maple P A2 

Ailanthus altissima Tree of heaven P A2 

Baccharis halimifolia Eastern baccharis P A2 

Branta canadensis Canada goose B A2 

Cornus sericea Red-osier dogwood P A2 

Cotoneaster horizontalis Rockspray P A2 

Hydrocotyle ranunculoides Water pennywort P A2 

Ludwigia grandiflora Water primrose P A2 

Myriophyllum aquaticum Parrotfeather P A2 

Pelophylax ridibundus Marsh frog A A2 

Persicaria wallichii Himalayan knotweed P A2 

Procyon lotor Raccoon M A2 

Pseudorasbora parva Topmouth gudgeon F A2 

Rhododendron ponticum Rhododendron P A2 

Rosa rugosa Rugosa rose P A2 

Spiraea spp. Meadowsweet P A2 

List 

category

Appendix (cont’d) – List of non-native species with a high detrimental impact on the 

environment 

Widespread species in Belgium 

Aster americ. North American asters P A3 

Carassius gibelio Prussian carp F A3 

Elodea canadensis Canadian waterweed P A3 

Elodea nuttallii Nuttall's waterweed P A3 

Fallopia japonica Japanese knotweed P A3 

Helianthus tuberosus Jerusalem artichoke P A3 

Heracleum mantegazzianum Giant hogweed P A3 

Impatiens glandulifera Indian balsam P A3 

Mahonia aquifolium Oregon grape P A3 

Ondatra zibethicus Muskrat M A3 

Prunus serotina Black cherry P A3 

Rattus norvegicus Brown rat M A3 

Solidago canadensis Canada goldenrod P A3 

Solidago gigantea Giant goldenrod P A3 

* Taxonomic groups: amphibians (A), birds (B), fishes (F), mammals (M) and vascular 

plants (P). 

17

Impact of Fallopia spp. on ecosystem functioning: Nitrogen and 

organic matter cycling and implicated soil biota 

Nicolas DASSONVILLE 1 , Sylvie DOMKEN 1 , Basile HERPIGNY 1 , Franck POLY 2 & Pierre MEERTS 1 

1 Laboratoire de génétique et écologie végétales, Université Libre de Bruxelles, Boulevard du 

Triomphe, campus de la Plaine CP 244 ; Brussels, Belgium. 

2 Laboratoire d’écologie microbienne (UMR5557), Université Claude Bernard Lyon I, France. 

Introduction 

Fallopia japonica is one of the most invasive alien plant species in NW Europe. Its impact on 

indigenous vegetation is high and well documented, whereas its impact on ecosystem 

processes has been less studied. It considerably increases primary productivity in invaded 

ecosystems (Dassonville et al., 2008). It has also been shown to increase cations and P 

availability in the topsoil by a nutrient uplift mechanism thanks to its very deep rooting 

system (Dassonville et al., 2007). In the ALIEN IMPACT project, we evaluated the impact of 

the species on different steps of nitrogen and organic matter cycling and on implicated soil 

biota. Considering the thick permanent litter layer often observed under clones of Fallopia, it 

was hypothesized that Fallopia slows down litter decomposition and nitrogen cycling rate. 

Results and Discussion 

The litter decomposition dynamic of Fallopia has been assessed using the litterbag technique 

(Dassonville et al., 2009). We measured the decomposition of Fallopia litter (stems and 

leaves) and of native vegetation of the study site (50-50 mixture of Eupatorium cannabinum 

and Calamagrostis epigejos). The litter of Fallopia decomposes much more slowly than that 

of native vegetation (Figure1), probably due to its very high C:N ratio (151 and 72 for 

Fallopia stems and leaves, respectively, against 33 for native litter). On the other hand, the 

decomposition of each litter type was slightly faster in invaded compared to uninvaded plots. 

This could be an effect of the moister microclimate under the dense canopy of Fallopia than 

on the soil surface of the uninvaded grassland. 

Remaining mass (%) 

100 

90 

80 

70 

60 

50 

40 

30 

20 

10 

0 

0 50 100 150 200 250 300 350 400 450 

Time (days) 

19 

Figure 1: Decomposition 

kinetics of Fallopia leaves 

(triangles) and stems 

(diamonds) and native litter 

(squares) during one year. 

All litter types were 

incubated in invaded (black) 

and uninvaded (white) 

environment; Decomposition 

is expressed as the 

percentage of initial mass 

lost. Values are means ± 

standard deviation.

20 

The evolution of the nitrogen stock in the litterbags was followed (Figure2). The N 

stock decreases rapidly in litterbags with native litter. This means that this litter easily 

releases N to the soil. On the other hand, N tends to accumulate over time in the litterbags 

containing Fallopia. This suggests that microorganisms living on the decomposing litter have 

to use mineral N from the soil to compensate for the low N concentration of their substrate. 

This leads to fixation of mineral nitrogen into organic matter. 

Nitrogen stock (mg) 

60 

50 

40 

30 

20 

10 

0 

0 50 100 150 200 250 300 350 400 450 

Time (days) 

Figure 2: Evolution of the N 

stock (=remaining mass x N 

concentration) in Fallopia 

leaves (triangles) and stems 

(diamonds) and indigenous 

litter (squares) 

decomposition in invaded 

(black) and uninvaded 

(white) environment. Values 

are means ± standard 

deviation. 

From N fluxes measurements in the invaded ecosystem, it has been found that the 

internal cycling of N in Fallopia is exceptionally efficient. Indeed, 80 % of the N present in 

aboveground biomass during summer is translocated to the rhizomes before the abscission of 

the leaves (This phenomenon explains the high C:N ratio of the litter). This process allows the 

plant to grow rapidly in spring, independently of soil N mineralization. This contributes to the 

high productivity of the species and to its competitive superiority. 

Fallopia has also been found to decrease the intensity of nitrification and 

denitrification in sites with high nitrification potential. Molecular analyses show that these 

differences in activity were partially explained by a decreased number of 

nitrifying/denitrifying bacteria (assessed by quantitative PCR targeting 

nitrification/denitrification genes). On the other hand, the structure of the soil microbial 

communities does not seem to be altered (PCR-DGGE analyses). Our results suggest a 

potential allelopathic effect of Fallopia on soil microbes. The reduction of nitrification and 

denitrification intensity result in the reduction of N loss from the ecosystem by nitrate 

leaching and NOx emissions. 

From the results mentioned above, it appears that Fallopia has a very economic N 

management, which tends to conserve N in the ecosystem (mineral N fixation on 

decomposing litter, efficient N retranslocation and reduced nitrification and denitrification 

intensity). This could be a key trait explaining the invasive success of the plant. 

Finally, Fallopia also impacts soil fauna. The invertebrate density is 50% lower under 

Fallopia compared to uninvaded grassland. The major groups of the mesofauna (0.2 to 4 mm) 

are similar (springtails, gamasid and oribatid mites) when comparing invaded and uninvaded 

plots. On the other hand, some groups of macrofauna (4 to 80 mm) differed between invaded 

and uninvaded plots. Typical forest taxa like diplopods and isopods or the earthworm 

Lumbricus terrestris are more frequent under Fallopia than under the uninvaded grassland 

vegetation. These taxa are important for litter fragmentation and incorporation to the soil. 

This could explain the higher decomposition rate under Fallopia. On the other hand, ants, 

aphids and the earthworm L. castaneus were totally absent under Fallopia. These taxa are

thermophilous grassland species. Based on these differences, the invaded and uninvaded plots 

are very well separated in the PCA analyses (Figure 3). The first axis of the PCA could 

represent a light and moisture gradient. The changes in soil fauna are thus mainly explained 

by a reduction of food diversity and a change in soil microclimate. 

Fact. 2 : 15.7% 

B 

A 

Fact. 2 : 15.7% 

-0,5 

5 

4 

3 

2 

1 

0 

-1 

-2 

-3 

0,5 

0,0 

Isopods 

Opilliones 

Diplopods 

Diptera 

Actinids 

Molluscs 

Heteroptera 

Chilopods 

Gamasids 

Coleoptera 

Araneids 

Oribatids 

Arthropleon 

Homoptera 

Symphiles 

-0,5 0,0 0,5 1,0 

Fact. 1 : 23.3% 

Sy mphy pleon 

Hymenoptera 

-4 

-4 -2 0 2 4 6 

Fact. 1 : 23.3% 

Figure 3: Principal Component Analysis (PCA). A: Projection of variables (taxonomic groups) on 

PC1 and PC2 for soil fauna but the earthworms. B: Projection of the sampling points from the 

invaded (white) and the uninvaded (black) plots. 

21

22 

References 

Dassonville N., Vanderhoeven S., Domken S., Meerts P., Chapuis-Lardy L., 2009. Impacts of 

Alien Invasive Plants on Soil and Ecosystem Processes in Belgium: Lessons from a 

Multispecies Approach. In Wilcox C.P. and Turpin R.D. (eds). Invasive Species: 

Detection, Impact and Control. Nova publishers. In press. 

Dassonville N., Vanderhoeven S., Vanparys V., Hayez M., Gruber W., Meerts P., 2008. 

Impacts of alien invasive plants on soil nutrients are correlated with initial site conditions 

in NW Europe. Oecologia 157:131-140. 

Dassonville N., Vanderhoeven S., Gruber W., Meerts P., 2007. Invasion by Fallopia japonica 

increases topsoil mineral nutrient concentrations. Ecoscience 14: 230-240.

Soil-dependent growth strategy of invasive plants: empirical 

evidence and model predictions using Carpobrotus edulis as target 

species 

Eduardo DE LA PEÑA & Dries BONTE 

Terrestrial Ecology Unit (TEREC), Department of Biology, Faculty of Sciences, Ghent 

University, K.L. Ledeganckstraat 35, 9000 Gent Belgium. 

Introduction 

Plants experience different soil conditions that may influence several aspects of their biology 

such as nutrient uptake, root competition, growth and even floral display (Mal & Lovett- 

Doust 2005). For clonal plants, the relative investment in sexual reproduction or vegetative 

growth can vary according to the context in which they grow (e.g. invaded soil vs. noninvaded) 

since environmental constrains may result in contrasting dispersal strategies due to 

differential cost-benefit balances (Kot et al., 1996). Carpobrotus edulis (L.) N. E. Br., is 

considered a highly invasive species in coastal areas of Southern Europe because it forms 

dense fast-growing mats that displace the native dune vegetation (Vila et al., 2006). 

Carpobrotus edulis can change drastically the characteristics of the invaded soil and the longterm 

occurrence of the species has been associated with a decrease in pH and increase in 

organic content (Conser in press; D'Antonio & Mahall, 1991). Nevertheless, and in spite of 

the large dense patches of C. edulis formed in invaded areas, there is no information available 

about how these changes in soil may affect the posterior growth and colonization rate of the 

species. The objective of the present study was to evaluate whether the residual effects on soil 

after C. edulis invasion affect the growth plasticity of the species and to model the long-term 

consequences of such growth responses. Our working hypothesis was that soil modification 

introduced by C. edulis leads to plant growth responses oriented to maximize the colonization 

rate of invaded areas. 

Material and Methods 

Using a lab experiment we assessed whether the residual effects on soil caused by the 

invasion of Carpobrotus edulis would affect the vegetative and reproductive traits of the 

species and ultimately the dynamics of establishment. We compared C. edulis performance on 

rhizosphere soil collected under native vegetation in the Quiaios dunes (Portugal) that has 

never been occupied by C. edulis (virgin soil, VS), which was used as a reference situation, 

with the performance of plants on soil collected under monospecific patches in the same 

locality where the species grew vigorously (Carpobrotus rhizosphere soil, CRS) or was 

dying-back after a long period of establishment (CDS). After four months growing C. edulis 

in controlled conditions plants were harvested and different plant-growth related features were 

assessed (e.g. biomass, root length, production of flowers). To understand the long term 

consequences of the observed plant responses, we built up a Monte-Carlo simulation model in 

which we integrated clonal growth and seed dispersal under different soil scenarios (absence 

or presence of residual effects).We modeled the spreading of the plant in a grid of 300*300 

23

24 

grid cells, with one grid equaling 0.3*0.3 m², being the species annual growth rate as assessed 

by Sintes et al. (2007). Each grid cell was characterized by soil type: either virgin (VS), 

occupied (CRS) or previously occupied soil (CDS). We consequently analyzed the rate of 

colonization (or coverage) in a hectare. The rate of colonization is followed for one seed 

entering the center of the landscape for three scenarios: colonization of virgin soil without 

residual effects, colonization of soil where the species is present and colonization of soil 

where C. edulis has been removed as a restoration measure (hence leaving behind formerly 

occupied soil). For each scenario we ran 100 replications and calculated average occupancy 

rates as the number of occupied grid cells/total number of grid cells. Soil status changes 

according to the emerging plant dynamics. Local plant spreading occurs by clonal growth and 

by seed dispersal (D'Antonio 1990). With exception of the initial seedling, we introduced a 

grid cell-extinction rate ε of 0.1 in all situations. Seed dispersal is modeled as a stochastic 

process where seeds are distributed according to a Gaussian distribution N(0,σ² from the 

mother plant in all directions. We chose σ²=10 (i.e. variance of the dispersal function 

approximates 3 meter in a thin-tailed dispersal kernel) in the simulation program because this 

accords with dispersal distances in similarly dispersed plant species (Cain et al. 1998). We did 

not model data on the effective number of produced seeds because earlier studies already 

indicated mass seed production per flower (>1000), with only moderate variance according to 

different environmental parameters (Suehs et al. 2004) and extremely low survival and 

germination rates (D'Antonio et al. 1993). Instead, we decided to emphasize on relative 

differences according to the differences in flower production (number of flowers per grid ~ 

number of flowers per plant in the experiment), which consequently would determine the 

number of seeds. Annual clonal growth in C. edulis follows fractal rules concentrically from a 

central branching node and is strongly related to biomass (Sintes et al. 2007). Because annual 

growth rates are estimated ~ 0.3 meter/year (Sintes et al. 2007) we allowed C. edulis to spread 

clonally one grid cell a year in the four direct neighboring grid cells. This corresponds with 

the star-shaped growth at maturity (Wisura 1993). Clonal growth was modeled as a decreased 

probability of 0.2 to colonize each of the adjacent cells based on the results of biomass 

obtained in the lab experiment with different rhizosphere soils (data not shown). 

Results 

Experiment 

The type of soil in which plants were grown affected the production of flowers (Figure 1A). 

In that sense, C. edulis plants growing on soil collected from the native plant community (VS) 

produced a greater number of flowers (P=0.001, F2,35=4.026 Figure 1A) than plants growing 

on soil collected from the two C. edulis patches (i.e. CRS, CDS). The average number of 

flowers produced in plants growing on virgin soil was 1.8 whereas for the other two types of 

soil, CRS and CDS, plants presented only 0.4 and 0.6 flowers respectively. 

Model 

The outcome of our simulations indicated that the local spreading of C. edulis is faster in the 

scenarios where there are no residual effects on soil (Figure 1B). Therefore, the fastest 

covering rate was observed in the model compatible with the responses observed 

experimentally in which the production of flowers resulted in a higher number of seeds in 

virgin soils. This is also reflected by the higher slopes of the fitted logistic function

Nr. of flowers/plant 

3.0 

2.5 

2.0 

1.5 

1.0 

0.5 

0.0 

A 

* 

VS CRS CDS 

Soil type 

Occupancy rate 

1.0 

0.8 

0.6 

0.4 

0.2 

No residual effect 

Residual effect 

Re-establishment 

Time (Years) 

25 

0.0 

0 5 10 15 20 25 30 

Figure 1. Mean number of flowers (A) for Carpobrotus edulis plants growing on Virgin soil (VS); C. 

edulis rhizosphere soil (CRS); C. edulis rhizosphere soil from dying mats (CDS). Error bars indicate 

±SE.* indicate significant difference after One-Way ANOVA. (B) Occupancy rates of C. edulis in a 

virtual landscape of 1ha under three different soil scenarios: no residual effect on flower production 

and growth rates, residual effect and residual effect after C. edulis removal (re-establishment). 

(respectively r=5.85±0.29SE). By contrast, on already occupied sites (with residual effects) or 

after C. edulis removal the covering rate decreased (r=5.19±0.23SE and r=5.23±0.25SE). 

Discussion 

The phenotypic plasticity of the plant (individual plant responses in function of soil 

characteristics) in combination with the resulting contrasting soil environments has interesting 

consequences in C. edulis invasion. In native areas where the plant is not present, investing in 

the production of flowers is the fastest strategy to occupy a given area. Increasing the number 

of flowers allows for the formation of a seed-bank after one growing season. A percentage of 

those seeds would manage to establish (even assuming particularly low germination rates) 

(D'Antonio, 1990; Vila & D'Antonio, 1998), which would result in a fast short-term spread. 

On the other hand, in patches were C. edulis is already present, investment in vegetative 

growth would increase the competitive potential against the native plant community 

(Otfinowski & Kenkel, 2008). If in those already occupied areas the production of seeds were 

increased, this would result in high kin competition or, alternatively, in less chances of seed 

germination due to the dense mat formed by the mother plant. As our model supports, in such 

situations, it is more advantageous to rely on clonal growth as the main means of dispersal. 

Previous studies have modeled the growth pattern of C. edulis in invaded areas however; in 

those cases the combination of clonal vs. sexual reproduction was not integrated. 

Interestingly, based on field observations and also model predictions those authors could see 

that the growth decreased after approximate twenty years (Sintes et al., 2007). This type of 

pattern is perfectly compatible with the residual effects on formerly occupied soil presented in 

our model, and therefore, rather than dispute it completes previous observations. The key 

result of our paper is that the residual effect on soil produced by the species Carpobrotus 

edulis leads to changes in individual plant responses that are oriented to maximize the 

colonization of invaded areas. In this sense, soil context (invaded vs. virgin) is the factor 

determining the shift from an enhanced flower production in virgin soils, never exposed to the 

exotic species, towards the induction of vegetative growth in previously occupied soils. The 

B

26 

changes observed at individual level as a function of soil-context have a dramatic effect on the 

rates of colonization of invaded landscapes. Our findings reveal a key mechanism to 

understand the invasion dynamics of C. edulis, a species that is a serious threat in the 

Mediterranean region, and, more importantly, illustrate how some invasive species can 

quickly respond to soil heterogeneity to maximize the probability of establishing long-term 

plant populations. 

References 

Cain M. L., Damman H. & A. Muir., 1998. Seed dispersal and the Holocene migration of 

woodland herbs. Ecological Monographs 68:325-347. 

Conser C. & Connor, E.F., in press. Assessing the residual effect of Carpobrotus edulis 

invasion, implication for restoration. Biological Invasions. 

D'Antonio C. 1990. Seed production and dispersal in the non-native, invasive succulent 

Carpobrotus edulis (Aizoaceae) in coastal strand communities of Central California. 

Journal of Applied Ecology 27:693-702. 

D'Antonio C. & Mahall. B. E., 1991. Root profiles and competition between the invasive, 

exotic perennial, Carpobrotus edulis, and two native shrub species in California coastal 

scrub. American Journal of Botany 78:885-894. 

D'Antonio C., Odion D. C. & Tyler C. M., 1993. Invasion of maritime chaparral by the 

introduced succulent Carpobrotus edulis. Oecologia 95:14-21. 

Kot M.,. Lewis M. A & van den Driessche P.. 1996. Dispersal data and the spread of invading 

organisms. Ecology 77:2027-2042. 

Mal T. K. &. Lovett-Doust J., 2005. Phenotypic plasticity in vegetative and reproductive traits 

in an invasive weed, Lythrum salicaria (Lythraceae) in response to soil moisture. 

American Journal of Botany 92:819-825. 

Otfinowski R. & Kenkel N. C., 2008. Clonal integration facilitates the proliferation of smooth 

brome clones invading northern fescue prairies. Plant Ecology 199:235-242. 

Sintes T., Moragues E., Traveset A. & Rita J., 2007. Clonal growth dynamics of the invasive 

Carpobrotus affine acinaciformis in Mediterranean coastal systems: A non-linear model. 

Ecological Modelling 206:110-118. 

Suehs C. M., Affre L. &. Medail F., 2004. Invasion dynamics of two alien Carpobrotus 

(Aizoaceae) taxa on a Mediterranean island: I. Genetic diversity and introgression. 

Heredity 92:31-40. 

Vila M. & D'Antonio C. M., 1998. Fruit choice and seed dispersal of invasive vs. noninvasive 

Carpobrotus (Aizoaceae) in coastal California. Ecology 79:1053-1060. 

Vila M., Tessier M., Suehs C. M., Brundu G., Carta L., Galanidis A., Lambdon P. et al., 2006. 

Local and regional assessments of the impacts of plant invaders on vegetation structure 

and soil properties of Mediterranean islands. Journal of Biogeography 33:853-861. 

Wisura W. & Glen H.F., 1993. The South African species of Carpobrotus 

(Mesembryanthema – Aizoaceae). . Contributions to Bolus Herbarium 15:76–107.

Detection of intraguild predation by Harmonia axyridis on native 

cocccinellids by alkaloids 

Louis HAUTIER 1 , Jean-Claude GRÉGOIRE 2 , Jérôme DE SCHAUWERS 3 , Gilles SAN MARTIN 3,4 , 

Pierre CALLIER 3 , Jean-Pierre JANSEN 1 & Jean-Christophe DE BISEAU 3 

1 Département Lutte biologique et Ressources phytogénétiques, Centre wallon de Recherches 

agronomiques , Rue de Liroux, 2, B-5030 Gembloux, Belgium 

2 Lutte biologique et Ecologie spatiale, Université Libre de Bruxelles, CP 160/12, Av. F.D. 

Roosevelt 50 - 1050 Brussels, Belgium 

3 Eco-Ethologie évolutive, Université Libre de Bruxelles, CP 160/12, Av. F.D. Roosevelt 50 - 

1050 Brussels, Belgium 

4 Present address: Écologie comportementale et conservation de la biodiversité , BDIV , 

Université catholique de Louvain, Croix du Sud 4 B-1348 Louvain-la-Neuve, Belgium 

Introduction 

In laboratory conditions, the invasive ladybird Harmonia axyridis Pallas is well known as an 

intraguild predator of ladybird species (Pell et al. 2008, Ware & Majerus 2008) but also of 

other aphidophages (Koch 2003). However, the real impact of H. axyridis intraguild predation 

(IGP) on native coccinellids is poorly known in natural conditions, where multiple prey 

species occur and where prey has the opportunity to escape. To follow predator - prey 

interactions, several techniques have been developed (Harwood & Obrycki, 2005): analysis of 

food remains by gut dissection (Triltsch, 1999), use of monoclonal antibodies (Hagler, 2006) 

and molecular analysis with detection of species-specific DNA sequences (Hoogendoorn & 

Heimpel, 2001). Molecular techniques have been used successfully to detect IGP in 

coccinellids, however, with some limitations regarding detection time as a result of the rapid 

digestion of prey DNA (Gagnon et al., 2005). A new method to monitor IGP of coccinellids 

in natural conditions is based on alkaloid detection by gas chromatography - mass 

spectrometry (Hautier et al., 2008). As many ladybird species use alkaloids as a chemical 

defence (Daloze et al., 1994), these compounds can be used as predation tracers. To apply this 

approach, we firstly developed a laboratory alkaloid detection technique for H. axyridis 

larvae. Next, the influence of several factors on the detection efficiency was studied under 

laboratory conditions: prey instar, risk of false positives and time of detection. Finally, the 

method was validated by testing it on H. axyridis larvae sampled in potato fields. 

Alkaloid detection 

H. axyridis larvae were crushed in 200 µl of methanol and soaked during 10 minutes. The 

extracts were subsequently filtered on cotton wool and the filtrate was concentrated under 

nitrogen. The residues were dissolved in methanol and an aliquot was analysed by gas 

chromatography - mass spectrometry. Analysis conditions are fully described in Hautier et al. 

(2008). 

27

28 

Laboratory results 

Several alkaloids from native ladybird species: adaline, calvine, precoccinelline, propyleine 

(Figure 1), were unambiguously detected in fourth instar larvae of H. axyridis that had 

ingested one first instar larva of Adalia bipunctata (L.), Adalia decempunctata (L.), Calvia 

quatuordecimguttata (L.), Coccinella septempunctata L. or Propylea quatuordecimpunctata 

(L.), respectively. 

Adaline Calvine Precoccinelline Propyleine 

Figure 1. Various alkaloids detected in H. axyridis according to the coccinellid prey 

consumed. 

The alkaloid quantity of a single first instar larva, as a prey, is the smallest amount of 

exogenous alkaloid that can be fed and detected in H. axyridis; it is inferior to that of a single 

egg (Figure 2). The method is sensitive enough to detect all instars. In addition, when the 

predator larva has had a contact with a prey but without biting, the quantity of alkaloid 

detected is significantly lower (4000x) than when consumption has occurred (Figure 3). The 

risk of false positives is thus very limited. After feeding, the exogenous alkaloid concentration 

in the predator decreases over time. However, when a first instar A. bipunctata larva was 

consumed by fourth instar H. axyridis larvae, the alkaloid was detected in the predator larvae 

during 96 h; traces were still detected later on, when the fourth instar H. axyridis larvae had 

become pupae and adults. Alkaloid traces were found in the exuviae as well (Figure 4). 

Adaline (µg) ± S.E. 

2.5 

2 

1.5 

1 

0.5 

0 

egg 1st instar larva 

(n=5) 

t 6= -3.13; p= 0.020 

Figure 2. Adaline in H. axyridis larvae after consumption of one A. bipunctata egg or one 1st 

instar larva. 

(n=10)

Adaline (µg) ± S.E. 

2.50 

2.00 

1.50 

1.00 

0.50 

0.00 

0.0004 µg ± 0.0007 

t 9= -4.62; p= 0.001 

1.67 µg ± 1.14 

Contact Bite 

(n=10) 

(n=10) 

Figure 3. Adaline in H. axyridis larvae after contact with, or after biting a A. bipunctata 

larva. 

Quantity of adaline (µg) 

5.0 

4.5 

4.0 

3.5 

3.0 

2.5 

2.0 

1.5 

1.0 

0.5 

0.0 

0 24 48 72 96 120 144 168 192 216 240 

Time after ingestion (h) 

R 2 =0.615 

p < 0.0001 

(n =10) (n =10) (n =10) (n = 8) 

pupae 

(n =10) (n =9) (n = 9) 

imago 

exuviae 

Figure 4. Temporal change of adaline measured in H. axyridis larvae after consumption of 

one first instar A. bipunctata. 

Field results 

In 2005, 28 H. axyridis larvae were sampled in potato fields in Belgium and analyzed by GC- 

MS. Nine larvae out of 28 were positive. Three alkaloids were detected: precoccinelline (n=4 

larvae), propyleine (n=4) or adaline and precoccinelline (n=1), indicating that the larvae had 

attacked and consumed native species such as C. septempunctata, P. quatuordecimpunctata 

and A. bipunctata. 

Conclusion 

29

30 

In conclusion, using alkaloids as predation tracers allows to follow IGP interactions between 

H. axyridis and native coccinellids under field field conditions. This new detection technique 

is very sensitive and permits to detect the smallest prey that can be attacked by H. axyridis. 

The risk of detecting a contact instead of a predation event is very limited. In addition, 

exogenous alkaloids are persistent in the predators and can still be detected several days after 

an IGP event. Results from potato fields confirm that H. axyridis is an intraguild predator of 

native coccinelids (A. bipunctata, C. septempunctata, P. quatuordecimpunctata) under natural 

conditions. 


We thank Evelyne Haulotte (Laboratory of Organic Chemistry, ULB) for the supply of synthetic 

adaline. Financial support was provided by the Belgian National Fund for Scientific Research, FNRS 

(grant FRFC 2.4615.06) 

References 

Daloze D., Braekman J. & Pasteels J.M., 1994. Ladybird defence alkaloids: Structural, 

chemotaxonomic and biosynthetic aspects (Col.: Coccinellidae). Chemoecology 5:173- 

183. doi: 10.1007/BF01240602. 

Gagnon A.E., Heimpel G.E. & Brodeur J., 2005. Detection of intraguild predation between 

coccinellids using molecular analyses of gut-contents. International symposium on 

biological control of aphids and coccids (Tsuruoka, Japan). 

Hagler J.R., 2006. Development of an immunological technique for identifying multiple 

predator-prey interactions in a complex arthropod assemblage. Ann App Biol 149: 153- 

165 

Hautier L., Grégoire J., de Schauwers J., Martin G., Callier P., Jansen J. & de Biseau J., 2008. 

Intraguild predation by Harmonia axyridis on coccinellids revealed by exogenous alkaloid 

sequestration. Chemoecology 18:191-196. doi: 10.1007/s00049-008-0405-4. 

Harwood J.D. & Obrycki J.J., 2005. Quantifying aphid predation rates of generalist predators 

in the field. Eur. J. Entomol. 102: 335-350 

Hoogendoorn M. & Heimpel G.E., 2001. PCR-based gut content analysis of insect predators: 

using ribosomal ITS-1 fragments from prey to estimate predation frequency. Mol Ecol 10: 

2059-2067 

Koch R.L., 2003. The multicolored Asian lady beetle, Harmonia axyridis: A review of its 

biology, uses in biological control, and non-target impacts. Journal of Insect Science 3:32. 

Pell J., Baverstock J., Roy H., Ware R. & Majerus M., 2008. Intraguild predation involving 

Harmonia axyridis : a review of current knowledge and future perspectives. BioControl 

53:147-168. doi: 10.1007/s10526-007-9125-x. 

Triltsch H., 1999. Food remains in the guts of Coccinella septempunctata (Coleoptera: 

Coccinellidae) adults and larvae. Eur. J. Entomol. 96: 355-364 

Ware R., & Majerus M., 2008. Intraguild predation of immature stages of British and 

Japanese coccinellids by the invasive ladybird Harmonia axyridis. BioControl 53:169- 

188. doi: 10.1007/s10526-007-9135-8.

Measuring the impact of Harmonia axyridis intraguild predation 

on native coccinellids in the field 

Louis HAUTIER 1 , Jean-Claude GRÉGOIRE², Pierre CALLIER³, Gilles SAN MARTIN 3,4 

Jean-Pierre JANSEN 1 & Jean-Christophe DE BISEAU 3 

1 

Département Lutte biologique et Ressources phytogénétiques, Centre wallon de Recherches 

agronomiques, Rue de Liroux, 2, B-5030 Gembloux, Belgium 

2 

Lutte biologique et Ecologie spatiale, Université Libre de Bruxelles, CP 160/12, Av. F.D. 

Roosevelt 50 – 1050, Brussels, Belgium 

3 

Eco-Ethologie évolutive, Université Libre de Bruxelles, CP 160/12, Av. F.D. Roosevelt 50 - 

1050 Brussels, Belgium 

4 

Present address: Écologie comportementale et conservation de la biodiversité , BDIV , 

Université catholique de Louvain, Croix du Sud 4 B-1348 Louvain-la-Neuve, Belgium 

Introduction 

The Multicoloured Asian Ladybird, Harmonia axyridis Pallas, was introduced in 1997 in 

Belgium for aphid biological control (Adriaens et al. 2003). In less than five years, this 

ladybird has invaded the whole of Belgium in urban, agricultural and semi-natural habitats 

and has an overlapping niche with that of several native species (Adriaens et al. 2008). In 

parallel, a decline of native ladybird species such as Adalia bipunctata (L.) and Adalia 

decempunctata (L.), was observed in tree habitats in Brussels (San Martin et al. in prep). The 

causes of this decline are not clearly identified and could be due to competition with or to 

intraguild predation by, H. axyridis. Intraguild predation (IGP) is defined as “killing and 

eating of species that use similar resources” (Polis et al. 1989); this interaction is very 

frequent in aphidophagous guilds because aphids are limited resources (Lucas 2005). Several 

laboratory studies have reported that H. axyridis acts as an intraguild predator of 

Cecidomyiidae, Coccinellidae and Chrysopidae (Koch 2003, Pell et al. 2008, Ware & 

Majerus 2008), except for interactions with the ladybird Anatis ocelata (L.) (Ware & Majerus 

2008). However, laboratory conditions are extreme circumstances for interactions and are 

considered a worst case. Under field conditions, IGP can be influenced by many factors such 

as extraguild prey (e.g; aphids), possibility of intraguild prey escape, and importance of niche 

overlap. 

To follow IGP by H. axyridis on native coccinellids in field conditions, a new method 

for detecting IGP was developed based on the detection of exogenous alkaloids from native 

ladybirds in H. axyridis larvae, using gas chromatography - mass spectrometry (Hautier et al., 

2008). With this detection method, IGP by H. axyridis in lime trees (Tilia spp.) was studied in 

twenty sites in Brussels between June and July 2008. 40 to 110 branchs of lime tree, 

depending on site size, were beaten with a stick above a sweep net (65 cm in diameter and 

130 cm in depth). H. axyridis larvae were isolated in microtubes and were kept in a freezer at 

-20°C until alkaloid analysis. 

31

32 

Ladybirds sampled 

From the 20 sites sampled, 13 species of adult ladybirds were collected (Figure 1). The most 

abundant species was H. axyridis: it was present in 18 out of the 20 sites sampled and was 

also the most abundant with 791 specimens collected. Next, but 15 times less abundant and 

each present on half of the sites, were 4 natives species: 3 aphidophagous tree species, A. 

decempunctata (L.), Calvia quatuordecimguttata (L.) and Calvia decemguttata (L.), and 1 

mycetophagous species: Halyzia sedecimguttata (L.). Further, 2 generalist species and 2 tree 

species were caught: A. bipunctata (L.), Propylea quatuordecimpunctata (L.) and Exochomus 

quadripustulatus (L.), Oenopia conglobata (L.), respectively. Anatis ocelata (L.), Myrrha 

octodecimguttata (L.) and Aphidecta obliterata (L.) live in coniferous trees and are not 

associated with lime trees, which explains the low catch of these species. 

Anatis ocelata (1/20) 

Myrrha 18-guttata (1/20) 

Psyllobora 22-punctata (1/20) 

Oenopia conglobata (1/20) 

Aphidecta obliterata (2/20) 

Propylea 14-punctata (2/20) 

Exochomus 4-pustulatus (2/20) 

Adalia 2-punctata (7/20) 

Halyzia 16-guttata (11/20) 

Calvia 10-guttata (11/20) 


Adalia 10-punctata (15/20) 

Harmonia axyridis (18/20) 

1 

1 

1 

1 

2 

2 

2 

9 

39 

47 

52 

52 

791 

0 100 200 300 400 500 600 700 800 900 

No adults/ 1540 beatings 

Figure 2. Numbers of adult ladybirds collected. In brackets: number of sites in which they 

were caught. 

An even lower number of species was present as larvae (Figure 2). Again, H. axyridis was the 

dominant species, present in all sites with 737 larvae. To the contrary, only 33 native 

coccinellids larvae (C. quatuordecimguttata, Adalia spp., C. decemguttata and P. 

quatuordecimpunctata) were collected in 7, 5, 2 and 1 out of 20 sites, respectively. All these 

aphidophagous species are in competition for food and are potential intraguild prey for H. 

axyridis.

Unidentified (9/20) 

Propylea 14-punctata (1/20) 


Adalia spp (5/20) 


Harmonia axyridis (20/20) 

1 

3 

17 

9 

20 

737 

0 100 200 300 400 500 600 700 800 

No larvae/ 1540 

beatings 

Figure 2. Numbers of ladybird larvae collected. In brackets: number of sites in which they 

were caught. 

Alkaloid contents 

The analysis of 590 H. axyridis larvae collected on lime trees revealed exogenous alkaloids in 

21% larvae coming from all of sampled sites, except for one site (LEO) (Figure 3). Positive 

larvae contained mainly one alkaloid but in 6% of the positive larvae, two alkaloids were 

detected in each individual, resulting from double predation on two different coccinellid 

genera. 

Three exogenous alkaloids were identified in the H. axyridis larvae analysed: adaline, 

propyleine and calvine. They are naturally present in Adalia spp., in P. 

quatuordecimpunctata, and in Calvia spp. (Laurent et al. 2005). The detection of these 

exogenous alkaloids in H. axyridis larvae confirms the existence of intraguild predation on 

these native species in the field. 

No H.axyridis larvae analysed 

90 

80 

70 

60 

50 

40 

30 

20 

10 

0 

LEO 

BAR 

0 Exogenous alkaloid (-) 

1 Exogenous alkaloid (+) 

2 Exogenous alkaloids (++) 

CHA 

VUB 

BAS 

SOL 

PAP 

MEL 

PLA 

ROS 

TEN 

CLE 

FOR 

Figure 3. Numbers of H. axyridis larvae with (black and grey) and without (white) exogenous 

alkaloids in each site. 

FLA 

DEN 

DIE 

PRA 

TRI 

WOL 

CAM 

33

34 

Conclusion 

In conclusion, we can state that the invasive ladybird H. axyridis is becoming the dominant 

coccinellid species on lime trees in Brussels, both in terms of presence in the sites and 

abundance. The analysis of exogenous alkaloid content of H. axyridis larvae reveals the 

existence of intraguild predation on native coccinellids in 21% of the individuals collected, in 

19 out of the 20 sites sampled. This intraguild predation concerns mainly Adalia spp. and, to a 

lesser extent, Calvia spp. and P. quatuordecimpunctata. These results indicate that intraguild 

predation is not a fortuitous event and support the hypothesis that H. axyridis intraguild 

predation on Adalia spp. could explain the observed decline of the latter species in trees. 

References 

Adriaens T., Branquart E. & Maes D., 2003. The multicoloured Asian ladybird Harmonia 

axyridis Pallas (Coleoptera: Coccinellidae), a threat for native aphid predators in 

Belgium? Belgian Journal of Zoology 133:195-196. 

Adriaens,T., San Martin y Gomez G. & Maes D., 2008. Invasion history, habitat preferences 

and phenology of the invasive ladybird Harmonia axyridis in Belgium. BioControl 53:69- 

88. doi: 10.1007/s10526-007-9137-6. 

Hautier L., Grégoire J., de Schauwers J., Martin G., Callier P., Jansen J. & de Biseau J., 2008. 

Intraguild predation by Harmonia axyridis on coccinellids revealed by exogenous alkaloid 

sequestration. Chemoecology 18:191-196. doi: 10.1007/s00049-008-0405-4. 

Koch R.L., 2003. The multicolored Asian lady beetle, Harmonia axyridis: a review of its 

biology, uses in biological control, and non-target impacts. Journal of Insect Science 3. 

Laurent P., Braekman J. C. & Daloze D., 2005. Insect chemical defense. Topics in Current 

Chemistry 240:167-230. 

Lucas E. 2005. Intraguild predation among aphidophagous predators. European Journal of 

Entomology 102:351-363. 

Pell J., Baverstock J., Roy H., Ware R. & Majerus M., 2008. Intraguild predation involving 

Harmonia axyridis: a review of current knowledge and future perspectives. BioControl 

53:147-168. doi: 10.1007/s10526-007-9125-x. 

Polis G.A., Myers C.A. & Holt R.D., 1989. The ecology and evolution of intraguild 

predation: potential competitors that eat each other. Annual Review of Ecology and 

Systematics 20:297-330. 

Ware R. & Majerus M., 2008. Intraguild predation of immature stages of British and Japanese 

coccinellids by the invasive ladybird Harmonia axyridis. BioControl 53:169-188. doi: 

10.1007/s10526-007-9135-8.

Comparing Fallopia japonica, F. sachalinensis and their hybrid F. 

xbohemica in Belgium: population ecology, functional traits and 

invasiveness 

Basile HERPIGNY 1, 2 , Grégory MAHY 3 & Pierre MEERTS 1 

1 Laboratoire d’Ecologie Végétale et Biogéochimie, Université Libre de Bruxelles, Belgium. 

2 Corresponding author ; e-mail : bherpign@ulb.ac.be 

3 Laboratoire d’Ecologie, Faculté Agronomique de Gembloux, Belgium 

Introduction 

In the course of the nineteenth century two Fallopia (knotweed) species were introduced to 

Europe from Asia. F. japonica (2n=88) has become one of the most proliferous invasive 

plants in Europe (Beerling et al, 1994). F. sachalinensis (2n=44) is much less invasive and 

still rare in Western Europe. Though F. japonica is clonal and male sterile in Europe, it can be 

pollinated by F. sachalinensis. This cross produces F. ×bohemica, a hybrid that is said to 

spread even more rapidly than its parents (Bailey & Wisskirchen, 2006; Gammon et al, 2007; 

Mandak et al, 2004). The biological reasons for this high invasiveness are unknown. The 

physiology of the hybrid has been less studied than the physiology of its parents (Adachi et al, 

1996; Marigo & Pautou, 1998; Price et al, 2001), but the cause could be genetic. Recent data 

(Tiébré et al, 2007) show that the hybrid is represented in Belgium (and other European 

countries: Mandak et al, 2003) by at least three cytotypes (2n= 44, 66, 88) and that it could be 

more variable than its parents. It is possible that the hybridisation in the japonicasachalinensis 

complex is a factor promoting quick evolution. Hybridisation and 

polyploidisation are well-known evolutionary mechanisms and their importance in the 

evolution of invasiveness has been shown in other polyploidy complexes (Lee, 2002; 

Ainouche et al, 2003). 

Our objective is to test if the three taxa have contrasting values of key functional traits 

that might explain their contrasting invasiveness. We also examine if the hybrid is 

intermediate between its parents or, alternatively, if it shows transgressive variation in some 

traits. 

Study outline and preliminary results 

In the course of 2008, we monitored, from April to August, the following traits in populations 

from six sites where two or three taxa coexist in sympatry: shoot height, number of leaves and 

ramifications (secondary and tertiary), number and length of internodes, leaf area. That the 

taxa coexist allows us to assume that phenotypic variation originates solely from genetic 

differences and is not environmentally induced. The first results reveal interesting differences 

among the three taxa. F. ×bohemica is intermediate between its parents for height (fig 1), 

number of internodes and ramifications, length of the internodes, and leaf area. It shows 

transgressive variation for number of leaves (Figure 2), outperforming both parental taxa. 

Other functional traits were measured once or twice during the year, on populations of 

the three taxa in a site called “Verrewinkel graveyard” (Uccle, Brussels). Specific leaf area 

(SLA) was measured during biomass peak (July). Leaf and shoot water, nitrogen and carbon 

content were measured during biomass peak and leaf senescence (October). Results show that 

35

36 

F. japonica has lower SLA (Figure 3) but higher N resorption efficiency (up to 70% N is 

resorbed from senescing leaves, vs. 40% in F. sachalinensis) than the two other taxa (Figure 

4). F. sachalinensis has higher N content than the two other taxa. F. ×bohemica is 

intermediate between its parents for most traits. 

This monitoring of functional traits will be carried on from April to July 2009. SLA, 

leaf water, nitrogen and carbon content will also be measured again. Field observations will 

be complemented by a “semi-controlled conditions” experiment to test if the three taxa show 

contrasting phenotypic plasticity of functional traits in response to different soil fertility 

conditions. For this, rhizomes were taken from seven knotweed populations (two F. japonica, 

two F. sachalinensis and three F. ×bohemica). They were planted in twenty litre pots, with 

different doses of NPK fertilizer. Their functional traits will be monitored from 2009 to the 

end of 2010. 

Height (cm) 

350,00 

300,00 

250,00 

200,00 

150,00 

100,00 

50,00 

0,00 

Growth 

0 14 28 42 56 70 84 98 112 126 140 

Time (days since the 17th of March) 

F jap 

F boh 

F sach 

Figure 1: Monitoring of the main axis height of the three Fallopia taxa. Bars indicate the standard 

error. 

Number of leaves on the 

main axis 

Variation of the number of leaves on the main axis 

30,00 

25,00 

20,00 

15,00 

10,00 

5,00 

0,00 

0 50 100 150 

Time (days since the 17th of March) 

F jap 

F boh 

F sach 

Figure 2: Monitoring of the number of leaves on the main axis of the three Fallopia taxa. Bars 

indicate the standard error.

SLA (m²/kg) 

30 

25 

20 

15 

10 

5 

0 

SLA (specific leaf area) 

F. japonica F. xbohemica F. sachalinensis 

Taxon 

Figure 3: Specific leaf area (m²/kg) of the three Fallopia taxa. Bars indicate the standard deviation. 

Nitrogen content (%) 

3 

2,5 

2 

1,5 

1 

0,5 

0 

F. jap 

F. xboh 

F. sach 

Leaves nitrogen content 

Summer Fall 

Season 

Figure 4: leaves nitrogen content (%) of the three Fallopia taxa, measured at biomass peak 

(summer) and during leaf senescence (fall). Bars indicate the standard deviation. 

37

38 

References 

Adachi N., Terashima I. & Takahashi M., 1996. Nitrogen translocation via rhizome systems 

in monoclonal stands of Reynoutria japonica in an oligotrophic desert on Mt Fuji: Field 

experiments. Ecological research 11: 175-186. 

Ainouche M.L., Baumel A., Salmon A. & Yannic G., 2003. Hybridization, polyploidy and 

speciation in Spartina (Poaceae) New Phytologist 161: 165-172. 

Bailey J., & Wisskirchen R., 2006. The distribution and origins of Fallopia x bohemica 

(Polygonaceae) in Europe. Nordic Journal of Botany 24: 173-200. 

Beerling D.J., Bailey J.P. & Conolly A.P., 1994. Fallopia japonica (Houtt.) Ronse Decraene. 

J Ecol 82: 959–979. 

Gammon M.A., Grimsby J.L., Tsirelson D. & Kesseli, R., 2007. Molecular and 

morphological evidence reveals introgression in swarms of the invasive taxa Fallopia 

japonica, F. sachalinensis, and F. xbohemica (Polygonaceae) in the United States. 

American Journal of Botany 94: 948-956. 

Lee C.E., 2002. Evolutionary genetics of invasive species. Trends in Ecology and Evolution 

17: 386-391. 

Mandák B., Pyšek P., Lysak M., Suda J., Krahulcova A. & Bimova K., 2003. Variation in 

DNA-ploidy levels of Reynoutria taxa in the Czech Republic. Annals of Botany 92: 265- 

272. 

Mandák B., Pyšek P. & Bímová K., 2004. History of the invasion and distribution of 

Reynoutria taxa in the Czech Republic: a hybrid spreading faster than its parents. Preslia 

76:15-64. 

Marigo G. & Pautou G., 1998. Phenology, growth and ecophysiological characteristics of 

Fallopia sachalinensis. Journal of Vegetation Science 9: 379-386. 

Price E.A.C., Gamble R., Williams C.G. & Marshall, C., 2001. Seasonal patterns of 

partitioning and remobilization of C-14 in the invasive rhizomatous perennial Japanese 

knotweed (Fallopia japonica (Houtt.) Ronse Decraene). Evolutionary Ecology 15: 347- 

362. 

Tiébré M.S., Bizoux J.-P., Hardy O.J., Bailey J.P. & Mahy, G., 2007. Hybridization and 

morphogenetic variation in the invasive alien Fallopia (Polygonaceae) complex in 

Belgium. American Journal of botany 94: 1900-1910.

Mediterranean container plants and their stowaways: A potential 

source of invasive plant species 

Ivan HOSTE* 1 & Filip VERLOOVE 2 

1, 2 National Botanic Garden of Belgium, Domein van Bouchout, B-1860 Meise. 

*Corresponding author: ivan.hoste@br.fgov.be 

Introduction: An expanding catalogue of neophytes 

A recently published catalogue of neophytes (Verloove 2006) lists 1969 non-native vascular 

plants recorded from Belgium between 1800 and 2005. By the end of 2008, only a few years 

later, 144 new taxa (+ 7.3 %) had been added to the catalogue, not including some 20 new 

taxa resulting from the current study. The composition of the group of 144 new additions 

differs markedly from the one of the catalogue 1800-2005. Species originally introduced as 

ornamentals make up 60 % of the additions, against 33 % in 1800-2005. Together the two 

most important families in the catalogue (Poaceae and Asteraceae) make up only 5 % of 

additions, as compared to 29 % in the catalogue. 

The clear differences between the catalogue and the additional recent dataset are linked 

with real changes and inevitable bias. The area of origin of diaspores of alien plant species 

and the routes and vectors involved have changed as a result of historical trends and events, 

which often reflect worldwide economic change. Certain categories have disappeared, 

whereas others are new or at least more important today than before. For instance, 14 % of the 

species in the 1800-2005 catalogue have only been recorded as wool aliens from the Vesdre 

valley. Yet, the historical bias in the cumulative dataset explains only part of the differences. 

The list of post-2005 additions also reflects an important bias. The increased number of 

introduced ornamentals in the list, recorded as garden escapes or locally naturalizing species, 

undoubtedly reflects more intensive fieldwork in urban areas in the past few years. The partial 

shift from grain terminal aliens to urban aliens has several different causes, including the 

attraction of formerly unexplored fields for the fieldworker, a possible real reduction in the 

number of imported grain aliens in Belgian port areas, a strongly diminished potential for 

finding novelties in the well-studied group of grain aliens in Belgium, and a long-standing 

neglect of certain groups of escapes from cultivation (e.g. shrubs and trees). 

Cardamine corymbosa, from New Zealand, is a rapidly spreading plant species. In 

2008, while doing fieldwork on this weed in nurseries and garden centres, we chanced upon a 

seemingly important and largely overlooked category of introduced aliens. We therefore 

decided to study this alien weed flora more in detail. 

Mediterranean container aliens in Belgium 

Propagule pressure has recently been described as “the new frontier in invasion ecology” 

(Richardson & Pyšek, 2008). However, our knowledge about the precise pathways followed 

by incoming aliens is often very incomplete. The information we gathered during a single 

year of prospection in garden centres in Belgium amply illustrates this. 

Between late spring and autumn a large number of garden centres were visited, with the 

aim of preparing a list of alien weeds growing in containers with Mediterranean plants, 

especially palms, olives and figs. Occasionally we also recorded plants that had obviously 

39

40 

escaped from such containers and that thrived on the ground in the direct vicinity of the 

containers. 

In Belgium, the highly increased popularity of Mediterranean container plants is a very 

recent, early 21 st century phenomenon. This popularity can be seen as the end result of a 

cascade of events and trends, linking the increased level of prosperity of the 1960s with 

tourism around the Mediterranean, a heightened esteem for gardening, and finally the desire 

to evoke in the home garden a tinge of the Mediterranean flavour and memories from summer 

holidays in the South. 

In Western Europe, most Mediterranean container plants are imported from Spain or 

Italy. Together with the ornamentals, large numbers of weeds (seeds as well as young plants), 

and frequently also other organisms such as snails, are unintentionally introduced in garden 

centres, situated all over the country. Once sold, the containers and their stowaway weeds find 

their way into hundreds and thousands of private gardens, parks, etc. While a lot of these 

weed species have also been recorded as grain aliens in port areas, this recently discovered 

pathway offers excellent opportunities for widespread dispersal. Furthermore, seeds can 

germinate in a microhabitat that is literally the same as the one in which the mother plant once 

grew. 

An overview of the results of our prospections is given in table 1. Of 122 identified 

species, 27 are indigenous to Belgium, and these are also indigenous to Spain and/or Italy. 

The remaining 95 species are naturalized in Belgium (28 species), casuals (44), or are 

recorded for the first time (23). A remarkably high number among these 95 species (33 = 

35 %) entered Belgium from a secondary distribution range in Spain and/or Italy, not from 

their natural range. (It should be kept in mind, though, that in the past some of these, e.g. 

Coronopus didymus, might have entered Belgium directly from their natural range too.) See 

for more details on the species list Hoste et al. (2009). 

Table 1. An overview of records of Mediterranean container aliens from garden centres in 

Belgium in 2008. 

Status in Belgium (*) 

Number of container aliens (records 2008) 

Indigenous to 

Spain and/or 

Italy 

Naturalized, casual or 

not yet recorded from 

Spain and/or Italy 

Indigenous s.l. 27 0 27 

Not indigenous, but 

rather widespread and/or 

more or less naturalized 

16 12 28 

Casual 30 14 44 

Not previously recorded 16 7 23 

Total 

Total 89 33 122 (**) 

(*) Based on Lambinon et al. (2004) and Verloove (2006). 

(**) Not including a number of crypto-aliens (that is, species that are common in both 

Belgium and at least part of the western Mediterranean), probable ornamental escapes, 

and taxa that could only be identified to genus level.

All species from our survey have in common that they followed more or less the same 

trajectory and used the same vector to travel from the Mediterranean to Western Europe. As 

such, they illustrate a new episode in the worldwide exchange of biota that has been going on 

for centuries. Apart from that, many species often previously had a rather different species 

history, as suggested by a few individual examples. 

In the mid-19 th century Bowlesia incana, indigenous to America, persisted for some 

years at a single location in southern France. It was first recorded from Spain in 1986, where 

it is now well established but rare. From Spain it reached Belgium in 2008 (figure 1). In 

contrast with other Bowlesia, this weedy species has little apparent morphological adaptation 

for seed dissemination. It is therefore remarkable that this plant, which has fruits devoid of 

glochids, has become a much more widespread alien in several parts of the world than all 

seven glochidiate-fruited Bowlesia combined (Mathias & Constance 1965). 

Today three American Chamaesyce frequently enter Belgium from Spain, but their pre- 

2008 history is different. Chamaesyce maculata (first recorded

42 

The new link between the Mediterranean and Western Europe has brought us some new 

species, along with a much larger group of species that previously followed other trajectories 

and used alternative vectors to enter Belgium. The active importation of a relatively 

homogeneous group of Mediterranean plants creates a propagule pressure bias toward species 

of warmer climates, and we should therefore not interpret this upsurge of Mediterranean 

aliens as a clear-cut illustration of global warming. Several species in the list apparently have 

not yet been mentioned in the Spanish or Italian botanical literature, indicating that the study 

of propagule pressure is indeed still hampered by insufficient data on the precise area of 

origin – primary or secondary – of introduced aliens. 

Given the large number of arrivals (both range of species and amount of diaspores), we 

can expect that at least some species will naturalize in Belgium as a result of the importation 

of Mediterranean container plants. The tens rule, a useful rule of thumb, states that 10 % of 

imported species become casuals, and 10 % of those casuals become naturalized (Williamson 

1996). There are indications that in the early 21 st century some candidates for naturalization 

are increasing in urban areas, including several species from our container aliens list: 

Piptatherum miliaceum, Polycarpon tetraphyllum, Sisymbrium irio, S. orientale. Species from 

our list have recently been recorded from similar habitats in England, e.g. Urtica 

membranacea (Boucher & Partridge 2006, anon. 2008) and Galium murale (Nicolle 2008), 

and in France (e.g. Chamaesyce prostrata; Bedouet 2008). 

Where do we go from here? 

The present study on Mediterranean container aliens clearly shows that there is still an urgent 

need for new data on propagule pressure. Again and again new data remind us of the 

complexity of the naturalization and invasion process. In our urge to find the laws and 

mechanisms that drive these processes, we should never forget that invasions are context 

specific (Richardson & Pyšek 2008). 

The early stages of invasion include long-distance transport and successful introduction 

of plants and animals outside their natural area of distribution. Especially in these early stages 

human activities largely determine what happens, where, when, and how. Among invasion 

ecologists these activities are often perceived as annoying interference with or disturbance of 

ecological processes. A more fitting approach accepts humans for what they really are: A 

primary agent in the bewilderingly rich and complicated succession of events that constitutes 

the essence of history. 

Both deficiencies in the available data and the fads and fancies of human history often 

make it difficult to interpret cumulative datasets on individual species. Before analysing 

results we should always carefully check whether the data is uniform. A single cumulative 

curve, based on historical data that span decades or more than a century, is often based on two 

or more subsets of records. This may result from changing global trade routes, from 

dwindling or increased trade volumes, etc. 

An example is the small dataset for Setaria adhaerens in Belgium: in the first half of the 

20 th century it was exclusively recorded as a rare wool alien, whereas in 2008 it was 

frequently recorded from garden centres. A quite different example is the invasion of Senecio 

inaequidens in Western and Central Europe. In a recent study, Bossdorf et al. (2008) argue 

that S. inaequidens could only start to spread after new frost-resistant and competitive 

genotypes had been introduced from mountainous regions in southern Africa, decades after 

the species had first been introduced in Central Europe. Such studies on the introduction and 

invasive history of an individual plant species, based on molecular research, illustrate the

dangers of uncritically using the term ‘lag time’. Close scrutiny of data that span long periods 

of time can prevent us from turning lag time into a black box. 

Barabási (2002) observed that we live in a small world. Our world is small because 

society is a very dense web. It is indeed humans who have created this small world in which 

plants and animals continue to disperse and propagate, basically following the same 

ecological rules as before. The science of networks described by Barabási offers opportunities 

for a new framework for the study of biological invasions. On all levels, from the local to the 

intercontinental, it is important to better understand how propagules are actively or passively 

being dispersed. How and in what numbers are seeds dispersed within and between plant 

nurseries, garden centres, private gardens, and their immediate vicinity? And how have 

pathways for the introduction of aliens changed over time? 

What explains the arrival of over 100 Bowlesia incana seedlings in a Belgian garden 

centre in the spring of 2009, thousands of kilometres away from its natural area of origin in 

America? Chances are small indeed for this to happen if dispersal were only possible through 

a distributed network (figure 2c). But that is not what networks look like in the real world. 

Real networks don’t have a homogeneous, mesh-like architecture. Ideas and goods, including 

ornamentals and weeds, spread across heterogeneous networks with numerous tiny nodes and 

a few large hubs characterized by an extraordinary number of links (figure 2b). For Bowlesia 

to reach Belgium, chances rise dramatically once it can travel through large decentralized 

networks. 

In theory, Bowlesia in its native range could be just three steps away from your home 

garden (figure 1). First it travels incognito from America to southeast Spain, where it settles 

as a persistent weed in a nursery that operates as a major node in an international horticultural 

trade network. The second step takes Bowlesia, still incognito, from Spain into a Belgian 

garden centre. There Bowlesia and its companion Trachycarpus palm abide their time, until 

one day they are sold and together transported into your private garden. 

“Networks are only the skeleton of complexity, the highways for the various processes 

that make our world hum.” (Barabási, 2002) Our world. And the container aliens’ world as 

well. 

Figure 2. Centralized, decentralized and distributed networks. (Source: A.-L. 

Barabási, Linked) 

43

44 


Our species list would have been much shorter without the input by C. Nagels and L. Andriessen, who 

visited numerous garden centres in the eastern part of Flanders. 

References 

Anonymous, 2008. Climate change could spread plants. Horticulture Week 17 October 2008: 

5. 

Barabási A.-L., 2002. Linked. The new science of networks. Cambridge (MA), Perseus 

Publishing. 

Bedouet F., 2008. Euphorbia prostrata Aiton (Euphorbe prostrée). Le jouet du vent 20: 3. 

Bossdorf O., Lipowsky A. & Prati D., 2008. Selection of preadapted populations allowed 

Senecio inaequidens to invade Central Europe. Diversity and Distributions 14: 676-685. 

Boucher A. & Partridge J., 2006. Urtica membranacea, an annual nettle, in Warwick: a first 

British record? BSBI News 103: 29-30. 

Hoste I., Verloove F., Nagels C., Andriessen L. & Lambinon J., 2009. De adventievenflora 

van in België ingevoerde mediterrane containerplanten. Dumortiera 97: 1-16. 

Lambinon J., Delvosalle L. & Duvigneaud J., 2004. Nouvelle Flore de la Belgique, du Grand- 

Duché de Luxembourg, du Nord de la France et des Régions voisines. 5 ème édition. Meise, 

Jardin botanique national de Belgique. 

Mathias M.E. & Constance L., 1965. A revision of the genus Bowlesia Ruiz & Pav. 

(Umbelliferae–Hydrocotyloideae) and its relatives. Univ. of California Publications in 

Botany 38. Berkeley & Los Angeles, Univ. of California Press. 

Nicolle D. J., 2008. Galium murale – a foothold in Eastbourne? BSBI News 109: 57-58. 

Richardson D. M. & Pyšek P., 2008. Fifty years of invasion ecology – the legacy of Charles 

Elton. Diversity and Distribution 14: 161-168. 

Verloove F., 2006. Catalogue of neophytes in Belgium (1800-2005). Scripta Botanica Belgica 

39. Meise, National Botanic Garden of Belgium. 

Williamson M., 1996. Biological invasions. London etc., Chapman & Hall.

Water frogs in Wallonia: genetic identification of the introduced 

taxa (Pelophylax ssp.) and impact on indigenous water frogs 

(Pelophylax lessonae and P. kl. esculentus). 

Christiane PERCSY 1,* & Nicolas PERCSY 2 

1 * 

ULB & UCL, private address: chemin du Bon Air 12, B-1380 Ohain. Corresponding author: 

cpercsy@gmail.com 

2 Institut Supérieur d'Architecture, UMONS 

Introduction 

Two taxa of water frogs are native in Wallonia: Pelophylax lessonae and P. kl. esculentus. 

But water frogs, from different origins, have been introduced in Wallonia during the last 25 

years, mainly as a consequence of aquatic horticulture: aquatic plants are imported from 

central Europe, in containers, with eggs, tadpoles or adults of water frogs; these frogs 

reproduce successfully in the horticulture ponds and are sold (or given) to people creating an 

ornamental pond in their garden. So, the frogs are introduced in many places in the country; 

then they spread in the neighbourhood and colonize (semi-)natural habitats. As a 

consequence, P. ridibundus (Figure 1) has become the most frequent green frog in Brabant 

wallon (Percsy & Percsy, 2002a and 2002b). It is also abundant in the neighbourhood of large 

cities (Brussels, Liège, Namur, Verviers) (Percsy & Percsy, 2007). 

Alien water frogs – i.a. P. ridibundus – occupy sites where P. kl. esculentus and P. 

lessonae are present, as well as other amphibians (mainly R. temporaria, Bufo bufo, Triturus 

sp.); it is a predator of these species. Since P. ridibundus is bigger than the two native green 

frogs, they outcompete the latter for territory, feeding and breeding Furthermore, because of 

the very particular genetic relationship between the three green frogs (hybridogenesis : see, 

e.g., Berger 1988), the introduction of P. ridibundus in P. kl. esculentus and P. lessonae 

populations may lead to genetic pollution of the latter species. Finally, foreign frogs may 

carry diseases, which is a threat for indigenous amphibians (Kok 2001). 

Figure 1. Lake Frog (Pelophylax ridibundus), 

Lasne, Wallonia. © C. & N. Percy 

45

46 

P. ridibundus has been introduced in other countries of western Europe and is known 

to have outcompeted other amphibians sharing the same habitat (see, e.g., Günther in Gasc et 

al. 1997, Grossenbacher 1988). Consequently, it is important to evaluate the impact of the 

introduced water frogs in Wallonia. Unfortunately, recognizing the different taxa of water 

frogs present in Wallonia is not easy. To insure the identifications we have made, we 

collected, in 2002, samples from 47 frogs from 8 different populations and submitted these for 

enzymatic and genetic analyzis. The « Laboratoire d'Ecologie des Hydrosystème fluviaux » 

(Prof. Joly) at the University of Lyon performed protein electrophoresis and the « Museum 

für Naturkunde » (Prof. Plötner) in Berlin investigated mitochondrial DNA. The results of 

these analyses allow to obtain: 

• a validation of the identification method of the taxa on the field and, thus, a reliable 

assessment of the evolution of the populations; 

• the determination of the geographic origin of the introduced frogs; 

• evidence of hybridization and/or introgression between Pelophylax ridibundus and the 

indigenous frogs P. lessonae or P. kl. esculentus; similar introgressions between 

taxons of water frogs have been observed in other European countries (e.g., Vorburger 

et al. 2003, Plötner et al. 2008, Holsbeek et al. 2008). 

Validation of the identification method 

We wanted to test a method of field identification that can be performed without capturing the 

frogs, hence relying on morphological characters (possibly by observing using field-glasses) 

and mating calls. We select, for Wallonia, the following easy criteria. 

Pelophylax lessonae P. kl. esculentus P. ridibundus 

� Length of the hind 

leg vs body length 

short intermediate long 

� Tibia length vs 

femur length 

shorter or equal equal longer 

� Colour of the back often with a vivid yellow with or without a vivid no vivid yellow tint 

of the thigh and/or 

of the groin 

tint 

yellow tint 

� Skin coarseness of 

the back 

weak weak generally strong 

� Vocal bags white, sometimes tinged 

with pink 

greyish or medium-grey dark grey 

� Call continuous, long and 

uniform 

shorter, modulated jerky 

Note that the evaluation of these criteria is somewhat subjective and may depend on the 

observer's experience. This is why identification is validated only if most of the criteria are 

coherently satisfied. By comparing our field identifications with the enzymatic- and DNAanalysis, 

we obtain the following results concerning the reliability of the above identification 

method of the taxa in the field (see Percsy & Percsy 2009):

• the distinction between the native green frogs, on the one hand, and the introduced 

water frogs, on the other hand, is valid in 100% of the cases; 

• the distinction between P. kl. esculentus and P. lessonae is less valid : two of the 

seven presumed « lessonae » are actually « esculentus »; 

• determination of the geographical origin of the alien water frogs is not reliable; 

• the distinction between native and non native male green frogs may be done using 

only the mating call. (This result has been confirmed by further bioacoustic analyses, 

using oscillograms and sonograms of the calls). 

Moreover, analyses of various photographs of the captured frogs allows to evaluate different 

other classical criteria for the identification of the water frogs (Percsy & Percsy 2009). 

Geographic origin of the alien water frogs 

All alien water frogs are determined as P. ridibundus or P. cf. ridibundus (from Anatolia). 

The mitochondrial analysis of our samples shows that introduced water frogs in Wallonia 

have at least two different geographic origins, corresponding to three different haplotypes: 

• haplotype C, present in Central Europe; 

• haplotype E1 and 

• haplotype E2, both typical for Anatolian, northern Greek or Bulgarian water frogs). 

Three of the eight populations we studied contain both indigenous and exotic water frogs (the 

latter having the same haplotype); two other populations contain only alien water frogs, all 

with the same haplotype; both haplotypes C and E1 are present on a same site; finally, two 

sites do not have exotic water frogs. 

Introgression 

Two samples corresponding to frogs that we have identified in the field as P. kl. esculentus 

and whose enzymec analyses concurs with P. kl. esculentus, have mitochondrial DNA of type 

E2. This shows that hybridization between native frogs and P. ridibundus has occurred. 

Concerning the « lessonae » haplotype of our samples, note that this type is different from all 

other « lessonae » haplotypes known (T. Ohst, pers. comm.). 

Conclusion and management strategy 

Alien water frogs have spread in certain regions of Wallonia: they come from at least two 

different origins (central Europe or Balkan and Anatolia). Such frogs have a negative impact 

on our native water frogs and, probably, on other amphibian species. 

Our work validates a method to separate, on the field, the native water frogs ( P. lessonae and 

P. kl. esculentus) from the introduced frogs P. ridibundus and P. cf. ridibundus. It also shows 

the existence of hybridation between alien and native frogs in Wallonia. 

Alien water frogs are already abundant in certain regions of Wallonia and it is impossible to 

eradicate them there. On the contrary, exotic frogs seem absent from a large part of the 

country, in particular in oligotrophic ecosystems where P. lessonae is dominant. Such 

47

48 

populations of water frogs have to be protected from invasion, because they are less frequent 

(Günther in Gasc et al. 1997); furthermore, the results above show that the « lessonae » 

haplotype of Wallonia is original, . Since, probably, P. ridibundus will not spontaneously 

colonize such ecosystems (Pagano et al. 2001), it is urgent to avoid introductions in these 

areas. Consequentely, the following measures should be implemented: 

• control of trade of water frogs in Belgium and in Europe (laws about trade should be 

enacted); 

• control of introduced frogs in target regions; 

• public awareness, to avoid the transfer of frogs from one place to another. 


We thank Prof. P. Joly, O. Grolet, Prof. J. Plötner and T. Ohst for the enzyme and genetic analyses.

References 

Berger, L., 1988. On the origin of genetic systems in European water frogs hybrids. Zoologica 

Poloniae 36 (1-4): 5-32. 

Gasc, J.P. et al. (eds), 1997. Atlas of Amphibians and Reptiles in Europe. Societas Europaea 

Herpetologica & Muséum national d’Histoire naturelle, Paris, 186 pages. 

Grossenbacher, K., 1988. Atlas de distribution des Amphibiens de Suisse. Documenta 

Faunistica Helvetiae n°8, Schweizerischer Bund für Naturschutz, Ligue Suisse pour la 

Protection de la Nature et Centre Suisse de Cartographie de la Faune, Bâle, 208 pages. 

Holsbeek, G., Mergeay, J., Hotz, H., Plötner, J., Volckaert, F. A. M. & De Meester, L., 2008. 

A cryptic invasion within an invasion and widespread introgression in the European water 

frog complex: consequences of uncontrolled commercial trade and weak international 

legislation. Molecular Ecology 17: 5023-5035. 

Kok, P. , 2001. Note sur l'introduction de Rana bedriagae Camerano, 1882 (Anura: Ranidae) 

en Belgique et ses possibles implications sur la batrachofaune indigène. Les Naturalistes 

belges 82 (1) : 25-30. 

Pagano, A., Joly, P., Plénet, S., Lehman, A. & Grolet, O. , 2001. Breeding habitat 

partitioning in the Rana esculenta complex: the intermediate niche hypothesis supported. 

Ecoscience 8 (3) : 294-300. 

Percsy, C. & Percsy, N., 2002a. Dix ans de suivi des populations indigènes et introduites de 

grenouilles « vertes » (Rana (Pelophylax) ssp., Anura, Ranidae) dans le bassin de la Lasne 

(Brabant wallon, Belgique). Bulletin de la Société Herpétologique de France n° 103: 59 – 

72. 

Percsy, C. & Percsy, N., 2002b. Evolution des populations indigènes et introduites de 

grenouilles « vertes » en Brabant wallon. Pages 213 – 218 in Peeters M. & J.L. Van 

Goethem (Eds), 2002. Belgian fauna and alien species. Actes du symposium: Faune belge: 

statut et tendances observées avec une attention particulière pour les espèces exotiques. 

Bulletin de l’I.R.S.N.B., Biologie 72, supplément, 297 pages. 

Percsy, C. & Percsy, N., 2007. Quatre chapitres sur les grenouilles vertes in Jacob J.P. et al., 

Amphibiens et reptiles de Wallonie. Aves-Raînne et CRNFB, série « Faune-Flore- 

Habitats » n°2. Namur, 384 pages. 

Percsy, C. & Percsy, N., 2009. Identification des grenouilles « vertes » (Pelophylax) en 

Wallonie : résultats de la confrontation de critères morphologiques et acoustiques avec des 

analyses enzymatiques et d'ADN (in press). 

Plötner, J., Uzzell, T., Beerli, P., Spolsky, C., Ohst, T., Litvinchuk, S. N., Guex, G. D., Reyer, 

H. U. & Hotz, H., 2008. Widespread unidirectional transfer of mitochondrial DNA: a case 

in werstern water frogs. J. Evol. Biol. 21: 668-681. 

Vorburger, C. & Reyer, H. U., 2003. A genetic mechanism of species replacement in 

European water frogs? Conservation Genetics 4: 141-155. 

49

Trends in the distribution of the Chinese mitten crab in the 

Scheldt estuary 

Maarten STEVENS 1 , Tom VAN DEN NEUCKER, David BUYSSE & Johan COECK 

Research Institute for Nature and Forest, Kliniekstraat 25, 1070 Brussels, Belgium. 

1 Corresponding author; e-mail: Maarten.Stevens@inbo.be 

Introduction 

The Chinese mitten crab (Eriocheir sinensis) is a catadromous species that spends most of its 

life in freshwater. E. sinensis is native to rivers and estuaries of central Asia, from North 

Korea to China. Since its arrival in Germany in the beginning of the 20 th century, the Chinese 

mitten crab has rapidly invaded coastal and inland waters throughout Europe. The species was 

first observed in Belgium in 1933 in the Sea Scheldt near Antwerp and is found nowadays in 

the main rivers and canals of the Scheldt basin. 

Results and discussion 

Macrocrustaceans were caught as bycatch during fish surveys using fyke nets in the Sea 

Scheldt in 1995, 1997 and 2008. During the surveys in the nineties, only a few mitten crabs 

were found. Ten years later, however, E. sinensis had spread throughout the estuary and more 

than 50 crabs can now be caught per fyke net per day. The expansion of the distribution and 

the increase of the population size of the Chinese mitten crab coincided with the improvement 

of the water quality in the estuary during the last decade. It is suggested that pollution may 

decrease mitten crab densities, by reducing the abundance of prey (Gollasch 1999). 

The abundance of mitten crabs in the estuary shows two distinct seasonal peeks 

(Figure 1). The first peek in spring coincides with the upstream migration of juveniles 

towards the freshwater reaches of rivers, where they burrow in the banks (Figure 2). The 

second peek in autumn coincides with the seaward spawning migration of adults. The highest 

densities are observed in the oligohaline and freshwater zone of the estuary. Their distribution 

is probably related to habitat availability and their ability to burrow (soft sediment banks). 

High population densities of this crab may have a significant adverse impact on the 

natural balance of the Scheldt ecosystem. Because of the crab’s flexible, omnivorous feeding 

habits, it may have a competitive edge over other bottom dwelling species. Their burrowing 

nature may also accelerate bank erosion and instability (Rudnick et al., 2005). 

References 

Gollasch S., 1999. Current Status on the increasing abundance of the Chinese mitten crab 

Eriocheir sinensis H. Milne Edwards, 1854 in German rivers. Proceedings of the First 

Annual Meeting on the Chinese mitten crab in California. 23 March. Sacramento, CA 

Rudnick D.A., Chan V. & Resh V.H., 2005. Morphology and impacts of the burrows of the 

Chinese mitten crab, Eriocheir sinensis H. Milne Edwards (Decapoda, Grapsoidea), in 

South San Francisco Bay, California, USA. Crustaceana 78(7): 787-807. 

51

52 

# crabs fyke -1 day -1 

150 

100 

50 

Liefkenshoek tunnel 

Rupelmonde 

Antwerp 

0 

0 

F M A M J J A S O N D J F M A M J J A S O N D J F M 

2007 2008 

Figure 1. Seasonal abundance of the Chinese mitten crab in the Scheldt estuary. 

Figure 2. Burrows of the Chinese mitten crab in the left side bank of a freshwater marsh creek in the 

Scheldt estuary. Inset: mitten crab in the opening of its burrow. 

350 

300 

250 

200 

150 

100 

50 

# crabs fyke -1 day -1 (Antwerp)

The invasion of ring-necked parakeet (Psittacula krameri) in 

Europe and Belgium: mechanisms and consequences for native 

biota. 

Diederik STRUBBE & Erik MATTHYSEN 

Evolutionary Ecology Group, Department of Biology, University of Antwerp, 

Groenenborgerlaan 171, 2020 Antwerp 

Establishment of ring-necked parakeets in Europe 

In Europe, at least 75 nonnative bird species established feral populations (Chiron et al. 2009) 

and, due in large part to their popularity as cage birds, parrots (Psittacidae) are well 

represented as invaders, accounting for about 18 % of Europe’s established exotic avifauna 

(DAISIE 2009). Among parrots, the ring-necked parakeet is undoubtedly the most successful 

invader. Although this parakeet originates from mostly subtropical regions in Africa and 

Asia, it has formed at least 65 populations in Europe (Lever 2005, Strubbe & Matthysen 

2009a), with population sizes varying from several tens to several thousands of birds (Strubbe 

& Matthysen 2007). In order to identify the mechanisms that allow this parakeet to invade 

Europe, we gathered data on parakeet releases and correlated the outcome of an introduction 

event with human population density and climatic factors, thereby testing two of the major 

hypotheses on the establishment success of non-native species, i.e. the ‘human activity’ 

(Taylor & Irwin 2004) and the ‘climate matching’ hypothesis (Williamson 1996). The former 

hypothesis states that human activity facilitates the establishment of alien species while the 

latter postulates that species have a higher probability to establish if they are introduced to 

regions with a climate similar to that in their native area. We found that parakeet 

establishment correlated positively with measures of human activity such as human 

population density, but negatively with the number of frost days, providing support for both 

hypotheses (Strubbe & Matthysen 2009a). Human activity is a root cause of species 

introductions (Westphal et al. 2008), but human-dominated habitats are often characterized by 

abundant food and we argue that this increased food supply could well explain the link 

between the parakeets’ establishment success and human population density, as food 

availability is one of the most important factors limiting bird populations and supplementary 

feeding may enhance breeding success (Robb et al. 2008), thus increasing population 

persistence probability. The negative relationship between parakeet establishment and the 

number of frost days indicates that parakeets may suffer from climate mismatch, although 

population crashes during harsh winter are rare, and establishment failure could also be 

caused by a reduced breeding performance, e.g. due to a lower body condition. Low 

temperatures can also impact on avian embryonic development, and further support for the 

climate matching hypothesis comes from Shwartz et al. (2009), who found that European 

parakeet populations have a much higher rate of egg infertility than parakeets in the native 

range (India), or populations introduced to warmer regions such as Israel. Predation pressure, 

however, is much lower in the introduced regions compared to the native range, and this 

‘predator release’ (Liu & Stiling 2006) partly offsets the climate mismatch, allowing the 

parakeet to flourish throughout much of temperate and Mediterranean Europe (Shwartz et al. 

2009). 

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54 

Ring-necked parakeets in Belgium 

Parakeet habitat selection and competition with native species 

Belgium, which is one of Europe’s exotic bird hotspots (Chiron et al. 2009), harbors one of 

the continent’s largest ring-necked parakeet populations. Originating from a deliberate 

release of about 40 parakeets in Brussels in 1974, the population numbered 8 000 to 8 500 

birds in 2006 (Weiserbs & Jacob 2007). The population doubles approximately every 4 year, 

and this growing population raises concerns for the loss of biodiversity of the native avifauna. 

Ring-necked parakeets are secondary cavity-nesters and as secondary cavity nesting 

communities depend on existing cavities and are hierarchically structured based on the 

production of and the competition for suitable nesting cavities (Martin & Eadie 1999), 

parakeets could pose a threat to native hole-nesting species. Based on preferred nesting cavity 

characteristics, parakeets might come into conflict with native nuthatches (Sitta europaea), 

starlings (Sturnus vulgaris) and great and middle spotted and green woodpeckers 

(Dendrocopus major, D. medius and Picus viridis). In order to assess the effects of 

competition for nesting cavities, we determined the abundance of parakeets and common 

native hole-nesters (nuthatches, starlings, great spotted and green woodpecker, jackdaw 

(Corvus monedula) and stock dove (Columba oenas)) in 44 study sites in the Brussels Capital 

Region from 2004 to 2006 using point counts. Results show that parakeet abundance was 

highest in forests or parks surrounded by built-up areas and parakeet numbers were strongly 

associated with cavity availability, suggesting that this may be a limiting factor (Strubbe & 

Matthysen 2007). After accounting for habitat and landscape variables influencing the 

abundance of native hole-nesters, we found that parakeet abundance was a significant 

predictor of nuthatch density, suggesting competition for nesting cavities (Strubbe & 

Matthysen 2007). In order to verify these findings, we set up an experimental manipulation of 

cavity availability in two city parks in the Brussels metropolitan area. In these sites, we 

blocked all cavities in which parakeets were found breeding in 2006, thereby severely 

reducing the cavity availability. Nuthatch breeding cavities were not blocked, but in the next 

breeding season we observed a significant decline in nuthatch breeding densities, largely due 

to cavity takeover by parakeets (Figure 1, Strubbe & Matthysen 2009b). 

Figure 1. Number of nests (mean ± SE) of (a) nuthatches and (b) 

parakeets at two experimental sites (black dots) and four control 

sites (white dots) before and after blocking of ring-necked 

parakeet breeding cavities.

Nuthatches defend their cavities by adjusting the entrance size of cavities to their own size by 

plastering up the entrance with mud. However, this does not protect the nuthatches from ringnecked 

parakeets, as parakeets start laying eggs already at the end of February (pers. obs.) 

while in western Europe, nuthatches start breeding only in the second half of April 

(Matthysen 1998). When breeding cavities were blocked, parakeets were forced to look for 

new breeding sites and the difference in timing of breeding enabled the parakeets to take the 

best nesting sites first. 

Parakeet range expansion in Belgium and consequences for native avifauna. 

Although the bulk of the parakeet population still occurs in the Brussels metropolitan area, 

parakeets can now be found breeding in a range of about 40 km around their release site and 

they are steadily increasing their range (Vermeersch et al. 2004, Vermeersch et al. 2006). To 

assess the potential distribution of parakeets in Belgium, we used a species distribution model 

(SDM) to identify areas suitable for parakeets. SDM use information on species occurrence 

data and environmental variables to generate statistical functions characterizing a species’ 

ecological requirements, and these functions are then projected onto the geographical area of 

interest to obtain the species (potential) distribution (Guisan & Zimmermann 2000). Using a 

dataset of ± 400 point locations of breeding parakeets and a number of habitat variables 

derived from region-wide land-use and forestry maps, we modeled parakeet distribution using 

an Ecological Niche Factor Analysis (ENFA, Hirzel et al. 2002). ENFA is a presence-only 

method and searches for an environmental ‘envelope’ characterizing the areas in which the 

species is present. Results show that parakeet distribution is mainly governed by variables 

representing cavity availability (i.e. the presence of parks and old forests) and parakeets are 

again strongly associated with urban habitats (Strubbe & Matthysen 2009c). The resulting 

distribution map (not shown here) shows that parakeets have ample room to spread into. Most 

highly suitable habitats are found along the urbanized north-south axis from Brussels to 

Antwerp, indicating that known nuthatch strongholds such as the regions south and east of 

Antwerp are highly likely to be invaded (Strubbe & Matthysen 2009c). 

However, as the real threat this parakeet poses depends on a combination of its 

potential distribution and abundance and its per capita impact on native birds, we not only 

need an assessment of its potential geographic distribution but also an estimate of its 

abundance (Parker et al. 1999). In order to obtain an estimate of the parakeets' expected 

abundance, we took advantage of recent improvements in the SDM field, and reanalyzed our 

data of parakeet abundance (see above, but point counts now converted to breeding densities, 

pairs/ha) using Boosted Regression Trees (Elith et al. 2008), a new technique capable of 

modeling abundance. BRT, coupled with the availability of exceptionally detailed geographic 

data layers containing information on land-use and vegetation types (the Biological Valuation 

Map and the Forest Reference Layer), allowed us to model parakeet abundance across the 

country at the level of individual forest patches. Region-wide predictions of nuthatch 

abundance are obtained using patch-level data obtained from a breeding bird atlas project. In 

order to quantify the parakeets' impact on nuthatches, we applied a regression model on our 

Brussels dataset of parakeet and nuthatch abundances, and we use the correlation coefficient 

between parakeet and nuthatch abundance as an indicator of competitive strength. The 

competition between parakeets and nuthatches was then quantified by superimposing their 

abundance maps and applying the competition coefficient, resulting in an estimate of the 

number of nuthatches that will be lost when parakeets have occupied all suitable sites. 

55

56 

Our models predict a mean number of about 22 000 ring-necked parakeet breeding 

pairs (90 % confidence limits from 9 000 to 39 800 pairs) indicating that these parakeets 

could become one of the most numerous cavity-nesting birds in the region. BRT results 

confirm that parakeet abundance is highest in older forests in urban environments, while there 

is also evidence that they prefer smaller, more fragmented forests (Strubbe, Graham & 

Matthysen in prep.). The resulting abundance map (Figure2) indicates that parakeets are 

expected to reach their highest densities along the strongly urbanized south-north axis from 

Brussels to Antwerp, while eastwards, there also is ample habitat for the parakeet to spread 

into, with high predicted abundances around the cities of Leuven and south of Hasselt. 

Westwards, there is less suitable habit and moderately high abundances are only found around 

the cities of Ghent and Bruges. Our nuthatch model estimates the number of nuthatches to be 

4 646 (2 929 - 6 776) breeding pairs, which agrees very well with the 4 740 to 5 750 pairs 

estimated in the Flemish and Brussels Breeding Bird Atlas (Vermeersch et al. 2006, Weiserbs 

& Jacob 2007). As expected from previous results, nuthatch numbers were negatively 

associated with parakeet abundance and we were able to extract a competition coefficient, 

quantifying the effect of parakeets on nuthatch densities. As our predictions of parakeet and 

nuthatch abundance and our estimate of the competition coefficient all have associated 

uncertainties (i.e. confidence intervals) , this leads to a number of possible scenarios. 

Figure 2. Region-wide predicted parakeet abundance. Number of breeding pairs per 5x5 km UTM 

grid were obtained by summing the predicted parakeet abundances for each forest fragment within 

each grid cell. 

Figure 3. Region-wide, long-term impact of parakeets on nuthatches according to a moderate 

scenario, i.e. taking the mean values for both parakeet and nuthatch abundance and competition 

strength. Number of breeding pairs lost per 5x5 km UTM grid were obtained by summing the 

predicted nuthatch losses for each forest fragment within each grid cell.

Thus, we calculated the region-wide, long-term parakeet impact using the mean estimates and 

the upper and lower confidence limits of our predictions and these calculations show that 

despite the high predicted parakeet abundances, total impact on nuthatches will probably only 

be small, and even in a worst case scenario, i.e. the maximum estimate for parakeet 

abundance and competition strength, only one third of the nuthatch population would be at 

risk. A more moderate scenario, i.e. taking the mean values for both parakeet and nuthatch 

abundance and competition strength, indicates a loss of 11 % of the nuthatches (i.e. 551 pairs, 

Strubbe, Graham & Matthysen, in prep.). In Figure 3, we visualized the predicted parakeet 

impact according to this moderate scenario. To conclude, we argue that the establishment of 

ring-necked parakeets should be prevented, but that in areas where they are currently present, 

there is no imminent ecological threat that calls for an eradication campaign. 

References 

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birds in Europe. Proceedings of the Royal Society B-Biological Sciences, 276: 47-53. 

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Elith J., Leathwick J. R. & Hastie T. 2008. A working guide to boosted regression trees. 

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Guisan A. & Zimmermann N. E., 2000. Predictive habitat distribution models in ecology. 

Ecological Modelling, 135: 147-186. 

Hirzel A., Hausser J., Chessel D. & Perrin N., 2002. Ecological Niche Factor Analysis : How 

to compute habitat suitability maps without absence data? Ecology, 83: 2027–2036. 

Lever C., 2005. Rose-ringed Parakeet (Ring-necked Parakeet) Psittacula krameri, p.: 124-130 

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Liu H. & Stiling P., 2006. Testing the enemy release hypothesis: a review and meta-analysis. 

Biological Invasions, 8: 1535-1545. 

Martin K. & Eadie J. M., 1999. Nest webs: A community-wide approach to the management 

and conservation of cavity-nesting forest birds. Forest Ecology and Management, 115: 

243-257. 

Matthysen E., 1998. The Nuthatches, London: T & A. D. Poyser. 

Parker I. M., Simberloff D., Lonsdale W. M., Goodell K., Wonham M., Kareiva P. M., 

Williamson B., Von Holle B., Moyle P. B., Byers J. E. & Goldwasser L., 1999. Impact: 

Toward a Framework for Understanding the Ecological Effects of Invaders. Biological 

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Robb G. N., McDonald R. A., Chamberlain D. E., Reynolds S. J., Harrison T. J. E. & Bearhop 

S., 2008. Winter feeding of birds increases productivity in the subsequent breeding 

season. Biology Letters, 4: 220-223. 

Shwartz A., Strubbe D., Butler C., Matthysen E. & Kark S., 2009. The effect of enemyrelease 

and climate conditions on invasive birds: a regional test using the rose-ringed 

parakeet Psittacula krameri as a case study. Diversity and Distributions, 15: 310-318. 

Strubbe D. & Matthysen, E., 2007. Invasive ring-necked parakeets Psittacula krameri in 

Belgium: habitat selection and impact on native birds. Ecography, 30: 578-588. 

Strubbe D. & Matthysen E., 2009a. Establishment success of invasive ring-necked and monk 

parakeets in Europe. Journal of Biogeography: Accepted manuscript. 

Strubbe D. & Matthysen E., 2009b. Experimental evidence for nest-site competition between 

invasive ring-necked parakeets (Psittacula krameri) and native nuthatches (Sitta 

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Strubbe D. & Matthysen E., 2009c. Predicting the potential distribution of invasive ringnecked 

parakeets Psittacula krameri in northern Belgium using an ecological niche 

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Patterns of Prunus serotina invasion in two contrasting forests 

Margot VANHELLEMONT 1 , Lander BAETEN 1 , Martin HERMY 2 & Kris VERHEYEN 1 

1 

Laboratory of Forestry, Ghent University, Geraardsbergsesteenweg 267, B-9090 Gontrode, 

Belgium 

2 

Division Forest, Nature and Landscape, K.U.Leuven, Celestijnenlaan 200e, B-3001 Leuven, 


Introduction 

Prunus serotina Ehrh., a North-American tree species, is considered an invasive species in 

Western Europe. Most studies in its introduced range focused on areas heavily invaded by P. 

serotina. Nonetheless, the presence and abundance of P. serotina in these areas still reflects 

the massive plantings of the past (Starfinger et al. 2003). The large-scale plantings resulted in 

a high propagule pressure of P. serotina. Consequently, P. serotina exhibited a considerable 

invasion rate in these areas, which lead to problems in silviculture and nature conservation 

(Starfinger et al. 2003). Since actual rates of invasion appear to be largely determined by 

propagule pressure (Von Holle & Simberloff 2005), we wanted to study the spread of P. 

serotina in an area characterized by a far lower propagule pressure where the species has not 

been introduced deliberately. Would we still label P. serotina an aggressive invader in these 

circumstances? Besides, P. serotina has not yet fully occupied its potential range in Europe, 

and the spread of the species is thought to be limited by dispersal (Zerbe & Wirth 2006, 

Verheyen et al. 2007). To develop appropriate management strategies, we should gain insight 

into the factors that affect the colonization rate of P. serotina in new sites. 

In this abstract, we compare the results of two studies on 70 years of forest 

development in areas with a low propagule pressure of P. serotina. The Liedekerke forest 

reserve (Belgium) and the Ossenbos forest reserve (the Netherlands) were particularly 

appropriate for our research because they have not been managed for over sixty years and P. 

serotina established spontaneously during the forest development. Based on the observed 

patterns of P. serotina colonization in these forests, we wanted to answer the following 

questions: which factors influenced the spread of P. serotina, and did P. serotina act as an 

invasive species in the studied forests? 

Materials & Methods 

A detailed description of the materials and methods can be found in Vanhellemont et al. 

(2009) and Vanhellemont et al. (in press) for the studies in the Liedekerke forest reserve and 

the Ossenbos forest reserve, respectively. Table 1 shows the main characteristics of the 

studied forest reserves. For the two forest reserves, we reconstructed the P. serotina invasion 

based on cadastral maps and aerial photographs, tree ring analysis, forest inventories and 

regeneration data. 

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Study area & data collection 

Liedekerke: The Liedekerke forest reserve has been forested until 1926, when all the trees in 

the study area were cut. The subsequent management resulted in heathland and coppice. After 

WWII, management ceased, and the vegetation developed into a mixed deciduous forest with 

a herb layer dominated by Rubus fruticosus agg. In 1986, a 12.9 ha study area was defined 

and 65 circular plots (radius 15 m) were installed at the intersections of a 40 m x 50 m grid. 

These plots were used to study the changes in structure and species composition of the forest, 

based on ten-year interval data. The 65 plots were inventoried in 1986; 31of the plots were 

sampled in 1996; and all the 65 plots were re-inventoried in 2006. Data were collected for the 

tree, shrub and herb layer. In addition, aerial photographs of the period 1944–1986 were used 

to gain insight into the vegetation development after WWII. The colonization of P. serotina 

within the study area was analyzed by (1) identifying and locating by GPS the initial points of 

colonization within the entire forest reserve, (2) setting up an age distribution for the P. 

serotina in the inventory plots, and (3) predicting the presence of P. serotina in the 65 plots 

based on the plot history, the connectivity to seed trees, the basal area, the percentage of basal 

area made up by shade-casting woody species, the change in basal area between 1986 and 

2006, and the percentage of cover by Rubus spp. in the herb layer. For 54 subcanopy P. 

serotina, diameter growth and age were determined based on stem cross sections or tree cores. 

Ossenbos: The Ossenbos forest reserve is situated in a landscape matrix with forest patches, 

heathlands, and bare sand. The forest developed on the heathlands and drift sands around a 

mound that had been planted with pine (Pinus sylvestris L.) and oak (Quercus robur L.) 

around 1832. The game densities are extremely high in the forest reserve: ca. 1 ha -1 of red 

deer (Cervus elaphus), roe deer (Capreolus capreolus), and wild boar (Sus scrofa). Data were 

collected in 40 circular plots (radius 12.6 m) located randomly at the intersections of a 50 m x 

50 m grid in the entire forest and in a 70 m x 140 m study area (the core area) in the eldest 

part of the forest reserve. Position, diameter, and height were measured for the living trees. 

The height and number of saplings were recorded in the circular plots. In the core area, 

regeneration was enumerated in four classes: seedlings-of-the-year, seedlings < 20 cm, 20 cm 

< seedlings < 120 cm, and seedlings > 120 cm. Besides, in the core area, the spatial position 

of the seed-bearing P. serotina trees was noted, and we counted seeds in litter samples. In 

addition, diameter growth was studied for P. serotina growing below P. sylvestris or Q. robur 

(13 samples), below P. serotina (11), or in a canopy gap (9) in the core area. 

Table 1 Characteristics of the two studied forest reserves: area (ha), geographic location, soil type, 

minimum and maximum monthly mean temperature (T, °C), mean annual precipitation (Precip., mm), 

mean basal area (BA, m 2 ha -1 ), and the main tree species 

Site Area Location Soil type T Precip. BA Tree species * 

(ha) N E (°C) (mm) (m 2 ha -1 ) 

Liedekerke 21 50°52’ 4°07’ sandy loam 2.5–17.2 821 31.1 

Betula, 

Quercus 

Ossenbos 54 52°08’ 5°48’ sand 2–17 850 26.6 Pinus 

* Betula pendula Roth & Betula pubescens Ehrh.; Quercus petraea (Matt.) Liebl., Quercus 

robur L. & Quercus rubra L.; Pinus sylvestris L.

Data analysis 

For both forest reserves, we first reconstructed the invasion process. Next, we tried to identify 

factors affecting the establishment and growth of P. serotina. 

Liedekerke: Forest development was analyzed based on the comparison of basal area and stem 

density for 1986–2006 (paired samples t-tests) and 1986–1996–2006 (repeated-measures 

GLM). We compared the characteristics of plots with and without P. serotina (t-tests) and 

predicted the presence/absence of P. serotina seedlings/saplings, shrubs, and trees in the plots 

(logistic regression). Apart from P. serotina, we also looked at Sorbus aucuparia L., a species 

that frequently co-occurs with P. serotina (Verheyen et al. 2007). In addition, diameter 

growth (multiple linear regressions) and allometric relationships between age and height or 

diameter at breast height (curve estimation procedure) were studied for P. serotina. 

Ossenbos: We focused on the core area, located in the oldest part of the forest, and used the 

data on the circular plots to check whether the patterns observed in the core area held for the 

younger parts of the forest. First, we analyzed the spatial patterns of trees and shrubs in the 

core area to determine the past establishment of P. serotina (bivariate Ripley’s L). Second, 

inverse modelling was used to calculate dispersal kernels for P. serotina seed and seedlings in 

the core area. Third, presence/absence of P. serotina was modelled with binary logistic 

regressions based on the plot characteristics such as basal area of the tree and shrub layers (m 2 

ha -1 ), stem density (ha -1 ), and canopy openness (%). Fifth, to explain the abundance of the 

seedlings/saplings of a species, we performed a data reduction (PCA) on the plot 

characteristics, and calculated Pearson correlations between the seedling/sapling densities and 

the principal components. Last, we studied allometric relationships for P. serotina age (curve 

estimation procedure) and diameter growth (multiple linear regressions). Interactive effects 

between the canopy tree neighbourhood and the age of the studied P. serotina on the achieved 

diameter or height were investigated with ANCOVA analysis. 

Results 

For the two forest reserves, we present the results on P. serotina spread, on the factors 

affecting its presence/abundance, and on P. serotina growth. 

Liedekerke: The first P. serotina established around 1970–1975 on sites with high light 

availability. Further P. serotina colonization started 10–15 years later, when these trees 

presumably started producing seeds. Prunus serotina has spread through the study area. Of 

the 65 plots, 14 were colonized in 1986, 33 in 2006. For S. aucuparia, the number of plots 

was 38 in 1986 and 60 in 2006. The largest increase in number of plots occupied occurred 

between 1986 and 1996. For the 30 plots which were inventoried thrice, P. serotina was 

present in 6 (1986), 15 (1996), and 17 (2006) plots and S. aucuparia in 17 (1986), 28 (1996), 

and 29 (2006) plots. In 2006, P. serotina seedlings and saplings (age < 12 yr) occurred in only 

10 % of the plots. 

Plots with P. serotina were characterized by a higher connectivity to seed trees, a 

higher overall basal area, and a higher basal area of shade-casting trees. Plots without P. 

serotina had a higher basal area of light-demanding trees and a high cover of Rubus. A plot 

was more likely to be colonized by P. serotina if its connectivity to the seed trees was high. 

The mean diameter growth of the subcanopy P. serotina in 2001–2006 was 2.8 ± 0.2 

mm yr -1 . Diameter growth (mm yr -1 ) was determined by the diameter at breast height (dbh, in 

61

62 

cm) and age of the tree and competition with neighbouring trees (CI): diameter growth = 

4.911 + 0.487dbh - 0.028dbh² - 0.077age - 0.071CI (R² = 0.80). The relationship between age 

(yr) and dbh (cm) was described most accurately by: age = 7.254dbh 0.402 (R² = 0.79). 

Ossenbos: Prunus serotina first became established in the Ossenbos around 1940. Successful 

recruitment of P. serotina into the tree layer occurred mainly in gaps of the P. sylvestris - Q. 

robur canopy layer. Prunus serotina shrubs occurred more often below the light-demanding 

P. sylvestris and Q. robur than below the shade-casting P. serotina. In 2006, P. serotina was 

by far the most abundantly regenerating species and the only species with seedlings taller than 

120 cm. Prunus serotina was found in all circular plots, and the high densities of seedlings 

smaller than 20 cm point towards the build-up of a persistent seedling bank. 

The dispersal kernels showed that seeds and small seedlings of P. serotina mostly 

occurred close to a source tree while large seedlings showed the highest densities between 

10–15 m from the source tree. Accordingly, the abundance of P. serotina regeneration was 

correlated significantly with seed density for small seedlings, and with basal area/canopy 

openness for larger seedlings. Ln-transformed abundances of P. serotina saplings were 

significantly correlated with the principal component that combined maximum tree height, 

basal area of tree and shrub layers, and stem density. 

Radial growth of P. serotina was related to dbh: ln growth = 0.083 + 0.491dbh (R 2 = 

0.66). Prunus serotina growing in gaps and below P. serotina showed a clear relationship 

between age and dbh (R 2 = 0.96 and 0.86) and between age and height (R 2 = 0.94 and 0.83). 

The increases of dbh and height with age were higher in gaps than below P. serotina 

(interaction: p = 0.013 and p = 0.034). 

Conclusion 

The initial P. serotina status was comparable in the two forest reserves: P. serotina had not 

been planted and the initial propagule pressure was low. Nonetheless, the outcome of the P. 

serotina invasion process contrasted sharply between the two studied forests: P. serotina was 

omnipresent and very abundant in the Ossenbos while the species did not act as an aggressive 

invader in the Liedekerke forest reserve. Consequently, it appears to be important to study an 

invasive species and the recipient ecosystem jointly and to gear the control measures to the 

characteristics of the recipient ecosystem. 

Long-distance dispersal events and windows of opportunity triggered the invasion of 

P. serotina. Further colonization was directed by connectivity to seed sources and light 

availability. In the Liedekerke forest reserve, the presence of native shrub species, the quick 

canopy closure, and the recalcitrant herb layer seemed to hamper further P. serotina 

establishment. Conversely, in the Ossenbos forest reserve, the high herbivore pressure 

favoured P. serotina above native species, which resulted in P. serotina dominance. 


The authors thank the Agency for Nature and Forests and the Dutch Ministry of Defence for the 

permission to work in the forest reserves; Bart De Cuyper, Diego Van Den Meersschaut, and Alterra 

for the data-collection in 1986, 1996, and 2003; Els De Lathauwer, the Research Institute for Nature 

and Forest (INBO), and Lotte Wauters for their help with the data-collection in 2006 and 2007. The 

first and the second author held a scholarship from the Research Foundation – Flanders (FWO) and the 

Institute for the Promotion of Innovation by Science and Technology in Flanders (IWT-Vlaanderen), 

respectively. The study was supported financially by the Special Research Fund of Ghent University 

(BOF).

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centuries. Biological Invasions 5:323–335. 

Vanhellemont M., Wauters L., Baeten L., Bijlsma R.-J., De Frenne P., Hermy M. & Verheyen 

K., in press. Prunus serotina unleashed: invader dominance after 70 years of forest 

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Vanhellemont M., Verheyen K., De Keersmaeker L., Vandekerkhove K. & Hermy M., 2009. 

Does Prunus serotina act as an aggressive invader in areas with a low propagule pressure? 

Biological Invasions Biological Invasions 11:1451–1462. 

Verheyen K., Vanhellemont M., Stock T. & Hermy M. 2007. Predicting patterns of invasion 

by black cherry (Prunus serotina Ehrh.) in Flanders (Belgium) and its impact on the forest 

understorey community. Diversity and Distributions 13:487–497. 

Von Holle B. & Simberloff D., 2005. Ecological resistance to biological invasion 

overwhelmed by propagule pressure. Ecology 86:3212–3218. 

Zerbe S. & Wirth P., 2006. Non-indigenous plant species and their ecological range in Central 

European pine (Pinus sylvestris L.) forests. Annals of Forest Science 63:189–203. 

63

Soil arthropods associated to the invasive Senecio inaequidens 

compared to the native S. jacobaea (Asteraceae) 

Valérie VANPARYS 1 , Pierre MEERTS 2 , Guy JOSENS 3 & Anne-Laure JACQUEMART 1 

1 

Research team in Genetics, Reproduction, Populations, UCL, Louvain-la-Neuve, Belgium. 

Contact: valerie.vanparys@gmail.com 

2 

Laboratoire de Génétique et Ecologie végétales, ULB, Bruxelles, Belgium. 

3 

Laboratoire de Systématique et Ecologie animales, ULB, Bruxelles, Belgium. 

Introduction 

Soil arthropods play a significant role in soil processes (e.g. nutrient cycling). Diversity of 

soil invertebrates has been shown to positively influence plant diversity (De Deyn et al. 

2003). In turn, plant diversity, and even plant species identity, can influence soil communities 

(Armbrecht et al. 2004, Ayres et al. 2006, Wardle 2006). In the particular case of plant 

invasions, several authors have shown that some animal groups in the soil can be affected by 

the change in plant community. For instance, Belnap & Philips (2001) showed that invasion 

by the annual grass Bromus tectorum (western US) decreased the overall abundance of soil 

invertebrates as well as the species richness of soil arthropods. These changes in soil 

communities during invasion have been proposed to facilitate the development of invasive 

plants, creating positive feedbacks (Callaway et al. 2004, Wolfe & Klironomos 2005). 

In the present study, the general hypothesis is that Senecio inaequidens (Asteraceae), 

invasive in Europe, and the native S. jacobaea (synonym: Jacobaea vulgaris) are not 

associated with the same community of soil arthropods, in terms of abundance, taxonomic 

assemblage and diversity. 

Material and methods 

Senecio inaequidens is a perennial chamaephyte native to South Africa, whereas Senecio 

jacobaea is a biennial to perennial hemicryptophyte native to Europe, spending the first year 

as a rosette (Harper & Wood 1957). 

This study was performed in a 25-ha wildland site, situated in Antwerp (51° 14’ 

36.40’’ N, 4° 23’ 15.03’’ E), northern Belgium. Vegetation was dominated by S. inaequidens 

(present for at least 5 years), S. jacobaea, Festuca rubra, Geranium molle and Tanacetum 

vulgare. The soil was described as a sandy brown soil. A previous study revealed no 

difference in soil chemical properties nor in granulometry between invaded and uninvaded 

patches (Dassonville et al. 2008). There was, however, a higher phytomass and K 

concentrations in invasive plots. 

Soil fauna was sampled in the Antwerp site in 2006 on October 13 th . Three distinct 

zones were chosen, separated by approximately 10 m. The first zone was dominated by S. 

inaequidens (hereafter referred to as I), the second one by S. jacobaea (J) and in the third one 

the density was similar for both species and the soil was more gravelly (M). In each zone, 

four pairs of plants S. inaequidens - S. jacobaea were chosen, among the oldest and largest 

ones. Within a pair, plants were separated by 1.5 m at most. One soil core of 8 cm diameter 

and 5 cm depth (volume 251 cm³) was extracted, as close as possible to the root crown. A 

total of 24 samples were taken and placed on a Berlese-Tullgren extractor. 

65

66 

The abundance of soil fauna, i.e. numbers of individuals per sample, was compared 

between S. inaequidens and S. jacobaea by an ANOVA for effects of species, zone and their 

interaction. For the main taxa, the numbers of individuals were compared between the two 

plant species by Mann-Whitney U tests. A principal components analysis (PCA) was used to 

extract the axes summarising the community structure of soil samples. These analyses were 

performed with STATISTICA 7 (Statsoft 2006). In addition, Shannon-Wiener’s index (H’) 

was calculated for soil arthropod communities associated to S. inaequidens and S. jacobaea. 

A t-test was applied on these data to test for a difference between the two Senecio species. 

Results 

The overall abundance of soil fauna was significantly lower (F= 7.46, p= 0.014) under S. 

inaequidens in comparison with S. jacobaea, with respectively 76 ± 62 (or 15,300 individuals 

m -2 ) and 186 ± 153 individuals per sample (or 37,000 individuals m -2 ). This difference was 

essentially due to the Collembola Arthropleona (Table 1), which were very abundant in S. 

jacobaea samples (19,000 individuals m -2 ), while they were six fold less abundant in S. 

inaequidens samples (3,500 individuals m -2 ). Gamasid mites were twice more abundant in S. 

jacobaea samples compared to S. inaequidens, but this difference was only marginally 

significant. None of the other taxa differed significantly in their abundance under the two 

plant species. Notably, no significant difference was detected in the number of homopteran 

and heteropteran herbivores (Table 1). The effect of the zone was marginally significant on 

the abundance of soil fauna, with a tendency of higher densities in the mixed zone (Figure 1). 

No significant interaction zone*species was detected. 

Taxonomic richness was similar between S. inaequidens and S. jacobaea (Table 1). 

However, the Shannon diversity index was lower for S. jacobaea (H=1.54 and evenness = 

0.52) than for S. inaequidens (H= 1.96 and evenness = 0.71) and this difference was highly 

significant (T-test: t=-10.19, p

Table 1. Abundance of soil fauna expressed as the total number of individuals (N) for each 

identified taxon in soil samples collected under S. inaequidens and S. jacobaea (12 samples 

per species) and percentage of the total number of individuals (%). Feeding groups: 

herbivore (H), microbivore-mycophagous (M), predator (P), honeydew (*), saprophagous (S), 

undetermined or diverse (?). Numbers in brackets are the minima and maxima per sample. 

Mann-Whitney U tests for difference between the two plant species (for taxa present in more 

than half the samples): * and + indicate respectively significant (p

68 

Fact. 2 : 15,70% 

2 

0 

-2 

-4 

-6 

Jj 

Jm 

Jm 

Ii 

Ji Ij 

Ji 

Jj 

Ij 

Im Ij Ii Ji 

Jj 

JmJj 

Im 

Ii 

Ji Ii 

Im 

Im 

-8 

-8 -6 -4 -2 0 2 4 

Fact. 1 : 18,76% 

Figure 2. Results of the PCA for the abundance of the different taxa in soil samples for S. 

inaequidens and S. jacobaea: projection of taxa (A) and samples (B). Labels of samples 

mention the plant species (J for S. jacobaea or I for S. inaequidens), the sample No (1 to 12) 

and the zone (i, j or m). 

As lower abundance and higher diversity of arthropods were found under S. 

inaequidens, the possible impacts of the invasion by S. inaequidens are difficult to assess. Our 

study needs to be repeated in other sites, as the response of soil arthropods may differ 

between sites (Yeates & Williams 2001). 

In conclusion, this study showed that S. inaequidens and S. jacobaea were associated 

with similar soil arthropod assemblages, but that arthropod abundance was lower under S. 

inaequidens, whereas the diversity was higher, essentially due to collembolans Arthropleona. 

Further studies should focus on verifying this result in other sites with contrasting soil 

characteristics. Finally, more thorough identifications may help to understand our results. 

Ij 

Jm

References 

Armbrecht I., Perfecto I. & Vandermeer J., 2004. Enigmatic biodiversity correlations: Ant 

diversity responds to diverse resources. Science 304(5668): 284-286. 

Ayres E., Dromph K.M. & Bardgett R.D., 2006. Do plant species encourage soil biota that 

specialise in the rapid decomposition of their litter? Soil Biology & Biochemistry 38(1): 

183-186. 

Belnap J. & Phillips S.L., 2001. Soil biota in an ungrazed grassland: response to annual grass 

(Bromus tectorum) invasion. Ecological Applications 11:1261-1275. 

Briones M.J.I., Ineson P. & Sleep D., 1999. Use of δ13C to determine food selection in 

collembolan species. Soil Biology and Biochemistry 31(6): 937-940. 

Callaway R.M., Thelen G.C., Rodrigez A. & Holben W.E., 2004. Soil biota and exotic plant 

invasion. Nature 427 731-733. 

Dassonville N., Vanderhoeven S., Vanparys V., Hayez M., Gruber W. & Meerts P., 2008. 

Impacts of alien invasive plants on soil nutrients are correlated with initial site conditions 

in NW Europe. Oecologia 157(1): 131-140. 

De Deyn G.B., Raaijmakers C.E., Zoomer H.R., Berg M.P., de Ruiter P.C., Verhoef H.A., 

Bezemer T.M. & van der Putten W.H., 2003. Soil invertebrate fauna enhances grassland 

succession and diversity. Nature 422, no6933: 711-713. 

de Jong T.J. & van der Meijden E., 2000. On the correlation between allocation to defence 

and regrowth in plants. Oikos 88(3): 503-508. 

Duda J.J., Freeman D.C., Emlen J.M., Belnap J., Kitchen S.G., Zak J.C., Sobek E., Tracy M. 

& Montante J., 2004. Differences in native soil ecology associated with invasion of the 

exotic annual chenopod, Halogeton glomeratus. Biology and Fertility of Soils 38(2): 77- 

77. 

Garcia-Serrano H., Escarré J., Garnier E. & Sans X.F., 2005. A comparative growth analysis 

between alien invader and native Senecio species with distinct distribution ranges. 

Ecoscience 12(1): 35-43. 

Gratton C. & Denno R.F., 2005. Restoration of arthropod assemblages in a Spartina salt 

marsh following removal of the invasive plant Phragmites australis. Restoration Ecology 

13(2): 358-372. 

Harper J .L. & Wood W.A., 1957. Biological flora of the British Isles: Senecio jacobaea L. 

Journal of Ecology 45: 617-637. 

Loizzo M.R., Statti G.A., Tundis R., Conforti F., Bonesi M., Autelitano G., Houghton P.J., 

Miljkovic-Brake A. & Menichini F., 2004. Antibacterial and antifungal activity of 

Senecio inaequidens DC. and Senecio vulgaris L. Phytotherapy Research 18(9): 777-779. 

Wardle D.A., 2006. The influence of biotic interactions on soil biodiversity. Ecology Letters 

9(7): 870-886. 

Wolfe B.E. & Klironomos J.N., 2005. Breaking new ground: Soil communities and exotic 

plant invasion. BioScience 55(6): 477-487. 

Yeates G.W. & Williams P.A., 2001. Influence of three invasive weeds and site factors on 

soil microfauna in New Zealand. Pedobiologia 45(4): 367-383. 

69

Non-indigenous freshwater fishes in Flanders: status, trends and 

risk assessment 

Hugo VERREYCKEN * , Gerlinde VAN THUYNE & Claude BELPAIRE 

Research Institute for Nature and Forest (INBO), Duboislaan 14, 1560 Hoeilaart 

*Corresponding author; tel. 02 658 04 26 & e-mail: hugo.verreycken@inbo.be 

Abstract 

At least eighteen non-indigenous freshwater fish species were reported to occur in the wild 

within the territory of Flanders. Nine are considered naturalized while the others are 

acclimatized and do not form self-sustaining populations. Nine of the introductions occurred 

prior to 1950, with the other nine species introduced since then. This contribution reviews the 

available information on these introductions, and evaluates a decade of data from fisheries 

surveys to assess the recent development of these non-indigenous populations. Gibel carp 

Carassius gibelio and topmouth gudgeon Pseudorasbora parva are the most widespread of 

the non-indigenous species in Flemish waters, and both continue to expand their ranges. A 

reduction in range has been observed in brown bullhead Ameiurus nebulosus only. Only four 

species occur in all eleven river basins while eight species are restricted to one or two basins 

and often only one specimen was found during fish stock assessments. We also discuss nonindigenous 

fish species that are likely to colonize Flanders inland waters in the near future. 

For all non-indigenous freshwater fish species present and expected to appear soon, different 

risk analysis tools (FISK and ISEIA) were used to screen these species for their possible 

invasiveness. Although scores from FISK and ISEIA differ for some species, gibel carp and 

topmouth gudgeon were in both assessments classified as ‘highest risk’ species in relation to 

their potential invasiveness. 

Introduction 

In 1995, the Research Institute for Nature and Forest (INBO) (then Institute for Forestry and 

Game Management) started a monitoring network on freshwater fishes. Fish stock 

assessments were performed on a regular basis at more than 800 locations all over Flanders. 

The data collected during these fish stock assessments were brought together in an online 

database VIS or Vis Informatie Systeem (Fish Information System) which can be consulted at 

http://vis.milieuinfo.be. At present, data of more than 200 000 native and non-native fishes 

with their individual lengths and weights are in this database. All these data are 

georeferenced. 

This paper discusses the presence and status of the non-indigenous fish species in 

Flanders (Belgium). Data of non-native fishes from the VIS-database were analysed for trends 

in numbers and biomass over the period 1996 - 2005. Potential invasiveness of all nonindigenous 

freshwater fish species present and expected to appear soon was assessed using 

the FISK tool by Copp et al. (2005a) and the ISEIA protocol by the Belgian Forum on 

Invasive Species (Branquart, 2007). Both tools classify the invasiveness of the species into 

low, medium or high, and results from both methods were compared. 

71

72 

Status of non-indigenous fish species in Flanders 

Louette et al. (2001) found evidence in the literature of 47 considered, attempted or successful 

introductions of fishes in Belgium since 1800; of these, 23 were partially successful (i.e. 

recorded in public waters after introduction or known to be reproducing). 

Since 1990 at least 18 non-native freshwater fish species were reported during fish stock 

assessments by INBO, Agency for Nature and Forest (ANB) and several universities in 

Flanders. Nine species are naturalised and maintain self-sustaining populations while nine are 

acclimatised only. Two of these acclimatised species however occur exclusively near cooling 

water discharges of power plants. These non-native species constitute more than 35 % of the 

total number of freshwater fish species in Flanders. Most species originate from Asia and N. 

America, with seven and six species respectively, three species come from Eastern Europe 

and two from Africa. Nine species were introduced pre-1950s and nine species since then. 

These 18 species are listed in table 1. 

To illustrate that non-indigenous fish species are not a localised phenomenon, table 2 

shows the distribution over the 11 Flemish river basins. Four species occur in all river basins, 

and one in all but one. Seven species occur in one basin only and the other six were found in 

Table 1: Species and common names of the 18 non-indigenous freshwater fishes occurring in 

Flanders, with their continent of origin (AS, Asia; EE, Eastern Europe; AFR, Africa; NA, North 

America), suspected pathways (AQ, aquaculture; OR, ornamental; AN, angling or bait fish; BC, 

biological control; UN, unintentional), date of introduction (year or period; c., century) and current 

status (Copp et al., 2005b; A, acclimatized; N, naturalized; A*, acclimatized only in restricted areas at 

cooling water discharges of power plants). 

Latin name Common name Origin Introduction Pathway(s) Status 

Acipenser baeri Siberian sturgeon AS 1990s AQ, OR A 

Ameiurus nebulosus Brown bullhead NA 1871 AQ, OR N 

Aspius aspius Asp EE 1984 AN N 

Carassius auratus Goldfish AS 17th c. OR A 

Carassius gibelio Gibel carp AS or EE 17th c. UN N 

Clarias gariepinus African catfish AFR 1980s AQ A* 

Ctenopharyngodon 

idella 

Grass carp AS 1967 BC A 

Cyprinus carpio Common carp EE 13th c. AQ N 

Hypophthalmichthys 

molitrix 

Silver carp AS 1975 BC A 

Hypophthalmichthys 

Bighead carp AS 1975 BC A 

nobilis 

Ictalurus punctatus Channel catfish NA 1884 AQ A 

Lepomis gibbosus Pumpkinseed NA 1885 OR N 

Oncorhynchus mykiss Rainbow trout NA 1884 AQ, AN A 

Oreochromis niloticus Nile tilapia AFR 1990 AQ A* 

Pimephales promelas Fathead minnow NA 1984 AN N 

Pseudorasbora parva Topmouth gudgeon AS 1992 UN, AN N 

Sander lucioperca Pikeperch EE 1890 AN N 

Umbra pygmaea 

Eastern 

mudminnow 

NA 1920 OR, AQ N

two to five river basins. The species with a wide distribution over Flanders often also are 

widespread within the individual river basins. Brown bullhead, eastern mud minnow and 

pumpkinseed occur at highest densities in the northeast of Flanders. This region is 

characterised by a high concentration of pond fish farms, which were created in abandoned 

peat diggings. All of these three North American species are common or widespread there. 

As Flanders is surrounded by countries with similar habitats and climates and in which 

the catchments of the rivers Danube, Rhine, Meuse and Scheldt are connected by canals, it is 

to be expected that new species as white-finned gudgeon Romanogobio belingi (present in 

Germany and the Netherlands), vimba Vimba vimba (reported from the Netherlands), round 

goby Neogobius melanostomus (also reported from the Netherlands), tubenose goby 

Proterorhinus semilunaris (the Netherlands, France and Germany) and bighead goby 

Neogobius kessleri (found in Germany) may enter Flanders in the very near future. Also the 

highly invasive Amur or Chinese sleeper Perccottus glenii was already observed in the 

Danube and may be a new invader in the years to come. 

Table 2: Occurrence of non-indigenous fishes in river basins of Flanders [n = 11; Lower Scheldt (LS), 

Upper Scheldt (US), Bruges Polders (BP), Demer (Dm), Dender (Dn), Dijle (Di), Ghent Canals (GC), 

Leie (Le), Meuse (Me), Nete (Ne) and Yser (Ys)] expressed as percentage of sites where a nonindigenous 

species is present compared to the total number of sample sites per river basin [VR, very 

rare (≤ 2.0 %); R, rare (2.1 – 10.0 %); C, common (10.1 – 25.0 %); W, widespread (> 25 %)] 

LS US BP Dm Dn Di GC Le Me Ne Ys N 

Gibel carp R C W W R C W C C C W 11 

Topmouth gudgeon R R C W C C C C R R C 11 

Pikeperch C VR R VR VR R R R R C C 11 

Common carp R R W C R C W R R C W 11 

Pumpkinseed C VR VR W R R VR R C W 10 

Brown bullhead VR C VR R C 5 

Rainbow trout VR R R R 4 

Goldfish VR VR VR VR 4 

Fathead minnow VR R VR 3 

Eastern mudminnow C C C 3 

Grass carp VR VR 2 

Asp VR 1 

Silver carp VR 1 

Bighead carp VR 1 

Siberian sturgeon VR 1 

Channel catfish VR 1 

African catfish VR 1 

Nile tilapia VR 1 

Nb 9 8 5 13 5 7 5 6 10 9 5 

Total number of basins where a species is present (N); Total number of non-indigenous pecies 

in a basin (Nb). 

73

74 

Trends 

At 487 site-specific surveys between 1996 and 2005, trends in frequency of occurrence and 

abundance were investigated. These sites were fished at least twice in this period, once 

between 1996 and 2000, and again between 2001 and 2005 with a minimum of 3 years in 

between. On 20 % of the sites no fish were caught. Standardised fishing techniques like 

electrofishing and fyke nets or a combination of both were used during these fish stock 

assessments. To evaluate the trends, a logistic regression was used to model the changes in 

frequency of occurrence and a linear mixed model for trends in abundance. 

Most non-indigenous species show an increasing trend in number of sites they occupy; in 

particular topmouth gudgeon (p

≥19 by the UK assessors and are considered high risk invasive species in FISK (Copp et al, 

2008). 

Two non-native fish species to be expected in Flanders, Amur sleeper and round goby, 

were categorized as potentially invasive and are thus on the Harmonia alert list (A0category)(Harmonia 

database, 2009). 

Despite the fact that both tools use different scoring systems, they manage to 

categorize the fishes more or less in the same ‘invasiveness classes’. FISK and ISEIA 

therefore both represent useful and viable tools to aid decision- and policymakers in assessing 

and classifying freshwater fishes according to their potential invasiveness. 

References 

Branquart E. (Ed.), 2007. Guidelines for environmental impact assessment and list 

classification of non-native organisms in Belgium, http://ias.biodiversity.be/ias/, last 

assessed 5 May 2009. 

Copp G.H., Garthwaite R. & Gozlan R.E., 2005a. Risk identification and assessment of nonnative 

freshwater fishes: a summary of concepts and perspectives on protocols for the UK. 

Journal of Applied Ichthyology 21, 371–373. 

Copp G.H., Bianco P.G., Bogutskaya N., Erős T., Falka I., Ferreira M.T., Fox M.G., Freyhof 

J., Gozlan R.E., Grabowska J., Kováč V., Moreno-Amich R., Naseka A.M., Peňáz M., 

Povž M., Przybylski M., Robillard M., Russell I.C., Stakėnas S., Šumer S., Vila-Gispert 

A. & Wiesner C., 2005b. To be, or not to be, a non-native freshwater fish? Journal of 

Applied Ichthyology 21 (4), 242-262. doi: 10.1111/j.1439-0426.2005.00690.x 

Copp G.H., Vilizzi L., Mumford J., Fenwick G.V., Godard M.J. & Gozlan R.E., 2008. 

Calibration of FISK, an invasiveness screening tool for non-native freshwater fishes. Risk 

Analysis (online: DOI: 10.1111/j.1539-6924.2008.01159.x). 

Harmonia database 2009, Belgian Forum on Invasive Species, accessed on 5 May 2009 from: 

http://ias.biodiversity.be 

Louette G., Anseeuw D., Gaethofs T., Hellemans B., Volckaert F.A.M., Verreycken H., Van 

Thuyne G., De Charleroy D., Belpaire C., Declerck S., Teugels G.G., De Meester L. & 

Ollevier F. (2001). Ontwikkeling van een gedocumenteerde gegevensbank over uitheemse 

vissoorten in Vlaanderen met bijkomend onderzoek naar blauwbandgrondel. Eindverslag 

van project VLINA 00/11. Studie uitgevoerd voor rekening van de Vlaamse Gemeenschap 

binnen het kader van het Vlaams Impulsprogramma Natuurontwikkeling in opdracht van 

de Vlaamse Minister bevoegd voor Natuurbehoud. D/2002/3241/136. [In Dutch] 

Pheloung P.C., Williams P.A. & Halloy S.R., 1999. A weed risk assessment model for use as 

a biosecurity tool evaluating plant introductions. Journal of Environmental Management, 

57, 239-251. 

Verreycken H., Anseeuw D., Van Thuyne G., Quataert P. & Belpaire C., 2007. The nonindigenous 

freshwater fishes of Flanders (Belgium): review, status and trends over the last 

decade. Journal of Fish Biology 71 (Supplement D), 160–172. 

Verreycken H., Vandenbergh K., Vilizzi L. & Copp G.H., (in prep). Initial application of the 

Fish Invasiveness Screening Kit (FISK) to freshwater fishes in Flanders. 

75

Non-indigenous species of the Belgian part of the North Sea and 

adjacent estuaries 

VLIZ Alien Species Consortium (listed in alphabetical order) 

Bauwens M. 1 , Braeckman U. 2 , Coppejans E. 2 , De Blauwe H. 3 , De Clerck, O. 2 , De 

Maersschalk V. 2 , Degraer S. 2 , Deprez T. 2 , Dumoulin, E. 3 , Fockedey N. 1 , Hernandez F. 1 , 

Hostens K. 4 , Lescrauwaet A.-K. 1 , Mees J. 1 , Rappé K. 2 , Sabbe, K. 2 , Seys J. 1 ,Van 

Ginderdeuren K. 1,4 , Vanaverbeke J. 2 , Vandepitte L. 1 , Vanhoorne B. 1 , Verween A. 2 , 

Wittoeck J. 4 

1 Flanders Marine Institute (VLIZ) 

2 Ghent University; Department of Biology 

3 Belgian Coastal Study Group (SWG) 

4 Flemish Government; Agriculture and Fisheries; Institute for Agricultural and Fisheries 

Research; Fisheries Research Domain (ILVO) 

Contact: Leen Vandepitte & Ann-Katrien Lescrauwaet 

Introduction 

Colonization of marine, coastal and estuarine environments by non-endemic or nonindigenous 

species is not a recent phenomenon in the Belgian context. Some of our history 

books contain references to early human introductions of non-endemic species, related to 

aquaculture and other economic purposes. For those aquatic species, it is relatively easy to 

define the year of introduction or ‘first observation’ for Belgian waters. Non-endemic species 

are also often discovered by coincidence. However, for a number of species in the marine and 

estuarine environment, it is difficult to say whether they are endemic or not. For some taxa it 

is difficult to distinguish between local and non-endemic species, which may lead to 

erroneous determinations. Our knowledge and research techniques have evolved enormously 

over the last centuries, and we can not always say with certainty whether a species may have 

been present before. This is particularly true for smaller and inconspicuous species such as, 

e.g., bacteria, microscopic algae or species with an elusive behavior. We refer to those species 

as ‘cryptogenic’. 

Once non-endemic or ‘alien’ species have settled as reproductive populations, it is 

difficult to turn back the clock. Aliens may cause impact of differing type and degree in their 

new environments. This impact can cause damage to economy, public health and biodiversity. 

In this case, the alien invaders are catalogued as ‘invasive’ species. Directed management 

efforts may mitigate or reduce the damage caused by non-endemic species, or even anticipate 

them. It is therefore important to collect data on these particular species, their location and 

changes in their distribution, and the type of impact they may cause. These factual data - 

together with knowledge on the species’ ecology - provides relevant information to policy and 

management. 

Flanders Marine Institute and its consortium of experts on non-indigenous species 

(‘VLIZ Alien Species Consortium’) conduct an ongoing effort to collect and maintain a list of 

aliens with documented established populations in the Belgian part of the North Sea and its 

adjacent estuaries. 

77

78 

The non-indigenous species list 

The discovery of America (1492) is set as the historical baseline for this assessment. This year 

marked the beginning of a strong increase in trans-Atlantic shipping. Associated with 

shipping, an increase in commerce between continents, their cultures and species evolved. 

Shipping and commerce are two important vectors for the ‘transport’ of non-endemic species. 

Species that arrived after 1492 in our waters are classified as non-endemic or alien. Those for 

which information indicates that resident populations existed before this reference year are 

considered as endemic. 

The list strives to include all currently known non-indigenous and cryptogenic species 

registered in salty and brackish environments in the Belgian part of the North Sea, the Belgian 

coastal zone and adjacent estuaries of the rivers Yser and Scheldt (see figure 1). The Ostend 

Sluice dock, an artificial water body connected to the port docks of Ostend, is also part of the 

study area. The initiative of the VLIZ Alien Species Consortium scrutinizes intentional and 

unintentional introductions by man or other vectors. Alien species for which there is no 

evidence of resident populations are not included in the list, nor are species that are limited to 

the fresh water environment. Vagrant species or occasional observations therefore are not part 

of the scope. New arrivals as a consequence of naturally induced migrations are also 

excluded. To this purpose the possible effects of global warming are not accounted for in the 

list. 

Figure 3: map of the study area 

The trend in the number of observed non-endemic species in the Belgian part of the 

North Sea and adjacent estuaries is on the rise (figure 2). The steep increase from the 1980’s 

onwards is largely due to an increased and improved observation effort. Therefore data before 

1980 can not be compared to data after 1980. 

Currently (May 2009), the list contains 64 resident non-endemic species. Figure 3 

shows the distribution of these resident non-endemic species by larger groups. The arthropods 

count the highest number of non-endemic species (23). In this group, the share of Cirripedia 

(barnacles) is significant (5). Molluscs (i.a. bivalves and gastropods) and algae contribute 

with respectively 11 and 10 species.

70 

60 

50 

40 

30 

20 

10 

0 

cumulative # of first observations 

# first observations 

80 

environment and potential mitigation measures. Pictures, schemes and diagrams add 

qualitative information. The information sheets are the result of a thorough literature study 

and the direct contribution of expert knowledge. They provide scientifically robust and 

reliable information. All sources (publications and documents) are included in a fully 

documented reference list - all digitally available at the VLIZ library. Contact information of 

the experts of the consortium is also available from the web pages. 

Web: http://www.vliz.be/NL/Cijfers_Beleid/Niet_inheemse (Dutch)

Brussels Psittacidae: impacts, risk assessment and action range 

Anne WEISERBS 

Département Etudes, Aves-Natagora (www.aves.be/coa)? Rue du Wisconsin 3, 5000, Namur, 

Belgium. E-mail: anne.weiserbs@aves.be 

Introduction 

Invasive species raise many questions regarding their environmental impacts. Literature 

provides ample advice to help field management choices, sometimes contradicting each other. 

For example, measures should be applied at low levels of abundance; but at the same time, it 

is advised to adapt action to impacts, which are often unknown at the moment of settlement. 

The necessity to define present and potential impacts of introduced species is increasingly 

obvious. In the case of the Psittacidae, there are only a few examples of feral population 

management. This could be linked to the fact that these species scarcely induce major 

economic damages. Furthermore, many people living in town welcome those birds which are 

to a certain extent a substitute to the contact between man and nature. 

The Brussels avifauna has been studied for many years by the birding society Aves, in 

collaboration with Brussels Capital-Region Environment Institute. Information is partly 

obtained through common birds monitoring (by point counts) within the framework of the 

Brussels Environment Survey, but also thanks to research conducted in 2002 to assess the 

impact of the Ring-necked Parakeet Psittacula krameri in Brussels (Weiserbs et al., 2002). 

Finally, a recent study carried out in 2008 reviewed the current status of Psittacidae 

populations in the Brussels Region and analysed their present and potential impacts in order 

to inform policy-makers about the best management practices to limit these impacts. This 

contribution stems from those different researches. 

Brussels Psittacidae 

Three Psittacidae species breed in Brussels: the Alexandrine Parakeet (Psittacula eupatria), 

the Ring-necked Parakeet (Psittacula krameri) and the Monk Parakeet (Myiopsitta 

monachus). The case of the Monk Parakeet will not be further developed here, the Brussels 

feral population being easier to manage (Weiserbs, in press). If Ring-necked Parakeet is 

known to be introduced in at least 35 countries, feral populations of Alexandrine Parakeet are 

much scarcer. In Brussels, both Ring-necked Parakeet and Alexandrine Parakeet are strongly 

increasing, although the second, having settled only recently, is much less numerous. Both 

populations are mixed in the field, sharing, for example, roost and feeding sites. Moreover, 

most of the Alexandrine Parakeet population is located in the North-West of Brussels, where 

the Ring-necked Parakeet is the most abundant. Feeding by man is supposed to have an 

important impact on demography, reducing winter mortality and increasing breeding success. 

Present and potential impacts of the Ring-necked Parakeet in Brussels could be 

summarized as followed: 

• Competition with indigenous fauna is at present the main threat of the species. In 

Brussels, a negative impact on Nuthatch has been suggested (Strubbe & Matthysen, 2007) 

and is observed when competition is experimentally forced (Strubbe & Matthysen, 2009). 

81

82 

• Point count survey between 1992 and 2008 indicates a favourable status of cavity nesting 

birds in Brussels (Weiserbs, 2008). Moreover, if no effect is observed for seven cavity 

nesting species, the Ring-necked Parakeet density has a significant positive effect on the 

trends of four other hole nesters: Green Woodpecker, Blue Tit, Great Tit (less significant) 

and Short-toed Tree creeper. This could be explained by the advanced age of most tree 

settlements in Brussels parks and the excavating behaviour of Ring-necked Parakeet, 

using any starting wounds on the trees to create new cavities. Besides, research conducted 

in 2002 showed extremely high cavity densities in parks inhabited by dense populations of 

Ring-necked Parakeet (Weiserbs et al., 2002). Nevertheless, a negative impact is feared in 

the short-term, linked to the regeneration of tree settlements and resultant shortage of 

cavity supplies. Moreover, impact on other groups, like bats, is unknown, but could be 

real. 

• On the fringe of the previous main threats, a local impact is possible on some fruit crops 

(as observed in Great Britain). 

• Finally, very localised impacts are linked to noise disturbance and dirt under the roosts. 

Present impacts of the Alexandrine Parakeet in Brussels are weak as the population is not 

very large, but are adding to those of Ring-necked Parakeet to which the Alexandrine is 

associated in the field. Moreover, a strong increase has to be expected in the future, which 

may result in growing impacts. 

The risk assessment is based on two schemes. The “UK non-native organism risk 

assessment scheme” (Anonymous, 2005), concerning risks for environment and socioeconomy, 

leads, for both species, to the conclusion of a weak to moderate impact. The 

“Guidelines for environmental impact assessment and list classification of non-native 

organisms in Belgium” (Branquart, 2007), assessing risks for Belgian biodiversity, leads to 

classify both species between categories B (Watch list) and C (low environmental risk). 

Since the two species are closely mixed (co-occur?) in the field, measures will have to 

consider both species in concert. The actions range reviews the possible management 

measures, from the weakest to the strongest: 

• Reduce and modify feeding by man to try to slow down demography. 

• Act at the cavity supply level to lower potential competition with native cavity-nester 

(nest boxes setting, old trees preservation,…). 

• Sterilization using a chemical substance (as Diazacon) could be possible; thisrequires 

catching the birds at roosts, for example by cannon netting 

• Eradication is difficult to plan in an urban context and discouraged, as the current impacts 

are assessed as low and as the public reaction could prevent future action against more 

problematic species. 

Acknowledgments 

We would like to thanks the Brussels Capital-Region Environment Institute who financed theses 

researches and whose collaboration is fruitful. We also are warmly grateful to all the volunteers 

contributing to the Brussels point counts monitoring.

References 

Anonymous 2005. UK non-native organism risk assessment scheme – Version 3.3 

(28.2.2005). Prepared by CABI Bioscience (CABI), Centre for Environment, Fisheries 

and Aquaculture Science (CEFAS), Centre for Ecology and Hydrology (CEH), Central 

Science Laboratory (CSL), Imperial College London (IC) and the University of 

Greenwich (UoG) under Defra 

Branquart E., 2007 (Ed). Guidelines for environmental impact assessment and list 

classification of non-native organisms in Belgium. ISEIA - http://ias.biodiversity.be 

Strubbe D. & Matthysen E., 2007. Invasive ring-necked parakeets Psittacula krameri in 

Belgium: habitat selection and impact on native birds. Ecography 30: 578-588. 

Strubbe D. & Matthysen E., 2009. Experimental evidence for nest-site competition between 

invasive ring-necked parakeets (Psittacula krameri) and native nuthatches (Sitta 

europaea). Biological Conservation (in press) 

Weiserbs A., 2008. Surveillance de l’état de l’environnement bruxellois. Groupe de travail 

Aves – Rapport pour Bruxelles Environnement – IBGE 2008. 

Weiserbs A., 2009. Espèces invasives : le cas des Psittacidés en Belgique. Incidences, 

évaluation des risques et éventail de mesures. Aves (in press). 

Weiserbs A., Jacob J. P. & Rotsaert G., 2002. Evaluation de l’incidence du développement 

des populations de perruches sur les habitats et les espèces indigènes en Région 

bruxelloise. Aves - Rapport pour l’Institut Bruxellois pour la Gestion de l’Environnement. 

83

Research on biological invasions: a Belgian perspective 

Etienne BRANQUART, Barbara GONZALEZ, Dimitri BROSENS & Hendrik SEGERS 

Belgian Biodiversity Platform, Brussels 

Introduction 

Three years after the SOS invasion milestone meeting, the Science facing Aliens conference 

was a new opportunity to produce an overview on Belgian research dedicated to biological 

invasions. In preparation to the conference, we inventoried the Belgian research linked to the 

conference theme and assessed its performance relative to European research, the results of 

which we present here. 

Methodology 

Inventory of the Belgian research 

Research projects dealing with biological invasions were surveyed through the BioBel 

database (Belgian Biodiversity Platform, http://biobel.biodiversity.be, accessed April 30, 

2009). Projects extracted from the database were sorted in two different categories. The first 

includes projects that focus on biological invasions and involve at least 1 full- time scientist. 

The second includes projects that deal with invasion ecology incidentally. 

Several attributes were assigned to the projects. This includes starting and ending date 

of the project, taxonomic affiliation, habitat type, research topic, and funding source. Five 

main research topics were considered based on the session themes of the last Neobiota 

conference (Prague, September 2008) : invasion and dispersion patterns, mechanisms and 

evolution of invasions, impacts of invasions, prediction and risk assessment and management 

practices. These topics are supposed to encompass the full spectrum of research activities 

linked to biological invasions. 

For all the analyses presented below, individual projects were weighted based on the 

number of research teams involved with a least 1 full-time scientist. This implies that more 

weight was attributed to large networks than to individual PhD theses. 

Although we made a great effort to include all invasion-related research projects in BioBel, 

some may have escaped our attention. The following results need to be interpreted with this 

caveat in mind. 

Performance of Belgian research compared to European research 

Bibliometric analyses were performed to compare Belgian with European research dedicated 

to invasion ecology. We used the terms “biological invasion*” or “invasive species” or “alien 

species” or “non-indigenous species” or “non-native species” or “exotic species” or 

“invader*” to search papers published between 1990 and 2008 included in Web of Science ® 

(WoS) . Then, we extracted those papers to which at least one European author contributed. 

This yielded 2796 papers, among which 88 were produced by Belgian authors, alone or in 

collaboration with international authors. 

Two different performance indicators were calculated using this information. Our first 

indicator intends to reflect the research attention dedicated to invasion-related issues in a 

85

86 

country, by calculating the number of publications divided by population size in each country; 

residuals of the linear regression of the number of publications on country population size 

were used to this purpose. The second indicator reflects the impact of the publications, and is 

based on the “times cited” count, or the number of times a published paper is cited by other 

papers indexed in WoS (figure 5). 

The Belgian research on biological invasions 

We identified 56 research projects dedicated to biological invasions being conducted by 

Belgian scientists, from 1990 to 2009. As shown in figure 1, a rising interest in invasion 

ecology is manifest in the exponential growth of research projects related to invasive species 

since 1999. 

In addition to these 56 projects, an additional 22 projects involve invasive species in a 

more incidental way. They are related either to biodiversity monitoring activities or to pest 

control studies. Such projects are not considered in further analyses. 

40 

35 

30 

25 

20 

15 

10 

5 

0 

1990 

Number of research projects 

1992 

1994 

1996 

1998 

2000 

2002 

2004 

2006 

2008 

Figure 1: Number of ongoing Belgian 

research projects dedicated to biological 

invasions since 1990. 

A majority of research projects (62 %) focus on invasive plants; vertebrates are considered in 

23 % of the projects and invertebrates in only 15 % (see figure 2). There is no project in our 

BioBel database that deals explicitly with invasive micro-organisms, fungi or algae. 

Although non-native species are known to invade most ecosystem types in Belgium, 

research dedicated to biological invasions is mainly conducted on terrestrial ecosystems (67 

%). Twenty-nine percent of projects target freshwater systems whereas only 4 % deal with 

marine areas. 

Invetebrates 

15% 

Vertebrates 

23% 

Vascular 

plants 

62% 

Freshwater 

29% 

Marine 

4% 

Figure 2: Share of Belgian research projects dedicated to biological invasions between main 

taxonomic groups (left) and major ecosystems (right) 

Terrestrial 

67%

Frequency 

70 

60 

50 

40 

30 

20 

10 

0 

1990 

1992 

Assessment 

Impacts 

Mechanisms 

Management 

Patterns 

1994 

1996 

1998 

2000 

2002 

2004 

2006 

2008 

Frequency 

40 

35 

30 

25 

20 

15 

10 

5 

0 

1990 

Other 

BELSPO 

PhD grants 

Institutes 

Regions 

Figure 3 – Evolution of ongoing project frequency from 1990 to 2009, by research topics (left) and 

research funding sources (right). 

Two out of the five main research avenues (invasion patterns and management) are 

investigated since the early nineties and constitute the baseline of Belgian invasion ecology 

research. The three other research topics were developed progressively later on. Prediction 

and risk assessment studies are the most recent and least developed subjects (see figure 3). 

Different funding sources supported the Belgian research effort on biological 

invasions (1990-2009) to different degrees 25% of projects received funding from BelSPO, 

25% concern PhD grants, 23% are funded by regional administrations, 18 % rely on core 

institute budgets, and the remaining 9% draw from from various other sources (see figure 3). 

International funding of the Belgian invasion ecology research was nearly absent during that 

period. 

Structural funding was available from regional administrations in charge of 

environment management and from biological research institutes (e.g. National Botanic 

Garden of Belgium, Research Institute for Nature and Forests, Royal Belgian Institute for 

Natural Sciences), allowing the development of long-term monitoring and research 

programmes. On top of that, more focused initiatives were developed from 1999 onwards 

based on the work of research teams involved in BelSPO projects and of PhD students. Those 

studies often focused on invasion mechanisms and impact of biological invasions; they often 

produced innovative results to be published in international journals (see further). The 

implementation of the BelSPO “Science for a Sustainable Development” programme, which 

included invasion ecology as a priority topic, resulted in a significant increase in research 

efforts and acted as a strong catalyst for the development of invasion ecology science in 

Belgium. 

Belgian research in a European context 

Fifty years ago, the publication of Charles Elton’s book The Ecology of invasions by animals 

and plants (1958) launched the systematic study of biological invasions. However, biological 

invasion- related scientific papers only started to be readily produced some 30 years later, 

mostly emanating from the SCOPE programme on biological invasions. This programme 

1992 

1994 

1996 

1998 

2000 

2002 

2004 

2006 

87 

2008

88 

raised awareness on the importance of the phenomenon at a world-wide scale (Richardson & 

Pysek 2008). 

In Europe, the number of publications dealing explicitly with invasion ecology is 

increasing exponentially since the early nineties (Figure 4). The top 10 international journals 

publishing papers on biological invasions is as follows (number of papers published between 

brackets): Biological Invasions (107), Diversity and distribution (90), Hydrobiologia (72), 

Journal of Applied Ecology (63), Molecular Ecology (59), Biological Conservation (58), 

Biodiversity and Conservation (41), Marine Ecology (38), Oecologia (38) and Journal of 

Biogeography (34). 

Number of publications 

1000 

100 

10 

1 

1990 

Europe 


1992 

1994 

1996 

1998 

2000 

Years 

2002 

2004 

2006 

2008 

Figure 4 - Growth in the 

number of publications 

registered on Web of 

Science, by European 

and Belgian scientists. 

Publications by Belgian scientists on biological invasions started to appear with a time lag of 

6 to 9 years compared to the overall production in Europe (figure 4). With a total of 88 

international publications, Belgium rates 7 th amongst European countries, after correcting the 

number of publication for population size of the country (figure 5A). The impact of 

publications by Belgian scientists, as measured by the average number of times they are cited, 

appears to be relatively low when compared to those by their European colleagues (figure 

5B). 

On top of international publications, Belgian scientists produce significant amounts of 

valuable “grey” literature. Notwithstanding the sometimes unjustified poor perception of 

such publications, they are particularly useful for risk assessment and management purposes. 

An example is the “Catalogue of the Neophytes in Belgium” (1800-2005) (Verloove 2006) or 

the various reports produced by scientists of the Research Institute for Nature and Forest. 

Most of these publications are referred to in the Harmonia information system of the Belgian 

Forum on Invasive Species and in the data centre of the Flanders Marine Institute. 

Twenty percent of Belgian papers published in WoS-indexed journals result from 

BelSPO funded projects. This concerns mainly research teams that are member of the 

INPLANBEL-PERINBEL-ALIEN IMPACT suite of projects. The following laboratories also 

contributed significantly to the production of Belgian publications related to invasion 

ecology: Animal Ecology (UA), Behavioural and Evolutionary Ecology (ULB), Biological 

Control and Spatial Ecology (ULB), Forestry (UGent) and Forest, Nature and Landscape 

(KUL). 

Despite the high level of the scientific research dedicated to invasion ecology in 

Belgium, only few research teams have been able to exploit European funding sources. The 

only institutions involved in European Research and Technology Development (RTD)

Residuals (N x HAB) 

Times Cited count 

300 

250 

200 

150 

100 

50 

0 

-50 

-100 

-150 

-200 

30 

25 

20 

15 

10 

5 

0 

UK 

UK 

FR 

FI 

IE 

SE 

NL 

RO 

CZ 

SL 

SE 

LI 

BE 

CZ 

FI 

DK 

ES 

NL 

PT 

FR 

ET 

ET 

DK 

LU 

Figure 5: Country ranking based on the two performance indicators: number of WoS publications 

corrected for population size (A) and Times Cited count of WoS publications (B). 

CY 

Countries 

DE 

ES 

IE 

Countries 

LI 

GR 

SL 

BE 

AT 

CY 

SK 

BG 

GR 

IT 

HU 

HU 

BG 

AT 

RO 

PL 

IT 

SK 

DE 

PT 

PL 

LU 

A 

B 

89

90 

Table 1 – Presentation of the different European RTD projects related to biological invasions and 

partnership with Belgian research teams. 

Program Acronym Title of the project Starting 

year 

FP5 GIANT ALIEN Giant Hogweed (Heracleum mantegazzianum) a 

pernicious invasive weed: Developing a sustainable 

strategy for alien invasive plant management in 

Europe 

FP6 ALARM Assessing LArge-scale environmental Risks with 

tested Methods 

Belgian 

partners 

2002 (NBGB) 

2004 (KUL, 

UCL) 

FP6 MarBEF Marine Biodiversity and Ecosystem Functioning 2005 VLIZ 

FP6 DAISIE Delivering Alien Invasive Species Inventories for 

Europe 

FP6 IMPASSE Environmental impacts of alien species in 

aquaculture 

2005 - 

2006 - 

FP6 REBECA Registration of Biological Control Agents 2006 - 

FP6 ALTER-Net A Long-Term Biodiversity, Ecosystem and 

Awareness Research Network 

FP6 FORTHREATS European Network on emerging diseases and threats 

through invasive alien species in forest ecosystems 

2006 INBO 

2007 ULB 

FP7 PRATIQUE Enhancements of Pest Risk Analysis Techniques 2008 - 

projects are VLIZ (MarBEF), INBO (ALTER-Net) and ULB (FORTHREATS). No Belgian 

partner was involved in the large EU projects that focused on biological invasions (ALARM, 

DAISIE, IMPASSE, etc.). The poor integration of Belgian scientists in EU research network 

is probably due to the publication time lag. On the other hand, many Belgian scientists are 

today involved in EU policy-oriented networks, linked to, amongst others, the Bern 

Convention, the European Environmental Agency (EEA), the European Food Safety Agency 

(EFSA) and the European Plant Protection Organisation (EPPO). The Harmonia information 

system and the ISEIA assessment protocol elaborated by the Belgian Forum on Invasive 

Species were welcomed by those initiatives (see e.g. EPPO 2008, Copp et al. 2009, Essl et al. 

submitted). But the contribution of Belgian scientists to the European alien species inventory 

has been on an exclusively voluntary basis, which is unsustainable in the long term. 

Perspectives for research 

Today, Belgian research dedicated to biological invasions enters a phase of maturity and is 

conducted according to high quality standards. Some research topics are well developed by 

Belgian teams and can be considered as very competitive within the international arena, e.g., 

studies dedicated to the evolutionary and ecological mechanisms of plant invasions or to those 

focusing on the spatial dynamics of invasions. Research effort should capitalise on that basis 

and try to integrate as much as possible within international networks. Scientists should also 

try to address the remaining gaps and consider invasion issues in less-studied ecosystems, 

such as freshwater and marine environments.

The Science Facing Aliens conference generated lively discussions on the future 

priorities for research on invasive alien species. If society wishes to limit the impact of 

biological invasions, then setting up an Early Detection and Rapid Response (EDRR) system 

in Belgium will be needed. This implies developing adequate monitoring activities, correctly 

identifying new invasive species, performing rapid risk analyses and implementing adequate 

management responses. Scientific research activities in support of EDRR should be promoted 

within emergent disciplines like bio-informatics, DNA barcoding and risk analysis. Best 

practices for the management of invasive species should also be identified. 

Furthermore, the invasion history of new invasive alien species will need to be 

documented , especially in the case of non-native species that have so far not or hardly been 

recognised as posing a threat in other countries. We need to quantify the impact of such 

species on biodiversity and ecosystem functioning, using both observational and experimental 

studies. To be effective, the results of these studies should be actively communicated to field 

managers and experts from Belgium and other countries. We therefore invite scientists 

involved in research on invasive species to participate to the risk assessment activities coordinated 

by the Belgian Forum on Invasive Species, and to attend meetings organised by 

EEA, EFSA, EPPO and other international initiatives. 

Finally, we advocate that invasion ecologists liaise with colleagues from other 

disciplines to reinforce interdisciplinary and integrative studies. Key areas where improved 

links with invasion ecology are needed are global change biology, restoration ecology, weed 

science as well as plant, animal and human health science, as exemplified by BelSPO 

MODIRISK and EPI-STIS projects. 


The authors would like to acknowledge all the Belgian scientists who were so kind as to provide 

information about their research projects in the BioBel database, without which this analysis would 

have been impossible. 

References 

Copp G.H. et al., 2009. Calibration of FISK, an Invasiveness Screening Tool for Non-native 

Freshwater Fishes. Risk analysis 29: 457-467. 

EPPO, 2008. Impact assessment of invasive alien plants in Belgium. EPPO Reporting service 

2008/087. 

Essl F., Klingenstein F., Milasowszky N., Nehring S., Otto C. & Rabitsch W., submitted. The 

German-Austrian black list information system (GABLIS) : a tool for assessing 

biodiversity risks of invasive alien species in Europe. Journal of Nature Conservation. 

Richardson D.M. & Pysek P., 2008. Fifty years of invasion ecology : the legacy of Charles 

Elton. Diversity and Distribution 14 : 161-168. 

Verloove F., 2006. Catalogue of the Neophytes in Belgium (1800-2005). Scripta Botanica 

Belgica 39, 89 pp. 

91


Proceedings of a scientific meeting on Invasive Alien Species 

Brussels, May 11 th 2009 

Edited by H. Segers & E. Branquart

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