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Build it and who will come?

Alien species risk assessment in river catchment management

Matthews J.M. (2017). Build it and who will come? Alien species risk assessment in river catchment management. PhD thesis, Radboud University, Nijmegen, The Netherlands.

© 2017 Jonathan Matthews, all rights reserved

ISBN: 978-94-6233-721-3 Cover photo: The spread of the plague by haystackengineering (creative commons license) Print: Gildeprint, Enschede

Build it and who will come?

Alien species risk assessment in river catchment management

Proefschrift ter verkrijging van de graad van doctor aan de Radboud Universiteit Nijmegen op gezag van de rector magnificus prof. dr. J.H.J.M. van Krieken, volgens besluit van het college van decanen in het openbaar te verdedigen op donderdag 5 oktober 2017 om 12:30 uur precies

door

Jonathan Matthews geboren op 23 juni 1973 te Epsom, Engeland

Promotoren: Prof. dr. R.S.E.W. Leuven Prof. dr. ir. A.J. Hendriks

Copromotor: Dr. G. van der Velde

Manuscriptcommissie: Prof. dr. M.A.J. Huijbregts (voorzitter) Prof. dr. L. Posthuma Dr. ir. J.L.C.H. van Valkenburg

Paranimfen: Dr. A. Benítez López Dr. R. Oldenkamp

Contents

1 General introduction ...... 1 1.1 Background ...... 2 1.2 Aims and scope ...... 8 1.3 Thesis outline ...... 10 2 Lessons from practice: assessing early progress and success in river rehabilitation ... 13 2.1 Abstract ...... 14 2.2 Introduction ...... 14 2.3 Methods ...... 16 2.4 Results ...... 19 2.5 Discussion ...... 24 3 A new approach to horizon-scanning: identifying potentially invasive alien species and their introduction pathways ...... 31 3.1 Abstract ...... 32 3.2 Introduction ...... 32 3.3 Materials and methods ...... 35 3.4 Results ...... 42 3.5 Discussion ...... 46 4 Inconsistencies in the risk classification of alien species and implications for risk assessment in the European Union ...... 53 4.1 Abstract ...... 54 4.2 Introduction ...... 54 4.3 Methods ...... 56 4.4 Results ...... 61 4.5 Discussion ...... 65 5 Rapid range expansion of the invasive quagga mussel in relation to zebra mussel presence in the Netherlands and Western Europe ...... 73 5.1 Abstract ...... 74 5.2 Introduction ...... 74 5.3 Methods ...... 75 5.4 Results ...... 79 5.5 Discussion ...... 90 6 A dominance shift from the zebra mussel to the invasive quagga mussel may alter the trophic transfer of metals ...... 97 6.1 Abstract ...... 98 6.2 Introduction ...... 98 6.3 Methods ...... 99 6.4 Results ...... 102 6.5 Discussion ...... 105 7 Synthesis ...... 111 7.1 Introduction ...... 112 7.2 Invasive species risk assessment in river rehabilitation planning...... 112 7.3 Implications of invasive species assessment to rehabilitation planning ...... 113 7.4 Assessing the potential for alien species establishment as a result of rehabilitation interventions ...... 117 7.5 Conclusions ...... 126 7.6 Recommendations ...... 127 Appendices ...... 129 Literature cited ...... 139 Summary ...... 175 Samenvatting ...... 181 Acknowledgements...... 187 Curriculum vitae ...... 193 List of publications ...... 197

Chapter 1

1 General introduction

General introduction

1.1 Background

1.1.1 Alien species and their impacts

Invasive alien species (IAS) are seen as an important cause of ecosystem changes, leading to major ecological and socio-economic impacts. In general, IAS are the second leading cause of global biodiversity loss, after loss of habitat (Moyle et al. 1986; Vitousek et al. 1997; García-Berthou et al. 2005). It has been estimated that IAS have contributed to nearly 40% of all animal extinctions since the 17th century where a cause has been established (Secretariat of the Convention on Biological Diversity 2006). The Secretariat of the Convention on Biological Diversity (2010) identified more than 542 alien species in a sample of 57 countries, including marine and freshwater fish, vascular plants, birds, mammals and amphibians, with a demonstrated impact on biodiversity. An average of over 50 such species were identified per country, ranging from nine to over 220. These figures are almost certainly underestimates as they exclude alien species whose impacts have not yet been assessed, neglect certain species groups such as bacteria, commensals, symbionts and parasites, and only include countries where data on alien species is available (Secretariat of the Convention on Biological Diversity 2010). In the European Union and other European countries, there are approximately 12,000 known alien species in the environment and around 10 to 15% of these are estimated to be invasive (European Commission 2014). Ecological impacts relating to species invasions occur at the individual, population, species, community and ecosystem level and occur through a variety of mechanisms including facilitation, competition, herbivory, predation, parasitism, vectoring of pathogens and physical or chemical habitat modifications (Van der Velde et al. 2006b). Human health can be affected by alien species either directly, for example in the case of poison ivy (Toxicodendron radicans), or indirectly if species act as vectors for human pathogens, e.g., the Asian tiger mosquito (Aedes albopictus) that is known to carry over 20 highly dangerous human pathogens, including dengue fever, yellow fever and chikungunya (Van der Velde et al. 2006b; Matthews et al. 2014a; Sundseth 2014). Alien species establishment may also result in economic costs. For example, the invasive zebra mussel (Dreissena polymorpha) attaches to hard surfaces with byssal threads thereby clogging water intake pipes, water purification plants, and the cooling facilities of electric generating plants (Pimentel et al. 2005). The invasive curly waterweed (Lagarosiphon major) can block the turbine screens of hydro-electric power stations in quantities too great for cleaning machinery to clear, causing temporary shutdowns and power shortages (Chapman et al. 1974). Estimates of the total cost of IAS to the global economy range from hundreds of billions to 1.4 trillion dollars (Pimentel et al. 2001; Secretariat of the Convention on Biological Diversity 2010). IAS are a global problem requiring responses at all administrative levels (Secretariat of the Convention on Biological Diversity 2006). Global efforts to reduce their introduction, establishment and spread are ongoing.

2

Chapter 1

A variety of terms are used as synonyms for ‘alien’ in the literature, e.g., neobiota, non- native, non-indigenous or exotic species (Verbrugge et al. 2016). Moreover, authors differ in their interpretation of the term ‘invasive’. In this thesis, two terms will be applied. The term ‘alien species’ refers to any live specimen of a species, subspecies or lower taxon of animals, plants, fungi or microorganisms introduced outside its natural biogeographical range; it includes any part, gametes, seeds, eggs or propagules of such species, as well as any hybrids, varieties or breeds that might survive and subsequently reproduce (European Commission 2014). This thesis does not include ‘climate shifters’ in its definition of alien species, i.e., species that have expanded their native range as a result of climate change alone. Generally, policy makers tend not to include climate shifters in definitions of alien species. The term ‘invasive’ is used to describe an alien species whose introduction or spread has been found to threaten or adversely impact biodiversity and related ecosystem services (European Commission 2014). The term ecological rehabilitation is used rather than ecological restoration as restoration aims to return an ecosystem to its original state, which is usually unachievable due to a lack of source populations and socio-economic constraints such as industry and agriculture, recreation, flood safety, changed water quality, or navigation. A lack of source populations may be remedied by introducing species. Ecological rehabilitation implies that an image of human-inclusive nature is assumed, i.e., that human society is integral to the design and goals of projects (Lenders et al. 1998). Ecological rehabilitation refers to a process that aims to increase natural biodiversity taking into account these restrictions. Abiotic conditions are recreated in the hope that the original flora and fauna will return. However, this approach may lead to the development of novel or emerging ecosystems that result when species occur in combinations and relative abundances that have not occurred previously within a given biome (See Hobbs et al. 2006). The chapters presented in this thesis feature research that was undertaken from the European perspective, with specific examples of risk prioritisation and analysis applied to the Netherlands. However, the content will also be relevant to nature managers, risk assessors and policy makers globally.

1.1.2 Alien species in international legislation and guidelines

The identification of effective methods for risk prioritisation and assessment of alien species in order to target preventative, early eradication and control measures on national, international and global scales has become increasingly important in recent years. The Convention on Biological Diversity (CBD) is an agreement signed by 150 government leaders at the 1992 Rio Earth Summit that addresses all aspects of biological diversity: genetic resources, species, and ecosystems (UNEP 2014a). It committed world leaders to achieve a significant reduction in the rate of biodiversity loss by 2010. According to the CBD, the maintenance and recovery of viable populations of species in their natural surroundings is a fundamental requirement for the conservation of biodiversity. The CBD specifically requires that signatories prevent the introduction, eradication or control of potential IAS (United Nations 1992).

3

General introduction

The CBD’s requirement in relation to IAS was further refined by the Secretariat of the Convention on Biological Diversity (2014) by obligating an increasing effort in identifying the main pathways responsible for species invasions, and controlling identified pathways through the development of border controls or quarantine measures, and by making full use of risk analysis (Secretariat of the Convention on Biological Diversity 2014). Further to this, in November 2014, the European Commission (EC) published a new European Union (EU) Regulation on Invasive Alien Species as foreseen under target 5 of the EU’s Biodiversity Strategy to 2020. The regulation provides a framework for the introduction of three types of measure: prevention, early warning and rapid response and management of already established IAS (Sundseth 2014). 2.4

2.2

2

1.8

1.6

1.4

1.2 Mean index of invasive index invasive Mean of species alien

1 1965 1975 1985 1995 2005 Year Europe The Netherlands

Figure 1.1. Trend indicator showing the mean increase in the cumulative number of alien species across 21 European countries and in the Netherlands, relative to 1970 figures (mean index). Adapted from Butchart et al. (2010) and Compendium van de Leefomgeving (2017).

However, most negative indicators of the state of biodiversity have increased, including those relating to IAS, despite IAS policy adoption (Butchart et al. 2010; Tittensor et al. 2014). An analysis of the average trend in numbers of alien species present in 21 European countries, selected for having at least 30 records of species with a known introduction date, showed that, on average, 1.7 times more alien species were present in 2012 than in 1970; in the Netherlands 2.09 times more alien species were present in 2003 than in 1970 (Fig.1.1; Butchart et al. 2010, Compendium van de Leefomgeving 2017). Moreover, the ecological and socio-economic impacts of IAS are increasing (Secretariat of the Convention on Biological Diversity 2014). Red list indexes for the impacts of IAS on birds, mammals and amphibians all show that the extinction risk of these groups has increased over time, specifically as a consequence of IAS (McGeoch et al. 2010). An extrapolation of current trends to 2020 indicates that the pressure posed by IAS introductions on global biodiversity is likely to increase (Tittensor et al. 2014), suggesting that additional efforts are required to prevent the impacts of IAS now and in future.

4

Chapter 1

1.1.3 River rehabilitation and alien species invasion

Biodiversity in riverine landscapes is dependent primarily on unhampered hydrological and morphological river dynamics (Amoros & Petts 1993; Petts & Amoros 1996). In sharp contrast to regulated river systems, natural systems feature hydro-morphological mechanisms that encourage the development of a mosaic of interdependent habitats, each with characteristic hydraulic conditions (Pedroli et al. 2002). According to Pedroli et al. (2002), any attempt to manage or restore rivers in favour of biodiversity should aim to achieve these conditions. Therefore, river rehabilitation often involves re-shaping the riverine landscape with the aim of reflecting natural conditions (Buijse et al. 2005). For example, rehabilitation efforts in the floodplains of the river Waal, a distributary of the river Rhine in the Netherlands, involved the creation of secondary channels and the lowering of the floodplain surface in order to encourage natural levee formation (Nienhuis et al. 2002; Wolfert 2002). Literature examining theories of ecological rehabilitation often refers to the pursuance of ecological integrity, which is defined as the ‘maintenance of all internal and external processes and attributes interacting with the environment in such a way that the biotic community corresponds to the natural state of the type-specific aquatic habitat, according to the principles of self-regulation, resilience and resistance’ (Angermeier & Karr 1994; Jungwirth et al. 2002). Large scale morphological change may reduce ecological integrity by encouraging the establishment of alien species thereby moving ecosystems further from their ‘natural state’.

Several biological invasion theories such as the theories of invasional meltdown, biotic resistance, disturbance, fluctuating resource availability, and propagule pressure (Box 1), suggest that rehabilitation interventions may increase an area’s vulnerability to species invasion and may allow the establishment or spread of a wide range of alien species (Strayer et al. 2005). For example, the removal of species in the process of altering local morphology may reduce species diversity and, therefore, biotic resistance. Disturbance resulting from plant species removal during river rehabilitation interventions may lead to an increase in the availability of resources such as space, light, nutrients and water, creating empty niches that are vulnerable to occupation by opportunistic invasive alien species (IAS). Furthermore, high propagule pressure originating from populations of alien species established beyond rehabilitation project boundaries may further increase the likelihood of IAS establishment following interventions.

The introduction of an alien species can be incorporated into rehabilitation efforts. For example, the invasive quagga and zebra mussels (Dreissena rostriformis bugensis and Dreissena polymorpha) have intentionally been introduced to lakes and urban ponds with the aim of improving water quality (McLaughlan & Aldridge 2013; De Hoop et al. 2015). However, introduction of dreissenid species may lead to abiotic changes that facilitate the establishment of other alien species. For example, brittle waternymph (Najas minor), an alien plant species, established following a fourfold increase in water clarity in Lake Erie

5

General introduction

Box 1: Theories of invasiveness

Invasional meltdown: process by which a group of nonindigenous species facilitate one another’s invasion in various ways, increasing the likelihood of survival and/or of ecological impact, and possibly the magnitude of impact (Simberloff & Von Holle 1999).

Biotic resistance: ecosystems with high native diversity can resist invasions better than those with fewer species (Elton 1958; Lodge 1993; Van der Velde et al. 2006b).

Disturbance: processes that lead to, for example, exposed ground or areas of altered light condition may be more vulnerable to species invasion than undisturbed areas (Moyle & Light 1996; Mack et al. 2000; Van der Velde et al. 2006b).

Fluctuating resource availability: A community becomes more susceptible to invasion whenever there is an increase in the amount of unused resources (Davis et al. 2000; Van der Velde et al. 2006b).

Propagule pressure: the number of individuals introduced and the number of introduction attempts is significantly positively associated with the establishment and spread of IAS (Williamson 1996; Colautti et al. 2006a). after the unintentional introduction of D. rostriformis bugensis and D. polymorpha (Stuckey & Moore 1995; GLANSIS 2017). Ecological rehabilitation measures in river floodplains may not reduce the impact of IAS due to their ability in occupying a wide range of ecological niches, and their superior establishment and competitive ability over native species (Peipoch et al. 2015).

In practice, it is observed that alien species colonize disturbed areas more often than pristine ones (Van der Velde et al. 2006b). Wetlands are vulnerable to invasions because they are frequently exposed to multiple introduction vectors and activities that promote species introduction (e.g., water diversions, shipping and recreation) (Strayer et al. 2005; Van der Velde et al. 2006b). Moreover, wetlands are subject to natural disturbance as a result of fluctuating water levels that leads to species rejuvenation and empty niches, encouraging alien species establishment. If transboundary disturbances that block the establishment of native species remain, e.g., diffuse pollution and eutrophication, increased colonization by IAS may occur as a result of local habitat quality improvements (Van der Velde et al. 2002, 2006a; Bij de Vaate et al. 2006). For example, there has been a marked improvement in water quality in the river Rhine since the implementation of the Rhine Action Programme that was put in place following the Sandoz toxic chemical spillage of 1986, resulting in improvements in species richness (Den Hartog & Van der Velde 1987; Admiraal et al. 1993; Leuven et al. 2009). The Rhine Action Programme was the first

6

Chapter 1

pollution abatement programme for the Rhine to include ecological objectives (Smit & Van Urk 1987; Admiraal et al. 1993). However, concurrent increases in diffuse thermal pollution together with warming associated with climate change, and increased connectivity of river catchments as a result of canal building, have been linked to an increase in the establishment of alien species in the region (Nienhuis et al. 2002; Bernauer & Jansen 2006; Van Riel et al. 2009; Leuven et al. 2011). Ironically, a reduction in certain sources of pollution, such as the discharge of industrial waste, may have removed additional barriers for the dispersal and establishment of alien species (Van der Velde & Bij de Vaate 2008). Bij de Vaate et al. (2006) observed that, despite efforts to improve the lateral and longitudinal connectivity of the Lower Rhine channel and reduce habitat fragmentation, macroinvertebrate communities were dominated by alien species that mainly originated from the Ponto-Caspian region linked to the river Rhine via the canals and rivers of the European network of waterways.

Evidence of facilitation of alien species by rehabilitation intervention in other locations is mixed. Information from outside Europe also suggests that rehabilitation measures can lead to increased alien species richness and abundance. A study examining the removal of willows (Salix sp.) from riparian areas in south-eastern Australia increased the likelihood of spread of the invasive aquatic grass Glyceria maxima (Loo et al. 2009). Moreover, Strayer et al. (2005) found, based on a study in the Hudson River in North America, that rehabilitation efforts in large-river ecosystems lead to heavy invasions by alien species. Van der Velde et al. (2006a) concluded that the rehabilitation of large rivers within current human constraints will lead to a dominance of IAS and limited return of native species. On the other hand, there is evidence to suggest that rehabilitation interventions have minimal impact on alien fish species. Schmutz et al. (2014) examined six rehabilitation projects covering 19 sites in the Austrian Danube River and found that alien fish species slightly increased in number following rehabilitation interventions. However, the relative abundances of fish species alien to that part of the river decreased significantly. This was particularly so for round goby (Neogobius melanostomus) and bighead goby (Neogobius kessleri) due to the removal of artificial riprap favoured by these species (Schmutz et al. 2014). A recent study examining the results of 62 reach-scale rehabilitation projects in 51 stream systems in Germany, Switzerland, and Liechtenstein concluded that the proportion of alien to native fish species remained stable in 96.8% of the projects examined (Thomas et al. 2015). Moreover, an analysis of fish species composition on a network scale at 158 sites in Hungarian streams found that alien species distribution was unrelated to habitat degradation (Erös 2007). However, Korsu et al. (2010) used a modelling approach to determine that stream habitat rehabilitation could contribute to the spread of alien salmonid species.

The above conflicting evidence suggests that local factors occurring beyond or within the spatial boundaries of rehabilitation projects, such as hydrological connectivity between

7

General introduction

rivers and the presence of source populations, may play a significant role in the establishment of alien species following rehabilitation interventions. This implies that nature managers should take a precautionary approach and initially assess the risk of alien species introduction occurring (partially) as a result of rehabilitation measures.

1.1.4 Alien species risk assessments in river rehabilitation planning

The CBD requires that participants introduce appropriate procedures requiring environmental impact assessment of its proposed projects that are likely to have significant adverse effects on biological diversity with a view to avoiding or minimizing such effects (United Nations 1992). Indeed, if species invasion is not foreseen, subsequent eradication efforts are costly. Most expenses generated by IAS in Europe result from eradication measures (Colautti et al. 2006b; Sinden et al. 2011). Therefore, ecological rehabilitation projects should be planned in such a way that they minimise the potential for alien species invasions (Strayer et al. 2005).

Buijse et al. (2005) state that prognostic tools should be developed to predict biotic responses to rehabilitation measures at the appropriate spatial and temporal scales. This suggests that identification and analysis of transboundary problems that occur at the catchment scale, such as potential IAS introduction, are an important consideration. However, current European guidelines on river rehabilitation planning largely ignore the potential for alien species invasion as a result of interventions aimed at improving hydro- morphological quality, and refer only briefly to the management of alien species present prior to rehabilitation interventions (Cowx et al. 2013; Peipoch et al. 2015). Current attitudes towards IAS in river rehabilitation practice in Europe tend to be passive or reactive in nature, focussing on the removal of IAS already present or monitoring their establishment following interventions, whose arrival leads to a reduction in perceived success (Angelopoulos et al. 2015). Generally, evaluation monitoring for alien species occurs infrequently (Woolsey et al. 2007; Cowx et al. 2013). For example, in a study of seventy river enhancement projects in the United States of America, geophysical attributes of rivers took priority when setting success criteria and criteria relating to alien species were entirely absent (O’Donnell & Galat 2008).

1.2 Aims and scope

In view of the potential risks that IAS pose to the ecological integrity of habitats subject to rehabilitation interventions, the perceived lack of risk prioritisation and assessment of potential IAS is remarkable. The aim of this thesis is to provide a framework for the assessment of potential risks that alien species pose to the achievement of river rehabilitation goals. The proposed assessment framework is applied in the assessment of D. rostriformis bugensis, a freshwater bivalve species alien to Western Europe. Subsequently, gaps in knowledge that are revealed during the assessment process are explored resulting in

8

Chapter 1

research focussing on the distribution and dispersal mechanisms of D. rostriformis bugensis and the implications of the establishment of this species on the trophic transfer of metals. The proposed assessment framework is based on the unified framework for biological invasions proposed by Blackburn et al. (2011) which is a synthesis of the earlier frameworks proposed by Richardson et al. (2000) and Williamson (1996). The unified framework consists of a number of stages and barriers (in parentheses) that a species has to overcome before becoming invasive in a region where it is classified as alien. The barriers consist of transport (geography), introduction (captivity or cultivation), establishment (survival and reproduction), and spread (dispersal and environmental) (Fig. 1.2). For the purposes of this thesis, the transport and introduction stages are combined into a single introduction stage that reflects the focus of this research on human facilitated introductions of alien species. The environmental barrier reflects conditions that vary from those occurring at the site of introduction that may reduce the potential for spread to other locations. Terminology used for species occurring at, and management interventions applicable to, each stage of the invasion process and referred to in this thesis are included in the figure. Alien Terminology Casual/introduced Naturalized / established Invasive Stage Introduction Establishment Spread

Barrier

Survival

Dispersal

or or cultivation

Reproduction

Environmental Geography, captivity Geography,

Prevention Containment Mitigation Management Eradication

Figure 1.2: The unified framework for biological invasions. Adapted from Blackburn et al. (2011).

The aim each chapter is defined by the following research questions:

Chapter 2: Which indicator groups are most appropriate for the measurement of short term success in river rehabilitation? Chapter 3: Which alien species may become invasive and are also amenable to early warning and rapid response measures in the Netherlands? Chapter 4: What are the main uncertainties associated with the use of protocols for detailed risk assessment and how can risk assessment methodology be improved?

9

General introduction

Chapter 5: Which mechanisms facilitate the rapid range expansion of D. rostriformis bugensis in the Netherlands and how will this affect the long established zebra mussel (D. polymorpha)? Chapter 6: How will a potential dominance shift from D. polymorpha to D. rostriformis bugensis and the trophic transfer of metals affect predator species?

1.3 Thesis outline

The coherence between various research activities and outcomes of this thesis are visualised in a flow chart (Fig. 1.3). The present introductory chapter provides a motivation for the work done, and describes the aims and scope of the study. Chapter 2 provides an assessment of the ability of indicator groups in the short-term (5 years) to demonstrate progress towards river rehabilitation goals. The study featured a survey of river managers that highlighted the lack of attention devoted to the threat that alien species’ invasions pose to the attainment of rehabilitation goals. Chapter 3 describes a new approach to horizon- scanning that identifies potential IAS and their introduction pathways. The method may be applied on a local, regional (e.g., river catchment) and continental geographical scale, or municipal, provincial, national and international (EU) political scales.

Horizon-scanning is the systematic search for potential IAS and their impacts on biodiversity and opportunities for impact mitigation that are currently poorly recognized, that informs policy and practice (Sutherland 2009, 2010; Roy et al. 2014a). This efficient approach stems from a requirement to prioritise a high number of alien species within a short amount of time, but produces less certain results than in depth risk assessments of individual species. Chapter 4 describes results derived from two approaches to alien species risk assessment applied to 18 aquatic alien species for the Netherlands. An alien species risk assessment is the technical and objective process of evaluating biological or other scientific and economic evidence to identify potential IAS, and determine the level of invasion risk associated with a species or pathway and specifically whether an alien species will become invasive (Genovesi et al. 2010).

Chapters 5 and 6 aim to explore uncertainty reduction in risk assessments by bridging knowledge gaps highlighted during the risk assessment of D. rostriformis bugensis, e.g., the species distribution, pathways of introduction, and potential impacts. Chapter 5 describes the rapid range expansion of D. rostriformis bugensis in relation to zebra mussel (D. polymorpha) presence in the Netherlands and Western Europe. Chapter 6 describes how a dominance shift from D. polymorpha to D. rostriformis bugensis may alter the trophic transfer of metals. The implications of the preceding chapters to river rehabilitation practice are discussed in chapter 7. Finally, chapter 8 draws conclusions and gives recommendations for management and further research. An English and Dutch summary, and appendices with raw data and background information complete this thesis.

10

Chapter 1

Incorporating alien species risk assessment in river catchment management

Lessons from practice: assessing early progress and success in river rehabilitation (2)

A new approach to Horizonscanning: identifying potentially invasive alien species and their introduction pathways (3)

Risk assessment protocols for aquatic plants, molluscs and fish (4)

Knowledge gaps with respect to environmental risks of specific alien species. Rapid range expansion of the invasive quagga mussel in The Netherlands and Western Europe (5) A dominance shift from the zebra mussel to the invasive quagga mussel may alter the trophic transfer of metals (6)

General discussion (7)

Conclusions and recommendations (8)

Figure 1.3: Flow chart visualising the coherence of various research lines (Chapter numbers are indicated in brackets).

11

Chapter 2

2 Lessons from practice: assessing early progress and success in river rehabilitation

Jonathan Matthews (Bart) A.J.G. Reeze Christian K. Feld A. Jan Hendriks

Hydrobiologia 655: 1-14, November 2010.

Assessing early progress in river rehabiltation

2.1 Abstract

This article comprises a literature analysis of 41 river rehabilitation projects to assess the short-term (5 years) ability of indicator groups to demonstrate progress towards river rehabilitation goals. Positive indications were compared to land-use, river size, rehabilitation intervention and time. A questionnaire was developed to investigate river manager’s interpretation of rehabilitation success and to assess their level of adherence to recommendations in the literature with regard to rehabilitation assessment on a conceptual level. A total of 54 responses were received from respondents based in Germany, the Netherlands and the United Kingdom. The results indicate that macroinvertebrate indicators, while widely used in assessing river rehabilitation efforts, exhibited a lower frequency of positive responses than most other indicator types in the short term. Conversely, terrestrial floodplain indicators exhibited the most frequent level of positive response for all ecological type indicators leading to recommendations for further investigations into their use for short-term monitoring. Assessment procedures recommended in literature are largely followed, illustrating the advances that have been made with regard to assessment planning. Indicator responses are influenced by scale factors, for example, land-use and river size, that are often not considered by rehabilitation managers. While an emphasis is placed on ecological, hydrological and morphological indicators in monitoring schemes, the socioeconomic perspective (emphasized in the literature as forming an integral part of the river system) is neglected.

2.2 Introduction

Worldwide, there has been increasing interest in the rehabilitation of freshwater systems in an attempt to mitigate against the effects of long-term degradation. An example of legislative formalization of the rehabilitation philosophy is The European Water Framework Directive (WFD). The WFD sets out goals for the attainment of a good ecological status or potential for all surface waters within the European Community to be achieved by the year 2015 (European Commission 2000). This target suggests that the success of river rehabilitation projects should be judged with respect to the WFD guidelines.

The WFD recommends monitoring to determine the level of degradation that is present in surface waters (European Commission 2000). However, monitoring designed to observe qualities that have developed over decades or even centuries, is not likely to be suitable for the observance of difference in the first few years following a river rehabilitation project (Jähnig 2007). The ultimate goal of good ecological status may, therefore, take decades to achieve. The variable response of different ecosystem elements also poses challenges for rehabilitation monitoring. River rehabilitation effects reveal spatial and temporal differences in the response of organism groups (Jähnig et al. 2009). In the Kissimmee River

14

Chapter 2

project in Florida, it was projected that aquatic plants would recover in 3-8 years, invertebrates in 10-12 years and fish in 12-20 years (Trexler 1995). This, combined with delays caused by temporal phenomenon such as drought and dispersal, do not cause rehabilitation to fail, but instead, may push response times beyond those over which monitoring is typically funded (Bond & Lake 2003). The challenge, therefore, is to identify indicators of progress towards, rather than achievement of the goal of good ecological status. Indicators of progress constitute measurable changes occurring within the first few years following project completion that suggest that a rehabilitated river reach is heading towards the condition of good ecological status.

Indicators of progress are required to demonstrate possibly small changes over a limited duration following rehabilitation interventions. This increases the possibility that scale factors will influence indicator response. Scale factors can include abiotic and biotic ‘filters’ that interact locally, as well as climatic and geological ‘filters’ that interact at larger spatial and temporal scales working in a hierarchical fashion (Poff & Ward 1990; Levin 1992; Palmer et al. 2005). In general, recovery appears to be related to stream size (catchment area) and hence colonization opportunities from upstream refuge areas (Friberg et al. 1998; Hansen et al. 1999). Land-use also influences the path of recovery. In un- rehabilitated streams, land-use has been cited as the main driver behind macroinvertebrate composition (Quinn & Hickey 1990; Young & Collier 2009). Specific to river rehabilitation, land-use stressors may influence the path or block the effects of smaller scale river rehabilitation (Moerke et al. 2004; Entrekin et al. 2009). The scale and type of rehabilitation intervention may also influence indicator response.

In general, where both large- and small-scale constraints occur, effective rehabilitation requires a coordinated attack on both (Palmer et al. 2005). However, rehabilitating all areas of a catchment is unrealistic (Brierley & Fryirs 2009), and the influence of land-use, catchment size and temporal factors all affect the condition of rivers and streams, and should be considered when planning rehabilitation (Ekness & Randhir 2007). While these type of effects are acknowledged in the literature, little attention has been focused on appraising their influence on indicators of short-term progress following river rehabilitation.

Conceptual elements required for an effective assessment of the results of river rehabilitation include the setting of objectives, initial monitoring prior to rehabilitation interventions, a comparison of the results of initial monitoring with the results of monitoring post implementation and the public reporting of results (Dahm et al. 1995; Bernhardt et al. 2007; Jansson et al. 2007; Woolsey et al. 2007). In reality, assessment is often fragmented, assessing only certain outcomes, avoiding a holistic assessment of ecology, economy and society (Gillilan et al. 2005). Additionally, little agreement exists on what constitutes a successful river rehabilitation effort (Palmer et al. 2005).

15

Assessing early progress in river rehabiltation

The focus of this study is to provide an initial impression of indicator groups that would be most appropriate for use as indicators of progress within the time limits set by the majority of monitoring schemes (5 years). In addition, the effects of scale factors on the results of ecological monitoring following river rehabilitation will be explored. A questionnaire is used to provide further insight into the attitudes of rehabilitation managers to the definition of project success and to explore possible reasons for indicator group choice. Questionnaire responses are used to investigate whether the poor implementation of the conceptual elements of assessment reported in the literature still persist in practice.

2.3 Methods

2.3.1 Literature review

Data were collected in a literature review and a questionnaire distributed amongst river managers. The literature review aimed to identify indicator groups that demonstrated progress within an initial five monitoring years and allowed an analysis of indicator response with respect to scale factors.

Peer reviewed scientific articles were favoured over other information formats. Papers were selected using the following criteria: (1) the results from the practical assessment of river rehabilitation projects are reported, (2) ecological rehabilitation is a main aim, (3) articles are ideally not published more than 10 years ago and (4) the results of monitoring within 5 years following rehabilitation intervention completion are reported.

Tables were generated, identifying the articles by author and listing scale factor categories that defined the land-use, river size, type of rehabilitation intervention and monitoring duration of each project. If reported separately, the results of articles that contained multiple rehabilitation sites were split into their individual locations.

All individual positive and negative indicator responses within the monitoring results of each article/report were collected and tabulated. A positive indicator response was defined as an indicator response demonstrating progress towards rehabilitation goals within the first 5 years of monitoring. Changes caused directly by interventions themselves, i.e. those involving no evolution away from the initial state created by the interventions, were not included in the analysis. An example of this could be an increase in sinuosity that is a direct result of digging fixed meanders. Any monitoring results that were recorded after the failure of an intervention were also omitted. Statistically insignificant, positive indicator responses were combined with negative responses where statistical analysis was undertaken. An example of statistically insignificant, positive indicator responses would be one where a positive trend was suggested but the sample was too small to provide a significant result. If no statistical analysis was undertaken, for example, where variables

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were not defined quantitatively, indicators that clearly demonstrated a move towards the objectives of river rehabilitation were assumed to be positive. Negative responses were defined as a cumulative movement away from rehabilitation objectives over the total monitoring period or a complete lack of indicator response.

Once tabulated, positive and negative indicator responses were totalled per indicator group and the number of positive responses expressed as a percent- age of the total number of responses found. Subsequently, positive and negative responses within ecological indicator groups were classified according to project river size, land-use, intervention type and monitoring duration scale factors and the results compared for each using bar graphs.

River size was classified qualitatively into the broad subgroups of small, medium and large. These were broad enough to allow the categorization of rivers described in the literature using diverse and qualitative methods, but also allowed a search for trends in the data. A subjective judgment was made using descriptions of catchment size, stream order, channel width measurements at the rehabilitation location, photographs and information obtained from internet sources. The results, therefore, give only an indication of indicator response relating to river size. Land-use was defined using the following sub- groups: urban, agricultural, forestry, limited anthropogenic input and mixed land-use. Projects that suggested a mixture of land-uses were put in a separate mixed land-use category. If required, supplementary information was collected from internet sources.

Rehabilitation interventions were classified into the subgroups of in-channel habitats (ICH), morphological and lateral connectivity. The ICH category consisted of small-scale interventions designed to improve local habitat conditions within the river channel itself e.g. artificial riffles. Morphological interventions were defined as larger scale changes to channel morphology including interventions such as re-meandering and bank and channel re-profiling. The category of lateral connectivity includes interventions that improve the connection between the river channel and its floodplain such as the removal of dikes. All other projects that could not be defined in these categories were allocated to the category ‘other’. These involved interventions that, for example, were difficult to categorize spatially such as changes in flow regime that may just involve changes in discharge but may also increase the inundation of the floodplain. Positive indicator response was also compared with time to establish if there was an optimum time scale within the first 5 years of monitoring where a majority of indicator groups showed progress.

2.3.2 Questionnaire

A questionnaire was developed to support and expand the scope of the data obtained from the literature study. Manager’s concept of what constitutes a successful rehabilitation project was explored. The concept of success was operationalized using the themes of ecological, learning and stakeholder defined in Palmer et al. (2005). These definitions were

17

Assessing early progress in river rehabiltation

supported by Gillilan et al. (2005) and Jansson et al. (2005). This operationalization method was chosen due to its recognition by a variety of authors and its comprehensive nature. Potential project objectives were derived from themes of success as described in Table 2.1. A coding system was used to facilitate the reporting of results.

Responses were generated from Likert type scales designed to gauge the level of importance that managers attached to each objective in relation to project success (1 being of little importance to 5 being very important). Respondents were split into their separate nationalities and analysed statistically to ensure that responses did not vary between countries.

Table 2.1: Classification and importance attributed to objectives and themes of success

1 Theme Objective Objective No. of Importance code responses Ecological No harm is done to the immediate and E1 51 5 success wider ecosystem during rehabilitation interventions Minimal maintenance to the river E2 52 4 section is required following implementation of measures The design of rehabilitation E3 52 5 interventions is based on a knowledge of ecological mechanisms Ecological improvements E4 53 5 Geo-morphological improvements E5 53 5 A more natural hydrological regime E6 52 5 Learning The project delivers a scientific L1 52 4 success contribution that will benefit future rehabilitation projects There is an increase in management L2 52 4 experience The results of the project assessment L3 51 5 are shared with others Stakeholder Increased opportunities for education S1 51 2 success Improved economic opportunities S2 51 1 Increased opportunities for recreation S3 50 1 Improvement in aesthetic value S4 53 4 1Objectives rated for importance to the success of river rehabilitation, mode average (1 being of little importance to 5 being very import).

Respondent speciality (ecology or non- ecology) and the size of river rehabilitation projects were statistically analysed to discover if these characteristics had a bearing on the application of monitoring methods and interpretation of success. The data were non-normal

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in distribution; therefore, analysis was undertaken using non-parametric statistical tests (Mann–Whitney U test and χ2 test). Finally, the results for individual objectives were combined to identify the relative importance placed on different themes by managers.

The second half of the questionnaire was devoted to discovering if conceptual steps taken to practically assess rehabilitation projects reflect theoretical recommendations. These steps can be summarized as an assessment of the initial state of degradation, development of a reference state, the formulation of objectives and the presence of a monitoring and assessment system (Dahm et al. 1995; Kondolf 1998; Lake 2001; Jansson et al. 2005; Palmer et al. 2005; Stoddard et al. 2006; Woolsey et al. 2007).

The questionnaire was piloted and then translated to German. Lists of potential respondents with experience in the management of river rehabilitation projects were generated for the United Kingdom (n = 100), Germany (n = 100) and the Netherlands (n = 20). Lists of questionnaire respondents were not taken from projects considered during the literature review. Instead, internet sources, reports, a database held by the UK River Restoration Centre and previous distribution lists were used to provide respondent lists. The respondent lists were analysed for representativeness by assessing if any respondents had worked on the same rehabilitation projects. Three out of the 27 UK respondents were found to have taken part in the same project. This project was extensive, however, covering 7 km of river over six sites within the UK River Avon catchment. Additionally, these respondents represented different disciplines suggesting that their responses would originate from different perspectives of practice. No German or Dutch respondents were found to have worked on the same projects. Once the questionnaire was sent, three reminders were delivered by email at 2 week intervals to improve response rates. The data obtained exploring the practical application of the conceptual methodology of river rehabilitation was summarized using content analysis and percentage response.

2.4 Results

2.4.1 Literature review

The results from the monitoring of 41 rehabilitation projects employing a total of 342 individual indicators were distilled from 239 publications found during the literature study. Details of the total number of indicators analysed related to the number of source articles and reports per indicator group are given in Table 2.2.

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Assessing early progress in river rehabiltation

Table 2.2: Overview of indicator sources. Indicator group No. of source articles / reports Total number of indications analysed Hydrological 11 46 Terrestrial fauna and flora 8 28 Morphological 10 25 Habitat structure 19 58 Functional 5 10 Fish 16 47 Macrophytes 8 27 Macroinvertebrates 17 73 Physico-chemical 5 27

Indicator groups that demonstrated progress towards rehabilitation goals within an initial five monitoring years The largest positive response was demonstrated by non-ecological indicators, with the exception of terrestrial species (Fig. 2.1). Macroinvertebrates demonstrated the poorest response of all indicator types with the exception of physico-chemical indicators. In this study, macroinvertebrates were the most frequently utilized indicator of any type for short- term monitoring. Functional indicators demonstrated a relatively frequent positive response; however, this must be viewed with respect to the relatively low sample number. In only one study analysed was an indicator of stakeholder satisfaction applied, no other examples of socioeconomic indicators were found. This group was therefore not included in the bar graph.

100

90 80 70 60 50 40 30 20 10

% indicators demonstrating indicatorsdemonstrating % 0 a move goals towards move rehabilitationa

Figure 2.1: Indicator response per indicator group within the first five years of monitoring following the completion of rehabilitation interventions.

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Overall response of ecological indicators per river size Data comparing river size with the response of ecological indicators (Fig. 2.2) demonstrate that in the first five monitoring years, the larger the river, the more likely ecological indicators will respond.

Overall response of ecological indicators per land-use type The number of positive response varies in a way that reflects the level of stress that the land-use confers on the ecosystem (Fig. 2.2). More natural systems (forestry and systems characterized by limited anthropogenic influence) appear to respond more slowly or have a more limited response to rehabilitation interventions. No respondents to the questionnaire reported a consideration of land-use stressors, specifically, in the formulation of objectives. Considerations to land-use were related to issues of spatial planning and stakeholder requirements in most cases.

Overall response of ecological indicators per rehabilitation intervention type The manipulation of ICH is less likely to produce positive responses in ecological indicators in comparison with other approaches that employ larger scale morphological interventions and reinstatement of lateral connectivity with the floodplain (Fig. 2.2).

100 90 80 70 60 50 40 30 20 10

0

large (n=37) large

small (n=69) small

urban (n=29) urban

mixed (n=27) mixed

forestry (n=21) forestry

medium (n=49) medium

% Indicators demonstrating a move towards project goals project towards move a demonstrating Indicators %

agricultural (n=58) agricultural

(n=33)

In channel habitat (n=65) habitat channel In

lateral connectivity (n=19) connectivity lateral

lateral connectivity +/- other (n=13) other +/- connectivity lateral

morphological, in channel habitat +/- habitat in channel morphological,

morphological +/- lateral connectivity lateral +/- morphological limited anthropogenic influence (n=17) influence anthropogenic limited River size Land-use Rehabilitation intervention

Figure 2.2: Ecological indicator response in relation to scale factors within the first five monitoring years following the completion of rehabilitation interventions.

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Assessing early progress in river rehabiltation

Overall response of ecological indicators per monitoring year The variation in the between year data indicates that there is no obvious optimum monitoring duration within 5 years that would provide the most positive indication of the progression of a rehabilitated river system towards ecological objectives (Fig. 2.3). The relatively large positive response at 5 years must be viewed in respect of the low number of indicators analysed.

100

90 80 70 60 50 40 30 20

10 a move towards rehabilitation goals rehabilitation towards move a % indicator response demonstrating response indicator % 0 1 (n=29) 2 (n=53) 3 (n=27) 4 (n=14) 5 (n=15)

Monitoring year (n = total number of indicators analysed)

Figure 2.3: Ecological indicator response per monitoring year following the completion of rehabilitation intervention.

2.4.2 Questionnaire responses

Importance of themes to project success Responses related to the importance of project objectives and respondent characteristics are described in Tables 2.1 and 2.3. Analysis demonstrated that ecological objectives were seen by rehabilitation managers as most important to a definition of project success. There was no significant difference observed between respondent specialties or German and British respondents (Table 2.4).

Table 2.3: Respondent characteristics. Respondent speciality Nationality No. of Mean project size (length in Non- No respondents Ecological metres) ecological response German 23 5 5 13 5148 British 27 8 8 11 4214 Dutch 4 0 3 1 5150

Ninety-four percent of rehabilitation managers involved stakeholders in the formulation of objectives for river rehabilitation projects (n = 53). However, only a small proportion of

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respondents included an indicator of stakeholder satisfaction in their assessment to gauge the success of their projects (Fig. 2.4).

Table 2.4: Mann–Whitney two-tailed analysis of themes of project success, scope of inclusion of spatial elements and respondent characteristics. Ecological success Learning success Stakeholder Scope of success inclusion of spatial elements Respondent U = 78.5, n1 = 13, U = 75, n1 = 13, n2 = 16, U = 75.5, n1 = 13, U = 101, n1 = 13, speciality n2 = 15, P < 0.05 P < 0.05 n2 = 16, P < 0.05 n2 = 16, P < 0.05

Respondent U = 244.5, n1 = 22, U = 218.5, n1 = 22, U = 226, n1 = 22, U = 148, n1 = 23 nationality n2 = 27, P < 0.05 n2 = 27, P < 0.05 n2 = 27, P < 0.05 n2 = 13, P < 0.05

Objectives that combined to form the stakeholder theme within the questionnaire were rated as being of less importance to the success of rehabilitation projects compared with the learning and ecological themes. There was no significant difference observed between respondent specialties or German and British respondents (Table 2.4).

Respondents rated objectives related to the importance of learning success second after ecological type objectives. Again, there was no significant difference observed between respondent specialty or German and British respondents (Table 2.4). The dissemination of project results was included in the objectives in 77% of rehabilitation projects referred to in questionnaire responses (n = 47).

Macroinvertebrate

Morphological

Fish

Hydrological Indicator group Indicator

Vegetation

Stakeholder satisfaction

Physicochemical

0 20 40 60 80 100

Percentage of respondents

Figure 2.4: Percentage of total respondents who reported incorporating particular indicator groups in their monitoring approach.

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Assessing early progress in river rehabiltation

Comparison of theoretical concepts and practical application The main conceptual elements of river rehabilitation assessment recommended in the literature were demonstrated to be present in projects carried out in practice (Tables 2.5, 2.6). One notable result is the relatively high number of respondents who undertook an assessment of outcomes but neglected to compare monitoring results with project objectives or the initial degradation state. There was no significant difference found between either respondent specialty and project size or the scope of inclusion of spatial elements (Table 2.4). An analysis examining possible relationships between the application of reference state and respondent specialty was invalidated due to insufficient sample size.

Table 2.5: Conceptual elements - Initial degradation assessment: inclusion of spatial elements. River River channel & River channel, riparian No initial channel riparian zone zone & floodplain assessment o N responses analysed 10 16 24 3 % projects where 19 30 45 6 element(s) included

Table 2.6: Conceptual elements – other. Assessment of outcomes by Reference Monitoring comparison with initial degradation state state or project objectives o N responses analysed 42 47 48 % projects where 82 94 73 element included

The macroinvertebrate indicator group was used most frequently for project monitoring, whereas socioeconomic and physicochemical type indicators were used the least (Fig. 2.4). There was no significant difference between respondent specialty and the application of morphological indicators χ2 (1, N = 42) = 0.73, P > 0.05. In other cases, however, analysis was not possible due to insufficient sample size.

2.5 Discussion

2.5.1 Methodological limitations

During the research process, a number of methodological limitations were identified. First, it could often not be ascertained whether a lack of indicator response was due to a non- response of the indicator or the result of an ineffective intervention. An example of this is the introduction of river bed material of the wrong grain size when attempting to encourage fish spawning. Also, the influence of other large-scale limiting variables, such as diffuse pollution, might have resulted in a lack of indicator response. The results are therefore expressed in terms of percentage positive indicator response. The use of percentages also

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introduces negative indicator responses to the analysis. However, the resultant uncertainty is the same for all indicator groups analysed which reduces its influence during group comparisons. Second, it was sometimes difficult to establish whether a development, particularly related to morphology, demonstrated a positive move towards river rehabilitation goals. Often morphological developments were not evaluated in relation to project goals. Only morphological development that was stated to be a progression towards project objectives or the general goals of river rehabilitation was considered a positive indicator response. This interpretation tends to reduce the number of positive indications taken from literature sources. However, the results demonstrate that morphological response was rated highly suggesting that the influence of this uncertainty is limited.

2.5.2 Indicator groups that demonstrated progress within the initial five monitoring years

Terrestrial flora and fauna indicators gave more positive short-term results following rehabilitation compared to other, within river indicators such as fish and macro-invertebrate indicator groups (Fig. 2.1). Potential explanations are differing sensitivities relating to water pollution, colonization potential and the presence of habitats in the locality prior to the implementation of rehabilitation interventions (Jähnig et al. 2009). Individuals of most (riverine) taxa move only short distances over their lifetime, this immediately constrains the probability that restored habitat will be colonized, particularly in the short term (Bond & Lake 2003). Floodplain inhabitants, on the other hand, have previously been noted to be sensitive to river degradation and rehabilitation. Carabid beetles react negatively to channelisation and floodplain connectivity and floodplain plants may benefit from seed banks which are mobilized if sediment is relocated (Kangas 1990; Lott 1996; Hölzel & Otte 2001). In general, riparian organism groups may be used as additional parameters to evaluate the short-term effects of rehabilitation, while aquatic organisms may be better suited to reflect long-term effects (Jähnig et al. 2009). In light of these results, the potential for terrestrial species to demonstrate early progress following river rehabilitation should be further investigated with a view to formalizing their use in short-term monitoring schemes.

Macroinvertebrates performed less well in comparison to the majority of other indicator groups. The poor response of macroinvertebrate indicators may be due to the natural variability that is attributed to this group. Natural variation has been demonstrated to reduce the sensitivity of macroinvertebrate indicators to change (Clarke et al. 2006; Blocksom & Flotemersch 2008). Variability in relation to space, time and sampling methods have been stated as significant factors that result in difficulties in identifying significant change in macro-invertebrate monitoring (Brooks et al. 2002; Sporka et al. 2006; Haase et al. 2008). The results suggest that macroinvertebrate indicators maybe too insensitive to consistently identify early progress following river rehabilitation interventions. Macroinvertebrates were the most frequently used of all indicator groups in projects examined. The application of this indicator group for short-term monitoring should therefore be re-evaluated.

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Assessing early progress in river rehabiltation

Non-ecological indicators of rehabilitation success were shown to respond most quickly to the interventions taken to rehabilitate river stretches. However, this result should be considered with a note of caution. Morphological changes are dependent on receiving a period of climactic normality including extreme events and/or climactic trends during the monitoring period (Downs & Kondolf 2002). Monitoring duration must, therefore, be tailored to the climactic norms of the region under surveillance.

The influence of scale factors on indicator groups Indicators were observed to respond more positively with increasing river size (Fig. 2.2). This has also been observed by others. Recovery appears to be related to stream size (catchment area) and hence colonization opportunities from upstream refuge areas (Friberg et al. 1998; Hansen et al. 1999). The dispersal ability of riverine organisms has been implicated as a possible reason for this. With the exception of most insects, and fish in lowland rivers, natural re-colonization of restored sites is only likely to occur from sites within the same stream (Hughes 2007). The results emphasize the importance of species colonization potential and the presence of local source populations in the choosing of rehabilitation locations.

The influence of land-use on river systems is widely documented. Stream ecosystems are constrained by phenomena acting at different scales in a hierarchical fashion and influences beyond the spatial limits of the project reach will have an effect (Douglas Shields et al. 2007; Schwartz & Herricks 2007; Kail & Hering 2009). The results suggest that rivers with surrounding land-use that is associated with higher degradation may exhibit a greater potential for early signs of recovery (Fig. 2.2). An explanation for this could be a lack of distinction between species reflecting a natural reference state and colonization by species that are tolerant of land-use stressors. Tolerant species are able to take advantage of empty niches within initially highly degraded systems following local improvements in habitat quality (Den Hartog et al. 1992; Van der Velde et al. 2002). The poorer response of rehabilitated stretches subject to lesser degrees of land-use stress may also be symptomatic of ecological resilience (Vugteveen et al. 2006). Ecological resilience is characterized by a buffering effect where ecosystems absorb moderate levels of stress before measurable degradation occurs. Caution should also be applied as changes that bring a system in poor condition to good condition maybe more easily detectable and achievable than changes that bring a system in good condition to high condition.

In general, the influence of stressors beyond the boundaries of the restored area should be taken into account during rehabilitation planning, and rehabilitation should occur at the appropriate spatial scale such that rehabilitation is not reversed by the prevailing disturbance regime (Lake et al. 2007). When asked about factors that were considered in the formulation of objectives for river rehabilitation (a question where land-use was given as an example answer), no questionnaire respondents referred to land-use stress in their

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responses. The influence of spatial factors, such as land-use stress, on monitoring results should therefore be emphasized to the managers of river rehabilitation.

The analysis of response of ecological indicators to rehabilitation intervention type indicated that interventions aimed at improving ICH induced fewer positive indicator responses on their own and in combination with other intervention types (Fig. 2.2). These observations support the notion that larger scale morphological interventions will produce more positive results from monitoring schemes within 5 years. The involvement of the floodplain in rehabilitation interventions may increase the potential for overall system recovery in the short term by exploiting the recovery potential of terrestrial fauna and flora observed by Jähnig et al. (2009). Larger scale morphological interventions such as re- meandering, if carried out correctly, may have a greater influence on natural habitat provision than smaller scale in-stream interventions that may also be more temporary in nature due to structural failure (Johnson et al. 2002).

A high degree of variation was observed when all indicator groups were combined and then viewed in terms of percentage positive responses from year to year (Fig. 2.3). During the initial phase (of rehabilitation), biotic elements and morphology will exhibit large spatial and temporal variation reflecting amongst-year differences in environmental conditions that have little to do with recovery from rehabilitation (Friberg et al. 1998; Muotka et al. 2002; Petranka et al. 2003). This suggests that monitoring should continue for longer than 5 years to gain a consistent view of progress towards rehabilitation goals.

Comparison of theory and practice: the conceptual building blocks of assessment and the interpretation of success The inclusion of objective setting, assessment of the degradation state, the use of a reference state and monitoring in questionnaire respondent projects suggests that river managers have taken note of past criticisms relating to the lack of proper assessment within river rehabilitation projects (Tables 2.5, 2.6). A potential area for improvement is in project assessment where 27% of respondents reported that an assessment was undertaken but monitoring results were not compared with project objectives or an initial degradation state.

Respondents reported that monitoring was under-taken in the majority of rehabilitation projects, contrasting with statements made in the literature. However, lack of standardisation in project reporting and monitoring continues to hamper abilities to compare and analyse the outcomes of similar projects. The importance of standardisation has been emphasized as a determining factor in the effectiveness of river rehabilitation (Gretchen & Allan 2006). Respondents emphasized, however, the importance of knowledge sharing to the success of river rehabilitation suggesting that practitioners would be open to efforts to standardize assessment in the future.

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Assessing early progress in river rehabiltation

An important observation that was made during the study was the lack of formal assessment of the achievement of stakeholder and learning objectives during project assessment. This is reflected in the relative lack of importance placed on the stakeholder theme (Table 2.1) and supports separate observations that criticise the lack of holistic assessments in river rehabilitation (Gillilan et al. 2005). Rivers cannot be considered as systems that are isolated from the surrounding human population, and society should be included in assessments of river condition. River system health has been defined as an expression of a river’s ability to sustain its ecological functioning (vigour and resilience) in accordance with its organization while allowing social and economic needs to be met by society (Vugteveen et al. 2006). The inclusion of socioeconomic elements to provide a holistic view of river systems is supported by others (Leuven et al. 2000; Lenders & Knippenberg 2005). In a more practical sense, socioeconomic elements are seen to permeate the planning and execution of river rehabilitation. Decisions related to river rehabilitation have to take into account trade-offs between ecological goals, ecosystem services, competing land-uses and costs (Reichert et al. 2007). Rehabilitation managers should, therefore, consider the relevance of socioeconomic factors and formulate objectives and indicators to measure success in relation to their projects.

Acknowledgements

Thanks go to Prof. Dr. Daniel Hering, Rico Gerfen, Mirjam de Groot and Kathrin Januschke for their technical assistance. Special thanks go to Martin Janes and Nick Elbourne at the River Restoration Centre in the UK for their assistance in providing lists of respondents for the questionnaire.

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3 A new approach to horizon-scanning: identifying potentially invasive alien species and their introduction pathways

Jonathan Matthews Ruud Beringen Raymond Creemers Hans Hollander Niels van Kessel Hein van Kleef Sander van de Koppel Adrienne J.J. Lemaire Baudewijn Odé Laura N.H. Verbrugge A. Jan Hendriks Aafke M. Schipper Gerard van der Velde Rob S.E.W. Leuven

Management of Biological Invasions 8: 37-52. March 2017.

Identifying potentially invasive alien species and their introduction pathways

3.1 Abstract

Invasive alien species (IAS) are considered an important threat to global biodiversity due to major ecological impacts. In 2014, the European Union (EU) introduced a regulation (EU) No 1143/2014 on the prevention and management of the introduction and spread of IAS. The first risk prioritized list of IAS of EU concern was adopted on the 3rd of August 2016. EU member states are required within 18 months to carry out a comprehensive analysis and prioritisation of the pathways of unintentional introduction and spread of these IAS in their territory. Horizon-scanning is a method of IAS prioritisation through the systematic analysis of potential future IAS and identification of new opportunities for IAS management. However, horizon-scanning has mostly been applied on a national basis only, leading to a fragmented approach within the EU and ignoring the potential for IAS to cross international borders. We present a novel framework for horizon-scanning applicable at a continental scale. Our method maximises the use of available data from climatically matched countries by applying a harmonisation and aggregation method, and elucidates the relationship between pathways, impact types and species groups for risk prioritised IAS. Application of the method produced a list of potential IAS for the Netherlands revealing that the international trade in plants and animals is the most important pathway for the introduction of IAS. The horizon-scanning provided a starting point for the design of preventative, early identification and rapid action measures for the effective management of potential IAS.

3.2 Introduction

Invasive alien species (IAS) are species whose human aided introduction and/or spread outside their natural past or present distribution threatens biological diversity, economy and/or public health (UNEP 2014a). IAS are considered to be one of the leading causes of global biodiversity loss (Moyle et al. 1986; Vitousek et al. 1997; García-Berthou et al. 2005). Examples of IAS that pose a high risk to native species as a result of, for example, competition and predation, include the striped skunk Mephitis mephitis (Schreber, 1776), curly waterweed Lagarosiphon major (Ridl.) Moss, fanwort Cabomba caroliniana A. Gray, and pumpkinseed Lepomis gibbosus (Linnaeus, 1758) (Kay & Hoyle 2001; Van Kleef et al. 2008; Van Belle & Schut 2011; Matthews et al. 2012b 2013a,b). For the purposes of this article, risk is defined as a combination of the consequences of an event (hazard) and the associated likelihood/probability of its occurrence (European Commission 2010).

Most economic costs generated by IAS in Europe result from reactive eradication and control measures (Colautti et al. 2006b; Vilà et al. 2009; Sinden et al. 2011). Moreover, once established, IAS can rapidly expand their range across national borders (Shirley & Kark 2006). Therefore, prevention of initial introduction and spread between member states through coordinated international action is seen as the most cost-efficient measure to

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combat potential impacts (Pyšek & Hulme 2005; Shirley & Kark 2006; Essl et al. 2011; European Commission 2013; UNEP 2014a). Additionally, an effective early warning system that promotes rapid action when species elude preventative measures is required (Leung et al. 2002; García-Berthou et al. 2005; Finnoff et al. 2007; Coetzee et al. 2009; European Commission 2013). For both prevention and rapid action, it is essential that IAS that are likely to invade new territories are identified (Shine et al. 2010). In 2014, the European Union (EU) introduced a regulation (EU) No 1143/2014 on the prevention and management of the introduction and spread of IAS (European Commission 2014). The first risk prioritized list of IAS of EU concern was adopted on the 3rd of August 2016. Currently, 37 species are present on the list (European Commission 2015, 2016). However, the list is open to future alteration that provides opportunities for methodological improvement and the application of new information. Moreover, within 18 months of the adoption of the list, EU member states are required to carry out a comprehensive analysis and prioritisation of the pathways of unintentional introduction and spread of IAS in their territory (European Commission 2014). Pathways are defined as the routes and mechanisms of the introduction and spread of IAS (European Commission 2014).

When applied to alien species, horizon-scanning is the systematic search for potential IAS, their impacts on biodiversity and opportunities for impact mitigation that are currently poorly recognized, that informs policy and practice (Sutherland & Woodroof 2009; Sutherland et al. 2010; Roy et al. 2014a). This approach is an important tool that contributes to the prevention, early identification and eradication of IAS in Europe (Caffrey et al. 2014; Shine et al. 2010). Generally, horizon-scanning has been applied on a regional rather than EU scale, e.g. in projects by the Great Britain Non-Native Species Secretariat (Parrott et al. 2009), the Belgium Forum on Invasive Species (D’hondt et al. 2015), and the RINSE (Reducing the Impacts of Non-native Species in Europe) project that focussed on Great Britain, France, Belgium and the Netherlands (Gallardo et al. 2013). This fragmented approach to risk prioritisation hampers effective international control and, in many cases, has relied on knowledge derived from small expert groups, which may reduce certainty (European Commission 2013; Roy et al. 2014a).

Invasiveness in locations with similar ecological and climatic conditions is considered one of the most relevant criteria in predicting the invasive behaviour of a species (Williamson 1996; EPPO 2012). However, previous horizon-scans often neglected this important driver of species invasion. For example, the RINSE horizon-scanning project that focussed on Belgium, the Netherlands and the United Kingdom (Gallardo et al. 2013) carried out a retrospective bioclimatic match for a proportion of analysed species but this did not influence their final species classifications. The recent horizon-scan and pathway analysis carried out by NOBANIS (2015) does not include a formal climate match. Only the European-scale horizon-scan carried out by Roy et al. (2015) considered the influence of European climate zones on the potential future establishment of IAS in different European

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Identifying potentially invasive alien species and their introduction pathways

regions. Previous horizon-scanning methods have relied on labour-intensive expert consultation and, while often presenting an overview of potential introduction pathways for individual species, many do not include systematic analyses aimed at prioritising introduction pathways (e.g., Parrott et al. 2009; D’hondt et al. 2015). The recent pathway analysis carried out by NOBANIS (2015) is limited to Nordic and Baltic countries and only examines species recorded in the NOBANIS database and Danish, Norwegian, German and Irish alert lists. While Roy et al. (2015) present a limited pathway analysis, they examine pathways at a relatively high level of abstraction, omit statistical analyses, and do not consider pathways in relation to ecological impact types. The importance of these potential shortcomings was emphasised during the 7th European Conference on Biological Invasions in Pontevedra (Spain) by the adoption of a resolution stipulating that, by 2020, IAS and their introduction pathways in Europe should be identified and prioritised, priority species are controlled or eradicated, and that pathways are managed to prevent the introduction, establishment and spread of new IAS (Neobiota 2012).

In this study, we present a novel approach to horizon-scanning that aims to address these limitations by: 1) designing a method that may be applied on any continental scale; 2) combining risk classifications derived from existing sources thereby reducing the need for expert consultation; and 3) determining the frequency and statistical significance of ecological impact types and introduction pathways to facilitate pathway prioritisation. Our method maximises the use of available data on risk classifications from countries climatically matched to the target region by applying a harmonisation and aggregation method that produces a prioritised list of potential IAS. Subsequently, an inventory of origins, pathways and impact types of these species is compiled and analysed. The inventory may facilitate policy decisions relating to prevention, early detection and eradication of the identified potential IAS.

The aims of our study were threefold: 1) To develop a horizon-scanning method that may be applied on a continental scale for identifying potential IAS absent from or with a limited distribution, amenable to prevention or early eradication measures. 2) To identify the most prevalent pathways, geographical origins and impact types associated with these potential IAS that may assist in the design and prioritisation of prevention or early eradication measures. 3) To test this method in a case study, by producing a list of potential IAS for the Netherlands and analysing their most prevalent pathways, geographical origins and impact types.

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3.3 Materials and methods

3.3.1 Horizon-scan framework

The horizon-scan aimed to identify a list of potential IAS most likely to be amenable to effective management interventions in the target region. Risk classifications of alien species obtained from regions climatically matched to the target region were harmonised, aggregated per species, and ranked to produce a risk prioritised list of alien species. The risk prioritised list was then screened using three criteria designed to identify new alien species that are most likely to be amenable to effective management interventions due to a limited distribution in, or absence from, the target region. Subsequently, an inventory of the origins, pathways and potential impacts of the high risk species (list of potential IAS) was undertaken, which serves as a basis for the design of preventative measures, and early detection and rapid response systems. The method consists of seven steps (Fig. 3.1).

Step 1: Risk classifications are collected using the sources listed in Appendix 3 for species that are applicable either to the target region or for any region that is climatically matched to the target region according to the Köppen-Geiger climate classification (Rubel & Kottek 2010). The climate match is made on a national scale, however, if horizon-scans or risk assessments have been carried out for a region that includes a number of different countries, e.g., Western Europe or North America, then these may be included as well. If a risk classification of a particular species already exists for the target region, then this is used instead of risk classifications from other regions. However, information underpinning risk classifications from other regions may still be relevant if it is not included in the risk analysis for the target region, and supports the results contained in the risk analysis for the target region.

Step 2: There are multiple assessment methods currently in use that apply different terms to classify the same level of risk (e.g., high risk, black list and priority species). Therefore, harmonisation of risk categories is required prior to the aggregation of classifications. Harmonisation occurs by attributing a score of 1 (low risk), 2 (medium risk) or 3 (high risk) to the risk classifications applied in the original risk assessments (see example harmonisation in the case study for the Netherlands presented in the results section and Table 3.1). Decisions to allocate a certain risk classification to a harmonised risk category are based on interpretation, e.g., a black listed species suggests a high risk species. The proposed classification system is then verified using expert consultation and consensus during a workshop. In the case study of the Netherlands, five to six experts per taxonomic group were consulted. Expert judgements were then verified by all contributors. In cases where there is discussion over category harmonisation, e.g., if the number of categories in a particular system varies from the three categories contained within the harmonised system, the precautionary principle should be applied (Raffensperger & Tickner 1999) and species should be assigned to the higher harmonised risk category.

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Identifying potentially invasive alien species and their introduction pathways

Horizon scan for potentially invasive alien species (IAS)

Step 1: (Opportunistic) species with ecological risk classifications from climatically similar regions to the Netherlands: 1425 species

Step 2: Standardisation of risk classifications to produce harmonised risk scores through expert consultation

Step 3: Aggregation and ranking of harmonised risk scores to produce risk prioritised species list

High risk, high uncertainty High risk, low alien species; medium and uncertainty alien Step 7: Periodical low risk alien species species: 243 species review of high risk, high (high and low uncertainty species uncertainty): 1182 species

Step 4: Screening of high risk, low uncertainty species list with horizon scan criteria i, ii and iii. Verification by expert review

Potential IAS amenable to prevention and early eradication in the Netherlands: 75a + 14b species

Step 5: Inventory of origins, pathways and impacts

Step 6: Prioritisation and implementation of management measures

Figure 3.1: Processes involved in the horizon-scan method including number of species considered at each step of the Netherlands case study. a Number of species selected in steps one to four; b Number of species selected by taxonomic experts in step four.

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Step 3: A single aggregated risk score is derived either by calculating the average of all harmonised risk scores for each species, or by applying the maximum harmonised risk score which reflects a more precautionary approach. The average score was applied in the case study of the Netherlands for two reasons: 1) Managers are limited in terms of the number of high risk species that can be targeted owing to resource limitations. The application of the maximum score reduces the ability of the method to discriminate between the highest risk species and other species thereby greatly increasing the number of species classified as high risk; 2) Application of the maximum score neglects lower risk classifica- tions from other climatically matched regions. The species are then ranked according to the aggregated score in order to produce a risk-prioritised list of species from climatically matched regions. If only one harmonised risk score exists for a particular species then an aggregation of harmonised risk scores is not required, e.g., if a risk analysis has already been carried out for the target region (see step 1).

Figure 3.2: Matrix displaying prioritisation method for harmonised risk classifications.

The matrix presented in Figure 3.2 is a visual representation of the allocation of a final risk classification to the aggregated risk score, and how a measure of certainty is applied. Species are risk prioritised using the aggregated risk score as follows: aggregated risk score >2: high risk; aggregated risk score ≤ 2 and >1: moderate risk; aggregated risk score ≤ 1: low risk. Uncertainty thresholds were set using an arbitrary method that assumes that an aggregated risk score derived from a certain number of risk classifications is certain. This method is also dependent on the level of risk that a species poses. “Low uncertainty” is

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Identifying potentially invasive alien species and their introduction pathways

assigned to low risk, and moderate to high risk species with aggregated risk scores derived from ≥ 4 and ≥ 2 harmonised risk scores, respectively. The low risk group is allocated a stricter limit because of the potential for unobserved negative impacts. “High uncertainty” is assigned to all other species (hatched areas, Fig. 3.2). Alternatively, certainty may be assessed by calculating the degree of variability from the mean of the underlying harmonised risk scores by deriving the standard deviation. Species ranked in groups other than the high risk, low uncertainty group should be reviewed periodically to determine whether a revision of either risk and / or certainty categories is required. For example, revisions may be required due to future climate change or if new introduction pathways are identified.

Step 4: High risk, low uncertainty alien species are then screened by experts using the following three criteria:

i. The alien species has not been recorded, but will be able to reproduce, in the target region. ii. The alien species has been recorded, is able to reproduce, but is kept only in captivity in the target region. iii. The alien species is able to reproduce and shows a limited distribution within the target region i.e. records exist for one or two locations only.

If the species is absent from the target area, the ability to reproduce is considered using available information from scientific literature and expert judgement in relation to climate match, habitat requirements, and the abiotic tolerances of the species in question. If the species is already present in the target region, ability to reproduce may be confirmed through the observation of a breeding population. Species fulfilling criterion i, ii or iii are classified as potential IAS for the target region amenable to prevention and early eradication. However, not all potential IAS (high risk species) will have been risk assessed for regions climatically matched to the target region and will have been excluded from steps 1 to 3. Therefore, expert consultation may be used to add species to the list of potential IAS for species where there is both a high level of certainty concerning their potential invasiveness, and their distribution complies with criteria i, ii and iii of the horizon-scan. In this case certainty is assessed according to the judgement of the relevant experts, which differs from the method applied in step 3 where the number of risk classifications is used to assess certainty.

Step 5: An inventory of the origins, pathways and potential ecological impact types of species on the list of potential IAS for the target region is undertaken by performing a literature search. The sources outlined in Appendix 3 were used in the literature search carried out during the Netherlands case study. Potential IAS are categorised according to habitat (e.g., terrestrial plant, freshwater animal, marine animal); taxonomy (i.e., mammals,

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fish, birds, amphibians, reptiles, macroinvertebrates and plants); introduction pathways utilized; and ecological impact type (i.e., competition, disease or other health effect / parasite carrier, habitat modification, predation, herbivory, introgression). Pathway information is classified according to the United Nations Environment Programme (UNEP) classification (UNEP 2014b): release in nature, escape from confinement, transport contaminant, transport-stowaway (i.e., the moving of live organisms attached to transporting vessels and associated equipment and media), and corridor (e.g., interconnected waterways). Theses pathways are divided into a number of sub-pathways that allow a more specific classification (Appendix 1). This system has been accepted as the official pathway classification for the EU.

Information may be obtained from horizon-scans, risk assessments, international databases and information portals, e.g., The European Network on Invasive Alien Species (NOBANIS), Delivering Alien Invasive Species Inventories for Europe (DAISIE), the Invasive Species Compendium, GB Non-native Species Secretariat), the World Register of Marine Species (WoRMS), and the European Alien Species Information Network (EASIN) (refer to Appendix 3 for a complete list of sources). It should be noted that, at the time of writing, DAISIE had not been updated since 2008.

Step 6: The recorded impact types of the potential IAS for the target region derived in step 5 are classified per introduction pathway and species group. The numbers of recorded impact types are then ranked to identify priority IAS groups and introduction pathways. Ranking is undertaken according to the frequency rather than severity of impact types. It is assumed that the frequency of impact types may be used to prioritise species groups and pathways for management interventions (preventative, early detection and rapid response systems), potentially leading to the greatest reduction in recorded impacts per intervention, thereby increasing cost-efficiency.

Step 7: A periodical review is recommended of all species identified as high risk with high uncertainty to determine if newly published risk assessments could lead to a reduction in the uncertainty scores. High risk species with reduced (low) uncertainty should then be screened using the criteria listed in step 4 to assess their suitability for addition to the list of potential IAS for the target region.

3.3.2 Case study

The horizon-scan framework was applied to create a list of potential IAS for the Netherlands that are most likely amenable to management intervention. All macro- organism species groups were considered for inclusion in the list (excludes viruses, bacteria, fungi and unicellular organisms). The Netherlands was considered climatically matched to Belgium, Germany, northern France, Denmark, Luxembourg, Ireland and

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Identifying potentially invasive alien species and their introduction pathways

England, and risk classifications for species were obtained for these regions (Step 1, Table 3.1). Due to resource limitations, the climate match was carried out with Western European countries only. We assume that risk assessments and horizon-scans carried out in Western Europe will incorporate species that are traded internationally and also imported from other parts of the world. Therefore, a global climate match is preferable if sufficient resources are available.

Table 3.1: Definition and harmonisation of individual risk classifications taken from regions climatically similar to the Netherlands. Classification Category Reference / website system / protocol Invasive Species Watch list Branquart (2007); Environmental Black list Alert list (Moderate Parrott et al. (2009); Impact Assessment (High risk) (High risk) risk) http://ias.biodiversity.be (ISEIA) Fish Invasiveness Copp et al. (2009); Low Medium High Scoring Kit (FISK) CEFAS (2010) Observation Danish Ministry of Danish list system Black list list Environment (2014) German- Austrian Rabitsch et al. (2013); black list White Black list, Nehring et al. information System list Grey list (2010a,b); Essl et al. (GABLIS) (2011) UK-adapted Low Moderate Urgent, Australian Weed Thomas (2010a) risk risk critical Risk Assessment RINSE meta-list Black list Alert list Gallardo et al. (2013)

North American Non- 1 2 1 Watch list GLANSIS (2014) Great Lakes indigenous

Priority list Watch list Irish risk (most (most Kelly et al. (2013) classification system unwanted, unwanted, amber)1 amber)1 Muséum national Invasive species list Listed1 d'Histoire naturelle (2013)

Standardised risk 1 2 3 3 classification 1Species classified as high risk according to the precautionary principle (Raffensperger & Tickner 1999); 2(National Oceanic and Atmospheric Administration – NOAA, Great lakes aquatic nonindigenous species information system - GLANSIS).

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A list of additional species was obtained for the North American Great Lakes, representative of eastern American regions such as the Hudson and Chesapeake Bay river catchments, because of shipping pathways originating from these areas that have been strongly associated with alien species introductions to the Netherlands (Leuven et al. 2009). Subsequently, risk classifications were harmonised using the standardised risk classification displayed in Table 3.1, and the harmonisation method critically reviewed by experts from a number of Dutch environmental organisations during a workshop (Step 2). The orga- nisations involved included consultancies in nature conservation (Bargerveen Foundation, Natuurbalans-Limes Divergens), non-governmental organisations specialised in ecological field surveys and data compilation for various taxonomic groups, The Mammal Society (Bureau van de Zoogdiervereniging); Reptile, Amphibian and Fish Conservation Netherlands (RAVON); Sovon, The Dutch Centre for Field Ornithology; The Netherlands Floristic Research Foundation (FLORON); and the Radboud University, Institute for Water and Wetland Research, Department of Environmental Science and Department of Animal Ecology and Physiology.

Harmonised risk scores were aggregated by calculating the average score (aggregated risk score). Species were risk prioritised according to their aggregated risk scores (Step 3, Fig. 3.2). The experts representing the organisations listed above were requested to screen species classified as high risk and low uncertainty by applying the horizon-scan criteria (Step 4). The resulting list comprised potential IAS that are amenable to prevention and early eradication measures in the Netherlands (Appendix 2).

Information on habitat type, taxonomical group, pathways and ecological impact types derived during the inventory and analysis (Steps 5 and 6) was collected during a literature search from a variety of sources (Appendix 3) using the official scientific name as the search term for each species according to the Integrated Taxonomic Information System (ITIS 2016). These data were supplemented by information found in all available Dutch, Belgian and British risk assessments carried out for these species. Additionally, a statistical analysis examining the significance of each introduction pathway, shared by different IAS, to the number of recorded ecological impact types recorded for those IAS was carried out by applying Chi squared and Fisher’s exact tests. Data for the statistical analysis was obtained during the literature search. A significant result suggests that management of the introduction pathway may lead to a reduction in the recorded impacts of the IAS related to it. Valid results were obtained from the Chi-squared tests of the escape from confinement and transport stowaway pathways only. All other pathway data sets violated the assumptions of Chi-squared and a Fisher’s exact test was applied instead of Chi-squared in these cases. The level of association between significantly related pathway types and ecological impact types was analysed using Cramer’s V statistic. Statistical analyses were carried out using IBM SPSS 20.

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Identifying potentially invasive alien species and their introduction pathways

3.4 Results

The number of species considered at each step of the case study horizon-scan for the Netherlands is displayed in Figure 3.1.

3.4.1 List of species that have been evaluated for ecological risk in regions climatically matched to the Netherlands

Risk classifications determined for regions climatically matched to the Netherlands were collected for 1425 species, hybrids, varieties and subspecies (Step 1; Fig. 3.1). The number of risk classifications collected per species ranged from one to eight. Following the harmonisation, aggregation and ranking of risk scores (Steps 2 and 3) 243 species were prioritised as high risk, low uncertainty species.

3.4.2 Risk prioritised species amenable to prevention or early eradication in the Netherlands

Of the 243 high risk, low uncertainty species, 75 satisfied the horizon-scan criteria and were classified as potential IAS amenable to prevention and early eradication in the Netherlands (list of potential IAS for the Netherlands). Fourteen additional potential IAS were added based on expert judgement (Step 4).

Table 3.2: Risk prioritised species. 1 Aggregated Number of Uncertainty Prioritisation matrix colour code risk score species

2 >2 75 Low

>2 600 High

>1 and ≤ 2 31 Low

>1 and ≤ 2 117 High

≤ 1 0 Low

≤ 1 434 High

1Refer to figure 3.2; 275 species were derived from the horizon-scan method. A further 14 were added based on expert judgement. In total 89 species were prioritised as high risk species (potential IAS for the Netherlands).

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The 1182 high risk (high uncertainty), and medium and low risk (high and low uncertainty) alien species were not screened with the horizon-scan criteria due to time and budgetary limitations. Six hundred of the 1182 species were allocated to the high risk, high uncertainty list; 31 species were allocated to the medium risk, low uncertainty list; 117 species were allocated to the medium risk, high uncertainty list; zero species were allocated to the low risk, low uncertainty list and 434 species were allocated to the low risk, high uncertainty list (Table 3.2, Fig. 3.2).

3.4.3 Inventory of list of potential IAS for the Netherlands (Steps 5 and 6)

Freshwater and terrestrial animals and terrestrial plants appeared most frequently on the list of potential IAS for the Netherlands, followed by marine animals and freshwater plants (Fig. 3.3a). Asia and North America were the most frequently listed geographical origins (native range) of concern, followed by Russia, South America, Africa and southern Europe (Fig. 3.3b). Hybrid species were considered not to have a native range. The pathway utilised most frequently for potential IAS for the Netherlands was escape from confinement (Fig. 3.4a).

A B

Eastern Europe (n=2) Northern Europe (n=1) Freshwater animals (n=25) Southern Europe (n=5) Ponto-caspian region (n=3) Freshwater plants (n=4) Russia (n=9) Africa (n=7) Arctic (n=1) Asia (n=33) Marine animals (n=7) Australia / New Zealand (n=3) North America (n=28) Terrestrial animals (n=25) Central America (n=2) South America (n=9) Terrestrial plants (n=28) Atlantic ocean (n=1) Pacific Ocean (n=2) Not applicable (n=2) Unknown (n=1)

Figure 3.3: (a) Relative contribution of ecosystem groupings to the list of potential IAS for the Netherlands; (b) Geographic origin of species present on the list of potential IAS for the Netherlands.

The most frequently utilised sub-pathways classified under the escape from confinement pathway were the pet and aquarium trade, and introductions for ornamental purposes, both of which are associated mainly with freshwater and terrestrial animals, and terrestrial plants. Terrestrial plants are associated with the horticulture sub-pathway, while the

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Identifying potentially invasive alien species and their introduction pathways

botanical and zoological garden sub-pathway is utilised mainly by terrestrial plants and animals. A 70 Release in nature 60 Escape from confinement Transport - contaminant 50 Transport - stowaway Corridor 40 Other

Number of species Number 30

20

10

0

B 90 Escape from confinement 80 Transport - stowaway

70 Transport - contaminant Release in nature 60 Corridor 50 Other

40

Number of impact types of impact Number 30

20

10

0 Competition Habitat Disease or Predation Herbivory Introgression modication other health effect / parasite carrier

Figure 3.4: (a) Possible introduction pathways of potential IAS to the Netherlands per taxonomic group; (b) Number of impact types recorded in regions climatically matched to the Netherlands per introduction pathway.

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Identification of introduction pathways associated with IAS that cause the highest number of ecological impact types allows prevention measures to be applied on a wider and potentially more cost-effective scale than measures aimed at individual IAS. Overall, the escape from confinement pathway was associated with the highest number of ecological- impact types attributed to the potential IAS for the Netherlands, followed by the transport stowaway and transport contaminant pathways (Fig. 3.4b). The most frequently occurring impact types that were attributed to species utilising the escape from confinement pathway were competition with native species, followed by habitat modification, and disease or other health effect / parasite carrier. Impact types associated most frequently with the transport stowaway pathway were competition, predation, and habitat modification. The most frequently occurring impact types relating to the transport contaminant pathway were competition, predation, and disease or other health effect / parasite carrier.

Competition Fur farms Disease or other health Other escape from confinement (fencing) effect / parasite carrier Habitat modication Forestry Predation Farmed animals, including animals under limited control Herbivory

Agriculture Introgression

Aquaculture

Research (in facilities)

Horticulture

Botanical garden/zoo/aquaria

Ornamental purpose

Pet/aquarium trade

0 10 20 30 40 50 60 Number of ecological impacts

Figure 3.5: Occurrence of impact types per pathway classified as escape from confinement.

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Identifying potentially invasive alien species and their introduction pathways

The results of the Chi-squared and Fisher’s exact tests indicated that there is a significant relationship between the tested pathways and the number of ecological impact types χ2 (5) = 14.052, P < 0.05 and P < 0.05 respectively. Cramer’s V statistic was 0.22 (P < 0.05) indicating a moderate association between the escape from confinement and transport– stowaway pathways, and the number of ecological impact types. Therefore, effective management interventions targeting the introduction pathways of potential IAS rather than individual species may be an effective approach leading to reductions in the number of ecological impacts recorded.

Figure 3.5 details the number of ecological impact types attributed to sub-pathways classified under escape from confinement. The identification of sub-pathways encourages further improvements in cost-effectiveness by targeting specific sub-pathways associated with IAS with the highest number of impact types. The highest frequency of impact types occurred for the pet and aquarium trade, followed by the trade in ornamentals and escapes from botanical gardens, zoos or aquaria. Of these three sub-pathways, competition, habitat modification and disease or other health effect / parasite carrier impact types were most frequently recorded.

3.5 Discussion

3.5.1 Horizon-scanning framework

This study presents a novel framework for horizon-scanning that maximally benefits from existing national and regional knowledge, such as horizon-scans, risk assessments and IAS lists, by applying a harmonisation and aggregation method to risk classifications that produces a list of potential IAS amenable to prevention and early eradication measures for a target region. The framework may be applied on a national, international or continental scale. Only species with an existing risk classification from climatically matched regions can be included in our horizon-scanning approach. Other species are assessed using expert judgement. However, the application of existing classifications greatly reduces the number of species requiring expert judgement compared with traditional risk prioritisation methods. Expert judgement may be supplemented by a literature study examining, for example, species traits that may increase the risk of invasiveness in the target region.

The horizon-scanning process aimed to produce a list of IAS that are likely to negatively impact biodiversity in the target region (high risk species with low uncertainty). The list provides a strong basis for cost-efficient management measures. Expert judgement was applied to verify this high risk list and identify other high risk species omitted from the list because they were not previously risk prioritised or assessed for climatically matched regions. Expert knowledge was not used to screen the 600 high risk, high uncertainty

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species because of time and budgetary limitations. However, further screening using expert judgement may be applied if time and budget allows.

A number of sources of uncertainty have been identified during the design and application of the horizon-scan framework. The principle that existing risk classifications may be applied to other, climatically matched regions is supported by Wittenberg & Cock (2001) who state that the only variable consistently correlated with invasiveness in one region is invasiveness elsewhere (Verbrugge et al. 2012a). However, regional differences in assessment methodology will introduce uncertainty. Ecological impacts vary and are weighted differently according to region-specific habitat characteristics and conservation aims (Verbrugge et al. 2012a). During the harmonisation process we assumed that the risk level is the same even though risk classifications differ. Moreover, variation between risk classifications may also stem from local environmental variables other than those controlled for when a climate match is applied, e.g., predator abundance, vegetation cover and soil type. To reduce this uncertainty and resolve possible inconsistencies between national risk classifications, only species that received a high risk rating in two or more of the surrounding countries are included in the list of potential IAS for the target region. Moreover, the application of the precautionary principle when harmonising risk categories prior to the aggregation of harmonised risk scores and when interpreting the potential impacts of IAS reduces the potential for underestimation of risks. Finally, it could not be established with certainty whether criterion 3 of the horizon-scan was fulfilled by a high proportion of alien macroinvertebrate species due to a limited direct availability of distribution data for the Netherlands.

During the application of the horizon-scan framework, either the average or the maximum harmonised risk score may be used to derive the aggregated risk score for a particular species. Uncertainty reduction is a requirement of the EU regulation on IAS and will also facilitate public support and acceptance by stakeholders. The average aggregated risk score was derived from all harmonised risk scores for a particular species that leads to a reduction in influence of individual methodologies, knowledge obtained from small expert groups, locally specific habitat conditions and varying national conservation goals. However, the application of the maximum risk score is in agreement with the precautionary principle (Raffensperger & Tickner 1999) and the observation that the only variable consistently correlated with invasiveness in one region is invasiveness elsewhere (Wittenberg & Cock 2001). It is expected that, contrary to an approach using the average score, the application of the maximum score will produce a relatively long list of high risk species which may require further risk prioritisation and may be impractical in view of limited budgets. The average aggregated risk score was used in the case study of the Netherlands in order to produce a relatively short list of potential IAS directly amenable to cost-effective management interventions.

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Identifying potentially invasive alien species and their introduction pathways

The horizon-scan framework omits alien species that have not been previously risk assessed in climatically matched regions. Moreover, due to resource limitations, the case study includes a climate match between the Netherlands and other European countries only, which may have resulted in the omission of risk classified alien species originating from climate matched regions outside Europe. It is recommended that a climate match is undertaken between the target region and all other countries or regions. High risk species assessed in regions not climatically matched to the target region may become invasive in the target region because of a climatic tolerance that is not reflected in their global distribu- tion. To account for this, more emphasis is focused on the identification of dominant introduction pathways during the meta-analysis. The results of the analysis can be used in the design of pathway-based measures that indirectly target potential IAS that use dominant pathways but not identified during the horizon-scan. In addition, expert knowledge is applied to add potential IAS that either have broad environmental tolerances or have not been previously assessed for climatically matched regions, but are considered, with a high degree of certainty, to pose a high ecological risk to the target region and fulfil at least one of the horizon-scan criteria. However, expert knowledge may not always be objective, accurate, consistent or reproducible (Hulme 2012). To counter this, our horizon-scan method utilises experts from multiple organisations representing as many taxonomic groups as possible. Taxon bias may be reduced by ensuring that specialists are equally distributed over taxonomic groups and have similar experience. The consultation of experts with complementary knowledge across taxonomic groups and environments ensures a broad collective knowledge base to undertake a comprehensive horizon-scan in an open, rigorous and time-efficient way (Roy et al. 2014a). Moreover, by using all available horizon-scans and priority lists from climatically matched countries, the results of a wide range of assessments based on multiple expert opinions are aggregated, significantly reducing the effect of potential individual errors. However, it is important to note the unpredictable nature of IAS introductions and the resulting imperfect nature of horizon-scanning lists (Roy et al. 2014a). For example, a determination of the actual temperature tolerance of a species can only be carried out under laboratory conditions due to the multitude of potentially confounding factors that exist in the field. Moreover, alien species that may not establish in an average year, may establish during years characterised by temperature extremes, with the advent of future climate change, or in urban locations where local climate conditions may be more favourable (Vermonden et al. 2010; Leuven et al. 2011; Verbrugge et al. 2012b; Collas et al. 2014). However, a precautionary approach to climate matching which takes inter-annual temperature extremes and climate change into account could address this issue. These potential limitations suggest that the horizon-scanning framework should be viewed as a single tool in a suite of measures to aid the design of management interventions aimed at preventing the impacts of potential IAS.

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3.5.2 Case study results

Our results pertaining to the geographical origin of the list of potential IAS for the Netherlands (Fig. 3.3c) are supported by Welcomme (1991), Vilà et al. (2009) and García- Berthou et al. (2005), who state that, in contrast to other taxa such as plants and terrestrial vertebrates, freshwater species introduced to Europe come mostly from North America and enter through mid-latitude countries in Western Europe (France, the UK, and Germany).

Trade pathways are frequently linked to potential IAS introductions globally (Bowmer et al. 1995; Randall & Marinelli 1996; Kay & Hoyle 2001; Perrings et al. 2005; Westphal et al. 2008; Brunel 2009; Matthews et al. 2013a). For example, one third of the aquatic species listed in the Invasive Species Specialist Group’s top 100 worst invasive species, and 40% of plant species introductions to Europe in general are attributed to the trade in ornamental plants (Padilla & Williams 2004; Gooijer et al. 2010; Martin & Coetzee 2011). This is similar to the combined contribution made by the ornamental and aquarium plant trade, and landscape or floral improvement pathways observed in our study (35% of all introductions of potential IAS). Internet retailing, and plant and animal hobbyist forums facilitate international transactions (Kay & Hoyle 2001; Westbroek 2014; Faulkes 2015; Humair et al. 2015; Mazza et al. 2015), and are implicated in the introduction of a number of species in the Netherlands, e.g., curly waterweed (Lagarosiphon major), fanwort (Cabomba caroli- niana), striped skunk (Mephitis mephitis), Reeves’ muntjac (Munctiacus reevesi (Ogilby, 1839)), sika deer (Cervus nippon Temminck, 1838) and non-native squirrel species (Van Belle & Schut 2011; Matthews et al. 2012b; Matthews et al. 2013a,b; Dijkstra 2014; Hollander 2015).

We observed a high frequency of impact types relating to the pet and aquarium trade, and the trade in ornamentals and escapes from botanical gardens, zoos or aquaria. This combined with the statistically significant relationships observed for introduction pathways and ecological impact types suggests that effective management interventions focussing on these pathways will yield the largest reduction in the number of ecological impacts of potential IAS for the Netherlands. A single pathway focussed intervention may prevent the introduction of many potential IAS. For example, voluntary covenants may prevent the introduction of multiple potential IAS via trade pathways (Verbrugge et al. 2014). In the absence of punitive tariffs, watertight trade regulations and certainty in risk assessment (Perrings et al. 2005), preventative measures, early detection and rapid response systems that focus on pathways management represent the frontline in the prevention of biological invasions (Hulme 2009).

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Identifying potentially invasive alien species and their introduction pathways

3.5.3 Implications for policy

The EU is introducing a system of prevention, early detection and rapid response measures to protect member states against the impacts of IAS (European Commission 2014, 2015). Important elements of this system are the comprehensive analysis of pathways of unintentional introduction and spread of IAS in member states and the identification of pathways which require priority action (European Commission 2013). The horizon-scan and inventory presented here address this need by firstly identifying potential IAS and secondly identifying pathways and impact types specific to these species. The identification of priority pathways will assist in the design of effective border controls consisting of a targeted and cost-effective surveillance system that facilitates the rapid response to invasions required by the EU (European Commission 2013). Regular updates to the list of potential IAS and identification of newly emerging introduction pathways by periodically repeating the horizon-scan will facilitate the early detection of potential IAS in the target region. For example, ecosystem change brought about by habitat modifiers (eco-engineers) is one of the most dramatic ecological effects of IAS (Crooks 2002). If habitat modifiers are introduced disproportionately through a particular pathway, that pathway should be prioritised. Furthermore, impact analyses per introduction pathway may be incorporated into a pathway-based risk analysis that is used to assess and manage all risks moving along the same pathway and, for example, provide input for the design of instruments to prevent IAS introduction such as blacklisting, the direct regulation of pathways such as ballast water, and codes of conduct (Shine et al. 2010; Matthews et al. 2012d; Verbrugge et al. 2014). The horizon-scan method may also be used to establish a white list of low risk species (Shine et al. 2010), that could be applied on an EU scale. Any species not on this white list would require screening prior to importation to an EU member state.

3.5.4 Further research

Uncertainty and the application of expert knowledge continue to pose a challenge for the assessment of potential impacts of IAS in regions where the species are absent. Moreover, the assertion that impacts that occur in one region will occur in another, based on a climate match alone, neglects the fact that successful invasion depends on multiple other environmental variables such as predator-prey interactions, competition with native species and diseases. Future research should be aimed at reducing these areas of uncertainty in the risk assessment of IAS. Additionally, an exploration of alternative approaches to the validation of standardised risk scores, additional to expert consultation, is desirable. Future methodological development may be aimed at revealing the contribution that risk assessment methods make to the aggregated risk score, relative to less rigorous risk prioritisation methods. Further research that identifies the relative severity of impacts in climatically matched regions may improve the accuracy of risk prioritisation and assessment of IAS and their introduction pathways. It is recommended that further research

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is conducted on the distribution of alien macroinvertebrate species with suspected limited distribution in the Netherlands and that future horizon-scans consider microorganisms and meiofauna. In addition, symbionts, parasites and commensals were not considered in this study and future updates to the horizon-scan for potential IAS in the Netherlands should incorporate these species groups.

3.5.5 Conclusions

The novel approach to horizon-scanning presented here may be applied on any continental scale, maximises the use of available data from climatically matched countries, and elucidates the relationship between species groups, pathways and impacts. Horizon- scanning provides a starting point for the design of preventative, early identification and rapid action measures for the effective management of potential IAS.

Acknowledgements

We would like to thank the Netherlands Food and Consumer Product Safety Authority, Invasive Alien Species Team (NVWA-TIE) of the Dutch Ministry of Economic Affairs for financially supporting this study (project 62001981 ‘Horizon-scanning for new invasive non-native species in the Netherlands’). We would also like to thank Gert Ottens of Vogelbescherming Nederland who allowed us to use his database on alien bird species, and Melanie Gillings and Polly Bryant of the RINSE project for the supply of data and information that contributed to the horizon-scan. We acknowledge two anonymous reviewers and Sarah Bailey who made constructive comments which led to significant improvements to this article.

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4 Inconsistencies in the risk classification of alien species and implications for risk assessment in the European Union

Jonathan Matthews Gerard van der Velde Frank P.L. Collas Lisette de Hoop K. Remon Koopman A. Jan Hendriks Rob S.E.W. Leuven

Ecosphere 8(6): e01832. 10.1002/ecs2.1832. June 2017. Inconsistencies in the risk classification of alien species

4.1 Abstract

Invasive alien species (IAS) are species whose introduction or spread outside their native range threatens biological diversity, ecosystem functioning, economy and/or public health. The recent European Union (EU) regulation on the management of IAS emphasises the need for a consistent approach to alien species assessment that will underpin international measures for the early identification of newly introduced IAS followed by rapid action aimed at the prevention of introduction, spread and negative impacts. The goals of the present study are (1) to present the risk classifications of 18 aquatic alien species for the Netherlands, using the Invasive Species Environmental Impact Assessment (ISEIA) protocol, (2) to compare these with available risk classifications made for countries spanning similar climatic and biogeographical regions to the EU, and (3) to provide explanations for inconsistencies between different risk classifications. Five species were classified as high risk: Carassius gibelio (Prussian carp), Cyprinus carpio (common carp), Sander lucioperca (pike-perch), Cabomba caroliniana (fanwort), and Dreissena rostriformis bugensis (quagga mussel). Of the 14 species with existing risk classifications for countries spanning similar climatic and biogeographical regions to the EU, all but two of the assessed species (Carassius gibelio and Dreissena rostriformis bugensis) were classified inconsistently. Reasons for these inconsistencies are the application of different risk assessment schemes, application on a national rather than biogeographical scale, differences in the definition and application of criteria, differences in habitat availability, and uncertainties that are intrinsic to risk assessment methodologies. Approaches that increase transparency by highlighting these methodological aspects, normative choices and uncertainty are vital to the legitimacy of any risk assessment method and will increase acceptance among decision makers, nature managers and stakeholders.

4.2 Introduction

Invasive alien species (IAS) are species whose introduction and/or spread outside their natural past or present distribution threatens biological diversity, economy and/or public health (United Nations 1992; Verbrugge et al. 2016). IAS are considered to be one of the leading causes of global biodiversity loss (Moyle et al. 1986; Vitousek et al. 1997; García- Berthou et al. 2005; McGeoch et al. 2010). Costs relating to monitoring, eradication, control and impact mitigation of IAS are significant. Reported estimates for the European Union (EU) and globally are 12.5 billion (Kettunen et al. 2008) and 1.4 trillion euros per year (Pimentel et al. 2001), respectively. Multiple introductions of potential IAS within the EU can be attributed to human mediated pathways. For example, the recent rise in (online) retailers offering species for sale internationally has been considered to be mainly responsible for the introduction of many IAS (Faulkes 2015; Humair et al. 2015; Mazza et al. 2015).

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Aquatic ecosystems are particularly vulnerable to the impacts of species invasion. Declines in biodiversity as a result of IAS introductions have been more severe for aquatic species relative to terrestrial species in the most affected ecosystems (Dudgeon et al. 2006). Impacts may be dramatic because freshwater systems feature the greatest density of species per unit surface area on the planet (Thomaz et al. 2015; Lozano & Brundu 2016). IAS negatively impact aquatic communities, particularly macrophyte, zooplankton and fish assemblages, and promote physical alterations in nitrogen and organic matter concentration, and changes to water turbidity (Gallardo et al. 2016). Examples of IAS that have been introduced to aquatic ecosystems in Western Europe as a result of the international trade in alien species are Lagarosiphon major (curly waterweed), Cabomba caroliniana (fanwort) and Lepomis gibbosus (pumpkinseed sunfish) (Kay & Hoyle 2001; Van Kleef et al. 2008; Matthews et al. 2012b, 2013a).

The Convention on Biological Diversity (CBD) specifically requires that signatories prevent the introduction, and control or eradication of potential IAS (United Nations 1992). Moreover, the European Commission (EC) has stated that a significant subset of alien species can become invasive, leading to serious adverse ecological, social and economic impacts which should be prevented (European Commission 2014). Most expenses generated by IAS in Europe result from reactive eradication measures (Colautti et al. 2006b; Vilà et al. 2009; Sinden et al. 2011). As a result, measures that prevent the initial introduction and spread of IAS have become the most attractive management approach (Pyšek & Hulme 2005; Shirley & Kark 2006; Essl et al. 2011; European Commission 2014). Therefore, predicting which alien species may become invasive by applying a risk assessment process has gained much interest as a method to support the management decisions of policy makers (Byers et al. 2002; Andersen et al. 2004; Verbrugge et al. 2012a; Vanderhoeven et al. 2015). Risk assessment has already proven to be an important tool for assessing which alien species should be subject to trade restrictions, eradication and control measures that are fundamental to the EUs strategy on the prevention and management of the introduction and spread of IAS (European Commission 2014; Roy et al. 2015).

In this study we define risk as the chance that a particular hazardous event (e.g., competition, hybridization) may actually cause damage, and regard it as a product of three factors: exposure * likelihood * consequence (D’hondt et al. 2015). Risk assessment is defined as the identification of risks and their assessment with regard to the likelihood and consequences of the introduction, establishment, spread, and impact of an alien species using science-based information (CBD COP6 Decision VI/23 2002). Risk classifications are the outcomes of risk assessments, and are usually expressed using three or five risk classes (e.g., low, medium or high).

The introduction of the new regulation (EU) No 1143/2014 of the European Parliament and of the council 22 October 2014 on the prevention and management of the introduction and

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spread of IAS requires the standardisation of alien species risk assessment and management across Europe. One of the major tasks relating to the regulation is the creation of a list of IAS considered to be of Union concern through the application of standardised criteria (European Commission 2014). Until recently, legislative and regulatory requirements applied to EU member states concerning the risk assessment and management of potential IAS have been fragmented (Hulme 2009, Verbrugge et al. 2010). This has resulted in the development of a number of different risk assessment approaches, e.g., the Invasive Species Environmental Impact Assessment (ISEIA) protocol (Branquart et al. 2009, Vanderhoeven et al. 2015), the German-Austrian Black List Information System (GABLIS) (Essl et al. 2011) and the Non-Native Species Risk Analysis Mechanism of Great Britain (NNSS 2016). Moreover, risk assessment methods are usually applied to areas defined politically instead of ecologically, i.e., per country rather than per biogeographical region. However, once established, IAS can rapidly expand their range across national borders (Shirley & Kark 2006). The lack of methodological consistency and differences in ecological context between risk assessments prevents direct comparisons of risk classifications and potentially hinders a standardised European approach that facilitates rapid preventative action in different countries within the same biogeographical region (Verbrugge et al. 2012a). In addition, limited attention has been devoted to the influence of different approaches to uncertainty applied by different risk assessment methods used to classify the risks of alien species in Europe (Essl et al. 2011; Verbrugge et al. 2012a), hindering the potential for future methodological improvement.

The aims of this paper are as follows:

1) Present the risk classifications of 18 aquatic alien species for the Netherlands, derived using the ISEIA protocol, that can be used for prioritisation and to design management measures in lowland North-Western Europe (e.g., the Netherlands, Belgium, Germany and Denmark). 2) Analyse the level of consistency between risk classifications from the Netherlands and those derived for countries spanning similar climatic and biogeographical regions to the EU. 3) Discuss possible reasons for potential inconsistencies between risk classifications, their implications for the prioritization of IAS on an EU level and opportunities for uncertainty reduction.

4.3 Methods

4.3.1 Species selection

The species were selected from a group that were risk assessed for the Netherlands Food and Consumer Product Safety Authority (NVWA) by Radboud University in association

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with partner organisations of the Netherlands Centre of Expertise for Exotic Species (NEC- E) who specialise in monitoring and field studies of particular species groups. The species were selected on the basis of the following criteria: (1) the species is alien to the Netherlands and potentially invasive, (2) no valid risk classification existed for the species in the Netherlands prior to the ones presented here, and (3) the species is recorded in freshwater habitats. The risk assessed species were chosen to facilitate government decisions regarding their management and to allow the sound evaluation of proposals for the listing of IAS of EU concern relating to the new regulation (EU) No 1143/2014 of the European Parliament and of the council 22 October 2014 on the prevention and management of the introduction and spread of IAS. The assessed species include twelve fish, four macrophyte, and two mollusc species (Appendix 4-6). While all of these species are alien to the Netherlands, eleven are also alien to Western Europe: Romanogobio belingi (northern whitefin gudgeon), Ctenopharyngodon idella (grass carp), Oncorhynchus mykiss (rainbow trout), Salvelinus fontinalis (brook trout), Umbra pygmaea (eastern mud- minnow), C. caroliniana (fanwort), Egeria densa (Brazilian waterweed), L. major (curly waterweed), Vallisneria spiralis (tapegrass), Bellamya chinensis (Chinese mystery snail), and Dreissena rostriformis bugensis (quagga mussel). All fish were assessed because they are included in the Dutch Fisheries Act that allows introduction by the holders of fish rights of these species if certain criteria are complied with. However, the potential risks of these species had not been previously assessed for the Netherlands (Schiphouwer et al. 2014). The macrophyte and mollusc species were assessed to provide more insight into their current distribution, probability of entry, establishment and spread, and potential impacts in the Netherlands.

4.3.2 Literature surveys for risk inventories

Literature reviews were carried out to provide an overview of the current knowledge on the distribution and invasion biology (i.e., a risk inventory) of each species listed in Appendix 4-6. Literature data were collected on the habitat, physiological tolerances, substrate preference, colonisation vectors, ecological and socio-economic impacts and potential measures for the management of the species using the official Latin name and frequently applied synonyms. The searches were internet based, supported by the use of a university library. Web of Knowledge and Google Scholar search engines were queried. Web of Knowledge includes only peer reviewed literature, a part of which is not available freely online, while Google Scholar includes both peer reviewed and grey literature that is often freely available online. All search results from the Web of Knowledge were examined while the first 50 results per search term from Google Scholar were examined due to the decreasing relevance of search results returned using this search engine.

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4.3.3 Consensus method for classification of risks

Each risk assessment was carried out by an expert team. Each expert completed an assessment form independently, based on the results of the literature review (i.e., a standard knowledge document with risk inventory). Therefore, each expert based their individual risk scores on the same information. Risks were scored using a hierarchical method where evidence from within the Netherlands was given priority over evidence derived from impacts occurring abroad. Following this individual assessment, the entire expert team met, elucidated differences in risk scores, discussed diversity of risk scores and interpreted key information during a workshop. Discussion during the workshop led to agreement on consensus scores and risk classification relating to the four sections contained within the ISEIA protocol. The consensus scores, risk classification and justifications for the scores were entered into a draft report that was reviewed by the expert team. If complete agreement on the contents of the draft report was not achieved, additional face to face or email discussions between team members took place until consensus was achieved.

4.3.4 Assessment scheme

The ISEIA protocol is a decision support tool that classifies species according to their potential invasiveness, informing the decisions of nature managers, policy makers and stakeholders (Vanderhoeven et al. 2015). ISEIA is applied in the risk assessment of alien species in Europe alongside the more recent Harmonia+ risk assessment protocol, and has been used in the Belgian early warning and rapid response system, an approach that is key to management approaches advocated in the new EU regulation on the prevention and management of IAS (European Commission 2013, 2014; Vanderhoeven et al. 2015). The ISEIA protocol is more extensive than Harmonia+ regarding the assessment of impacts on biodiversity and ecosystem functioning, elements that are key impact categories requiring assessment in relation to the new EU regulation (European Commission 2013, 2014). Moreover, assessments undertaken after the introduction of Harmonia+ also applied the ISEIA protocol to allow direct comparisons between the results of all risk assessments. Risk assessment is carried out through the application of ten criteria that match the last steps of the invasion process (i.e., the potential for spread, establishment and adverse impacts on native species and ecosystems) (Branquart et al. 2009; Vanderhoeven et al. 2015). These criteria are divided over the following four risk sections: (1) dispersion potential or invasiveness, (2) colonisation of high conservation habitats, (3) adverse impacts on native species, and (4) alteration of ecosystem functions. Section 3 contains sub-sections referring to (i) predation / herbivory, (ii) interference and exploitation competition, (iii) transmission of diseases to native species (parasites, pest organisms or pathogens) and (iv) genetic effects such as hybridisation and introgression with native species. Section 4 contains sub-sections referring to (i) modifications in nutrient cycling or resource pools, (ii) physical modifications to habitats (changes to hydrological regimes, increase in water

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turbidity, light interception, alteration of river banks, destruction of fish nursery areas, etc.), (iii) modifications to natural successions and (iv) disruption to food-webs, i.e., a modification to lower trophic levels through herbivory or predation (top-down regulation) leading to ecosystem imbalance. The potential positive impacts of alien species are not considered by the ISEIA protocol.

Each criterion of the ISEIA protocol was scored. Scores ranged from 1 (low risk) to 2 (medium risk) and 3 (high risk). If knowledge obtained from the literature review was insufficient, then the assessment was based on expert judgement and field observation leading to a score of 1 (unlikely) or 2 (likely). If no answer could be given to a particular question (no information) then no score was given (dd - deficient data). Finally, the total score for the species was derived by adding the highest score of each section.

Subsequently, the Belgian Forum Invasive Species (BFIS) list system for preventive and management actions was used to categorise the species of concern (Branquart et al. 2009). This list system was designed as a two dimensional ordination (Environmental impact * Invasion stage; Fig. 4.1).

Figure 4.1: The Belgian Forum Invasive Species (BFIS) list system to identify species of most concern for preventive and mitigation action (Branquart et al. 2009).

It is based on guidelines proposed by the Convention on Biological Diversity (CBD decision VI/7) and the EU strategy on IAS. Species’ ecological impact was classified based on the total risk score which is converted to a letter / risk classification: C – low risk (score 4-8), B – moderate risk (9-10) and A – high risk (11-12). This letter is then combined with

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a number representing the invasion stage: (0) absent, (1) isolated populations, (2) restricted range, or (3) widespread. Absent species scoring moderately or highly for ecological impact are placed on an alert list. Species that have been recorded in the area under assessment and that score a moderate or high risk are placed on a watch list or black list, respectively.

4.3.5 Comparison with other risk classifications

Risk classifications derived from risk assessment methods were obtained from other lowland North-Western European countries (Germany, Belgium, the Netherlands, and the United Kingdom) and from the state of New York (USA) that, according to the Köppen- Geiger climate zones of Kottek et al. (2006), is climatically matched with large parts of Europe (Table 4.1). These risk classifications were then compared to risk classifications derived for the Netherlands with the ISEIA protocol and inconsistencies between classifications were identified. The protocols compared vary in their approach but many share the aim of measuring ecological risk which suggests that their outcomes should be broadly similar when assessing the same species, in similar biogeographical and climate regions.

The FISK and Harmonia+ protocols, and the New York Invasiveness Ranking System contain criteria addressing both ecological and socio-economic impacts. To maintain consistency, we limited comparisons to the ecological risk components of the FISK, Harmonia+ and New York Invasiveness Ranking System assessments. All criteria received equal weighting during all the assessments considered. A number of species were classified according to systems that do not allocate a low, medium or high risk score, applying different terminology to define risk. This prevents direct comparisons of risk classifications produced by different protocols for the same species. Therefore, the following standardisation was applied. Species allocated to the grey and black lists of GABLIS were considered as high risk species. White list species were considered low risk (Essl et al. 2011). Thomas (2010b) applied the categorisation critical (red) in the horizon-scanning for invasive alien plants in Great Britain if a more detailed risk assessment was required. Species allocated to this risk category were assumed to be high priority (high risk) species. Species allocated to both the high and very high risk categories of the New York Invasiveness Ranking System were considered high risk.

In a separate statistical analysis, contingency tables were derived based on the risk classifications for the same set of species that were assessed 1) in the United Kingdom and in the Netherlands, 2) using the New York Invasiveness Ranking System and ISEIA protocol, and 3) using the FISK and ISEIA protocols. Subsequently a Pearson’s Chi- squared test was applied to both contingency tables using R statistics version 3.3.1 (R Project for Statistical Computing 2016).

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Table 4.1: Overview of protocols included in the comparison of risk classifications for the Netherlands and regions with similar climate and biogeography. Protocol Assessment Geographical Taxonomic Ecological Socio- References type scope scope assessment economic assessment Fish Invasiveness Rapid risk United Fish X X2 Copp et al. Screening Kit assessment Kingdom, (2005, 2009) (FISK) Belgium

German Austrian Risk Central All species X - Essl et al. Blacklist assessment Europe (2011) Information (Germany, System (GABLIS) Austria) Harmonia+ Risk European All species X X2 D’hondt et al. assessment Union (2015)

Invasive Species Risk Belgium, the All species X - Branquart et Environmental assessment Netherlands al. (2009) Impact Assessment (ISEIA) Modified Rapid risk United Plants - - Thomas Australian Weed assessment Kingdom (2010b) Risk Assessment (Rapid Risk Assessment)1 New York Rapid risk New York Plants X X2 New York Invasiveness assessment State, USA Invasive Ranking System Species Information (2017) Rapid risk Rapid risk United All species X - Sutherland et assessment and assessment Kingdom al. (2011); consensus method Roy et al. (Horizon-scan) (2014b) 1Defines invasiveness in terms of spread only; 2Scores relating socio-economic risk were removed to allow a direct comparison of classifications relating to ecological impact scores only.

4.4 Results

The risk classifications for all aquatic species prioritised using the ISEIA protocol are displayed in Table 4.2. Five species were classified as high risk, a plant species: C. caroliniana, a mollusc: D. rostriformis bugensis, and three fish species: Carassius gibelio (Prussian carp), Cyprinus carpio (common carp) and Sander lucioperca (pike-perch). D. rostriformis bugensis, C. gibelio, C. carpio and S. lucioperca are widely distributed, while C. caroliniana has a restricted range in the Netherlands.

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Table 4.2: Overview of risk scores and classifications for species assessed with the Invasive Species

Environmental Impact Assessment (ISEIA) protocol.

1 f f parasites and

Scientific name

Predation/herbivory Interference or exploitation Transmission o Genetic effects / (hybridisation Disruption to food webs

. Modification to natural

Taxonomic Taxonomic group

Dispersion potential or Colonisation of high Direct or indirect adverse Direct or indirect alteration

invasiveness conservation value habitats impacts on native species competition diseases introgression with natives) of ecosystem functions cycling or resource pools habitat succession Total score BFIS list category

1. 2. 3. 3.1. 3.2. 3.3. 3.4. 4. 4.1. Modification of nutrient 4.2. Physical modifications of 4.3 4.4. Cabomba 2 3 3 3 na 3 dd 1 3 2 3 2 2 12 A2 caroliniana Lagarosiphon 3 2 2 na 2 dd 1 2 2 2 2 2 9 B1 major3 Egeria densa4 2 2 2 na 2 dd 1 2 2 dd dd dd 8 C1

Macrophytes Vallisneria 3 2 1 na 1 dd dd 1 1 1 1 1 7 C1 spiralis5

Dreissena rostriformis 3 3 3 3 3 dd dd 3 3 3 3 3 12 A3 bugensis6

Molluscs Bellamya 2 3 3 2 3 2 dd 2 2 2 2 2 10 B1 chinensis7 Carassius gibelio8 3 3 3 3 3 1 2 2 29 1 1 1 11 A3

Cyprinus carpio8 3 3 2 2 2 2 2 3 3 3 3 3 11 A3

Sander lucioperca8 3 3 3 3 3 2 1 2 dd dd dd 2 11 A3 Cyprinus carpio X 2 2 3 2 2 1 3 2 2 2 2 2 9 B1 Carassius sp.8 Salvelinus 2 3 2 2 2 2 2 2 dd dd dd 2 9 B1 fontinalis8 Romanogobio

3 3 2 dd 2 dd dd 1 1 1 1 1 9 B2 belingi8

Fish Ctenopharyngodon 1 3 3 3 3 2 1 3 3 3 3 3 10 B3 idella8 Leuciscus aspius8 3 3 2 2 dd 1 1 1 dd dd dd 110 9 B3 Oncorhynchus 2 3 3 3 2 2 1 2 1 1 1 2 10 B3 mykiss8 Salvelinus alpinus8 2 1 2 2 2 2 1 1 1 1 1 1 6 C0

Coregonus albula11 2 2 2 2 2 1 1 1 1 1 1 1 7 C1

Umbra pygmaea8 2 3 2 2 dd dd 1 1 1 1 1 1 8 C2 na: not applicable; dd: deficient data; Numbers in italics indicate scores determined using expert knowledge alone; 1: Belgian Forum on Invasive Species (BFIS) list category - A: high environmental hazard (black list); B: moderate environmental hazard (watch list); C: low environmental hazard (unclassified); 0: absent from the Netherlands; 1: isolated populations in the Netherlands; 2: restricted range in the Netherlands; 3: widespread in the Netherlands; 2: Matthews et al. (2013b); 3: Matthews et al. (2012c); 4: Koopman et al. (2014); 5: Matthews et al. (2012a); 6: De Hoop et al. (2015); 7: Collas et al. (2017); 8: Schiphouwer et al. (2014); 9: Based on a single study containing correlation evidence linking nutrient enrichment with the species; 10: Has some impact but this is not large enough for the species to be categorised under medium risk; 11: Schiphouwer et al. (2014), M. Schiphouwer (personal communication).

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C. caroliniana and D. rostriformis bugensis were the only species that received a maximum overall risk score. Eight species were assessed as moderate risk. These were a plant species: L. major, a mollusc: B. chinensis, and six fish species: Cyprinus carpio X Carassius sp. (cross carp), S. fontinalis, Romanogobio belingi (northern whitefin gudgeon), C. idella, Leuciscus aspius (asp) and O. mykiss. Five species were considered low risk, two plant species: E. densa and Vallisneria spiralis (tapegrass), and three fish species Salvelinus alpinus (Arctic charr), Coregonus albula (vendace) and U. pygmaea.

4.4.1 High risk fish species

The highest scoring fish species, C. gibelio, C. carpio and S. lucioperca all scored maximally for both dispersion potential and colonisation of high conservation value habitats (Table 4.2). C. gibelio and S. lucioperca scored maximally for the category direct or indirect adverse impacts on native species due to their high impacts relating to the sub- categories predation / herbivory, and interference or exploitation competition. C. carpio was the only high risk species to score maximally for the category direct or indirect alteration of ecosystem functions. All sub-categories, i.e., modification of nutrient cycling or resource pools, physical modifications of habitat, modification to natural succession and disruption to food webs, were scored maximally for this species.

4.4.2 High risk molluscs

D. rostriformis bugensis was classified as high risk in all four categories of the ISEIA protocol (dispersal potential and invasibility, colonisation of habitats with a high conservation value, direct and indirect effects on native species, and direct and indirect effects on ecosystem functioning; Table 4.2). All sub-categories were scored maximally apart from one where there was insufficient data (transmission of disease to native species: parasites, pest organisms or pathogens) and a second that was not considered applicable as there are no native species that are likely to interbreed with D. rostriformis bugensis (genetic effects such as hybridisation or introgression with native species). Of all the aquatic species risk assessed using the ISEIA protocol, D. rostriformis bugensis was one of only two species that received a maximum overall score of 12 for ecological impact in the Netherlands.

4.4.3 Comparison of risk classifications

Only two out of the 14 species compared were classified consistently across all regions (C. gibelio and D. rostriformis bugensis). Risk classifications for the Netherlands were either the same or lower than other available classifications for climatically similar regions in all but three cases (S. lucioperca and C. caroliniana and B. chinensis) (Tables 4.3a,b). Classification consistency of the most frequently assessed species (four or more

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classifications) was ranked C. gibelio and D. rostriformis bugensis > C. caroliniana and C. idella > O. mykiss and S. lucioperca > U. pygmaea. U. pygmaea has been assessed in Belgium twice using different protocols and received different risk classifications. All species assessed in the United Kingdom were attributed an equal or higher risk classification than classifications from other countries.

Table 4.3a: Comparison of risk classifications of aquatic plants and molluscs for the Netherlands and regions with similar climate and biogeography. Species Location Protocol Classification Reference Cabomba The Netherlands ISEIA1 High This study caroliniana Belgium ISEIA1 Medium Baus et al. (2009) New York New York Invasiveness High New York State, USA Rank System Invasive Species Information (2017) United Modified Australian Weed High Thomas (2010b) Kingdom Risk Assessment Egeria densa The Netherlands ISEIA1 Low This study Belgium ISEIA1 High Branquart et al. (2007) New York New York Invasiveness High New York State, USA Rank System Invasive Species Information (2017) Lagarosiphon The Netherlands ISEIA1 Medium This study major United Modified Australian Weed High Thomas (2010b) Kingdom Risk Assessment

Vallisneria The Netherlands ISEIA1 Low This study spiralis United Modified Australian Weed High Thomas (2010b) Kingdom Risk Assessment

Bellamya The Netherlands ISEIA1 Medium Breedveld (2015) chinensis The Netherlands Harmonia+ Medium Collas et al. (2017) New York New York Invasiveness High New York State, USA Rank System Invasive Species Information (2017) Dreissena The Netherlands ISEIA1 High This study rostriformis The Netherlands Harmonia+ High De Hoop et al. bugensis (2015) United Rapid risk assessment and High Roy et al. (2014b) Kingdom consensus method (Horizon-scan) New York New York Invasiveness High New York State, USA Rank System Invasive Species Information (2017) 1: Invasive Species Environmental Impact Assessment.

4.4.4 Statistical analysis

United Kingdom based risk classifications were found to be significantly higher compared to ISEIA classifications from the Netherlands for the same species pool, χ2(2, N = 11) = 10.27, P = 0.006. Risk classifications derived using the New York Invasiveness Ranking System were not found to be significantly higher compared to ISEIA classifications from

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the Netherlands for the same species pool, χ2(2, N = 8) = 1.6, P = 0.45. Similarly, risk classifications derived using the FISK protocol in the United Kingdom yielded significantly higher scores for the same species pool compared to classifications derived using the ISEIA protocol in the Netherlands, χ2(2, N = 7) = 7.78, P = 0.02.

Table 4.3b: Comparison of risk classifications of fish species for the Netherlands and regions with similar climate and biogeography. Species Location Protocol Classification Reference Carassius gibelio The Netherlands ISEIA1 High This study Belgium FISK2 High Verreycken et al. (2009)

2 Belgium ISEIA High Anseeuw et al. (2007a) United Kingdom FISK2 High Copp et al. (2009) Ctenopharyngodon The Netherlands ISEIA1 Medium This study idella Germany GABLIS3 High Nehring et al. (2010a)

United Kingdom FISK2 High Copp et al. (2009) New York State, New York Invasiveness High New York Invasive Species USA Rank System Information (2017) Cyprinus carpio The Netherlands ISEIA1 High This study New York State, New York Invasiveness High New York Invasive Species USA Rank System Information (2017) Leuciscus aspius The Netherlands ISEIA1 Medium This study United Kingdom FISK2 High Copp et al. (2009) Oncorhynchus The Netherlands ISEIA1 Medium This study mykiss Germany GABLIS3 High Nehring et al. (2010a)

New York State, New York Invasiveness Medium New York Invasive Species USA Rank System Information (2017) United Kingdom FISK2 High Copp et al. (2005) Salvelinus fontinalis The Netherlands ISEIA1 Medium This study Germany GABLIS3 High Nehring et al. (2010a) United Kingdom FISK2 High Copp et al. (2005) Sander lucioperca The Netherlands ISEIA1 High This study Belgium ISEIA1 Medium Anseeuw et al. (2007b)

United Kingdom FISK2 High Copp et al. (2009) New York State, New York Invasiveness Medium New York Invasive Species USA Rank System Information (2017) Umbra pygmaea The Netherlands ISEIA1 Low This study Belgium ISEIA1 Low Anseeuw et al. (2007c)

2 Belgium FISK Medium Verreycken et al. (2010) Germany GABLIS3 Low Nehring et al. (2010a) United Kingdom FISK2 High Copp et al. (2009) 1: Invasive Species Environmental Impact Assessment; 2: Fish Invasiveness Screening Kit; 3: German-Austrian Black List Information System.

4.5 Discussion

This study presents the results of risk assessments using the ISEIA protocol of 18 aquatic species alien to the Netherlands, and highlights inconsistencies with risk classifications between climatically similar regions. We chose to analyse aquatic species, however, the

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conclusions drawn here apply to a wide range of risk assessment methods and habitat types. In this section methodological differences and other sources of uncertainty which may cause the observed inconsistencies are discussed. Many of the uncertainties discussed are generic and apply to a wide range of risk assessment methods.

4.5.1 High risk species

Of the most frequently assessed species considered in the present study, a fish species, C. gibelio, and a bivalve mollusc, D. rostriformis bugensis were most frequently allocated the highest risk scores in the Netherlands, and in other regions. The high risk scores of C. gibelio may be attributed to its high dispersal potential and ability to colonise natural habitats which has led to a decline in native cyprinid species, invertebrates and plants, and potential hybridisation with the threatened crucian carp (Carassius carassius) (Paschos et al. 2004; Lusková et al. 2010; Lenhardt et al. 2010; Perdikaris et al. 2012). D. rostriformis bugensis has a strong reproductive potential and spreads via hydrochory or facilitated by human vectors such as watercraft. Due to a highly efficient filtering capacity, high densities of D. rostriformis bugensis exert significant influence on the integrity of the ecosystem by affecting biotic factors (e.g., decrease in algal biomass) and abiotic factors (e.g., increase in transparency and an accumulation of benthic organic matter in the form of (pseudo)faeces) (De Hoop et al. 2015).

4.5.2 Consistency of risk classifications

The need for a consistent assessment of risk is emphasised in the recent EU regulation on the prevention and management of the introduction and spread of IAS which demands the standardisation of risk assessment approaches across the EU (European Commission 2013, 2014). Despite differences in methodological approaches and species groups analysed, ecological risk assessments should result in classifications that are broadly similar for similar biogeographical and climate regions, and the same species. However, risk classifications from climatically similar regions generally showed a low level of consistency with risk classifications for the Netherlands. Only risk classifications for the fish C. gibelio and the mollusc D. rostriformis bugensis were consistent. Moreover, systematic inconsistency was observed between risk classifications obtained from the United Kingdom and risk classifications from other countries (consistently equal or higher risk in all cases), and from the Netherlands and other countries (consistently equal or lower risk in the majority of cases). Only in the cases of S. lucioperca and C. caroliniana was the risk classification for the Netherlands higher than that of other countries. These observations are supported by Verbrugge et al. (2012a) who showed that risk assessments of the same species but undertaken in different European countries and with different protocols resulted in differing risk classifications for 18 out of 25 species reviewed. Inconsistencies in risk classifications for the same species and region but for different

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assessment methods have been observed by a number of authors. Křivánek & Pyšek (2006) used three previously developed risk assessment schemes to categorize 180 alien woody species commonly planted in the Czech Republic to assess their consistency when applied in Central Europe: the Australian weed risk assessment scheme (WRA); the WRA with additional analysis by Daehler et al. (2004); and the decision tree scheme of Reichard & Hamilton (1997) developed in North America. The WRA+Daehler model produced results that were most consistent (85.5% consistency) with classifications of invasiveness defined in Richardson et al. (2000) and Pysek et al. (2004). Reichard and Hamilton’s decision tree model produced the least consistent results (61.6% consistency). Copp et al. (2009) compared classifications based on expert opinion from Fishbase (Froese & Pauly 2009), a global information system on fish, with the results of risk analyses of the same fish species using the Fish Invasiveness Screening Kit (FISK) applied to the United Kingdom. Out of 60 species classified as high risk by FISK, 39 were categorized on FishBase as harmless, with the remaining 21 categorized as potential pests (Copp et al. 2009).

Risk classifications from neighbouring countries may not always produce consistent results, even when the same methodology is applied. Nehring et al. (2010a) provided a risk prioritised list of 31 alien fish species for Germany. In total, 30 of these species were also selected for a list for Austria (Nehring et al. 2010a). Both lists were assessed using GABLIS. In total, 90% of species were classified equally. However, three species were classified differently which means that 10% of these fish species classifications are not transferable between neighbouring countries. Moreover, both Verreycken et al. (2010) and Verbrugge et al. (2010) observed that risk classifications generated by FISK for alien fish species in Belgium were consistently lower than mean United Kingdom scores for the same species. Leuven et al. (2016) analysed the results of risk analyses of twenty alien species (aquatic plants, molluscs and fish) carried out by different risk assessors who applied the same protocol, in a dissimilar biogeographical setting and with access to the same information. The resulting risk classifications varied between 50 and 90% for various assessment criteria (Leuven et al. 2016).

4.5.3 Potential sources of inconsistencies between risk classifications

The large number of variables included in the comparison of risk classifications precludes the direct attribution of a single determining factor for the inconsistencies observed in this study. Differences in classifications may be related to the number and differences between criteria, and the way in which they are applied. Firstly, risk classifications are usually determined and applied according to political borders and not biogeographical boundaries. For example, risk classifications derived for Belgium using the ISEIA protocol incorporate both the Belgian Atlantic and continental biogeographical regions. The Harmonia+ protocol has been promoted as a method suitable for assessing alien species risk on the EU scale (D’hondt et al. 2015). However, the EU spans a wide range of different biogeographical

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regions. Risk classifications derived using the GABLIS methodology apply to both Germany and Austria, countries that span the Atlantic, continental and Alpine biogeographical regions. Species distributions are strongly determined by biogeographical characteristics (Wallace 1876; Lomolino et al. 2006; Kreft & Jetz 2010). Therefore, a single risk classification is unlikely to be consistently accurate across an assessment area containing multiple biogeographical regions. Secondly, different assessment methods may apply different types of data when deriving risk classifications. GABLIS allocates species to particular categories according to the quality of available information as well as the potential invasiveness of the species (Essl et al. 2011). ISEIA considers only ecological impacts, whereas assessment schemes such as Harmonia+, FISK and the New York method consider ecological, economic and social aspects (Copp et al. 2005; D’hondt et al. 2015; New York Invasive Species Information 2017). Thirdly, different rating systems applied by assessment schemes may influence risk classifications due to the application of normative cut-off thresholds. Small changes in assessment, e.g., slight differences in the interpretation of the same information by different assessors, or differences in the number of cut-off thresholds, may lead to different risk classifications (Verbrugge et al. 2012a). The New York method applies insignificant, low, moderate, high and very high risk categories, whereas other assessment methods, such as ISEIA, apply low, medium and high risk categories. The application of weighting factors may also lead to different outcomes. The Harmonia+ protocol allows the application of individual weighting factors to individual questions, among modules and impact types (D’hondt et al. 2015). Finally, assessment schemes that are specific to particular taxonomic groups (e.g., FISK) may more accurately assess potential species invasiveness compared to generic assessments such as ISEIA.

Blackburn et al. (2014) recently proposed a tool that defines the impacts of alien species according to five sequential categories. In ascending order of impact, these categories are Minimal (ML), Minor (MI), Moderate (MO), Major (MR), and Massive (MA). Impacts are classified based on the level of biological organisation affected, i.e., individuals, populations, communities (reversible), and communities (irreversible), setting this methodology apart from other existing risk assessment methods such as Harmonia+, the New York method, GABLIS, FISK and ISEIA. Similar to many of the protocols considered here, the tool proposed by Blackburn et al. (2014) does not consider economic or societal impacts or ecosystem services which, according to Roy et al. (2014b), are elements required to achieve the minimum standards for risk assessments in Europe. The tool proposed by Blackburn et al. (2014) was implemented by Hawkins et al. (2015) within the Environmental Impact Classification for Alien Taxa (EICAT). To attribute a risk classification, EICAT requires information on the current impact as well as the potential maximum impact level that may be caused by a particular taxon.

Other sources of uncertainty intrinsic to the application of risk assessment protocols relate to risk perception, species environment matches or invasion histories that vary between

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countries, and variability between assessors. Uncertainty relating to qualitative and semi- quantitative approaches were defined by Leung et al. (2012): (1) linguistic uncertainty, (2) stochasticity (also referred to as irreducible uncertainty or natural variation), and (3) epistemic uncertainty (also referred to as reducible uncertainty or incertitude). Linguistic uncertainty occurs because verbal and written communication is frequently open to interpretation, and even exact language may be deciphered in different ways by different assessors (Verbrugge et al. 2016). A lack of unifying definitions within invasion biology potentially magnifies these issues (Verbrugge et al. 2016). Linguistic uncertainty may be increased when risk assessments are applied internationally due to the requirement of a common language that may be non-native to many contributors. For example, in section 5.1 of the ISEIA protocol ‘dispersal potential and invasiveness’, the criteria for medium risk applies vague terminology such as ‘remote places’, and ‘rarely exceeds’ leaving assessors to quantify the terms ‘remote’ and ‘rarely’ so that a score may be applied. A German assessor may interpret the term remote on a different spatial scale than an assessor in a relatively small country like the Netherlands, making comparisons difficult.

Stochasticity results from temporal and spatial variability, and the probabilistic mechanisms that originate from this variability. For example, the use of cases that are similar in ecological or geographical circumstances when direct evidence appears lacking, as advocated by D’hondt et al. (2015) may be problematic when considering that several species are known to expand to other habitat types once outside their native range (Wittenberg & Cock 2001; Verbrugge et al. 2012a). There is no universal explanation of successful invasion of alien species into native communities (Dawson et al. 2015), and only limited research is available on factors that determine invasion success such as particular species traits or combinations of traits (e.g., Moravcová et al. 2015; Pyšek et al. 2015).

Epistemic uncertainty reflects the level of available knowledge and its reliability related to sample size, surrogate measurements or observer error (Leung et al. 2012). Semi- quantitative risk assessments of alien species are particularly vulnerable to data error relating to observational gaps (e.g., Gassó et al. 2010; Verbrugge et al. 2012a). Vilà et al. (2009) state that of the more than 10,000 European alien species registered in the Delivering Alien Invasive Species Inventories for Europe (DAISIE) database, ecological impacts are only documented for 1094 species (11%) and economic impacts for only 1347 species (13%). The establishment of alien species does not automatically lead to negative impacts, however, a lack of recorded impacts may also be due to a lack of observation. Data gaps lead to a heavy reliance on expert opinions and interpretations (Maguire 2004; Strubbe et al. 2011; Leung et al. 2012; Verbrugge et al. 2012a). The ISEIA protocol applies strict criteria to define risk categories which may result in greater instances of data deficiency. For example, in the section ‘dispersion potential or invasiveness’, the criterion for medium risk includes the sentence: ‘natural dispersal rarely exceeds more than 1 km per year’. Seldom does information describing alien species contain such specific data on dispersal

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and terms such as ‘rarely’ are often not defined resulting in the frequent application of expert judgement with a resulting increase in epistemic uncertainty. Moreover, risk prioritization and assessment methods often do not include an assessment of impacts relating to diseases and pathogens carried by some alien species that may result in additional, severe impacts within the alien range. The transfer of diseases and pathogens to native species increases morbidity and mortality, thereby reducing their ability to compete. For example, crayfish pest (Aphanomyces astaci) is a fungal pathogen that was introduced to Europe with American crayfish species or through water transport. The disease spreads more quickly than its resistant American host, leading to widespread and severe impacts on European crayfish species, which at the same time paves the way for widespread establishment of the American species in vacated habitats (Spitzy 1971; Müller 1973). Another example is the introduction and spread of rosette agent (Sphaerothecum destruens), a disease which is associated with invasive topmouth gudgeon (Pseudorasbora parva). S. destruens is spread through invasion by P. parva to native cyprinids which are, in contrast to P. parva, not resistant to the disease (Gozlan et al. 2005, 2009; Spikmans et al. 2013).

4.5.4 The precautionary principle

The precautionary principle, or applying the worst-case scenario when different scenarios are possible, is advocated in ISEIA (Branquart et al. 2009). However, the application of the precautionary principle together with potential linguistic biases may encourage assessors to select information that portrays alien species in the worst possible light. A reduction in IAS management effectiveness may result if the same financial budget is applied to the increased number of species classified as high risk. Therefore, it is vital that the application of the precautionary principle is accompanied by an awareness of the potential implications to uncertainty and reduction in discriminatory power.

4.5.5 Conclusions

The semi-quantitative methods for the risk assessment of alien species examined here convert what is frequently qualitative data to a quantitative value to enable the calculation of a final risk score and the determination of a risk classification. If not transparently applied, semi-quantitative approaches hide methodological differences and the complexity and potentially variable quality of supporting information may obscure the application of expert opinion, and may give a false impression of legitimacy. However, the inclusion of normative aspects in the valuation of ecological effects in qualitative risk assessments of non-native species is unavoidable. Therefore, approaches that increase transparency by highlighting uncertainty are vital to the legitimacy of any assessment method. An awareness on the part of assessors of the influence of methodological difference and sources of uncertainty that accompany a particular assessment method, and the

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communication of uncertainties during reporting, will contribute to an increased acceptance of risk classifications by decision makers, nature managers and stakeholders.

The risk classifications examined here were determined for areas defined politically (within national borders) rather than ecologically (per biogeographical region). This potentially reduces the accuracy of risk classifications if the area assessed falls within multiple biogeographical regions. The application of assessment methodologies to biogeographical rather than political regions or the inclusion of biogeographical differences within current assessment approaches should be considered in an effort to increase the accuracy of risk classifications and provide a basis for more targeted and cost-effective management interventions.

Acknowledgements

The authors would like to thank the following contributors to the risk analyses described in this article: M. Dorenbosch (Bureau Waardenburg); L.M. Dionisio Pires (Deltares); M.C.M. Bruijs (DNV-GL); R. Beringen, B. Odé, L.B. Sparrius (FLORON); A. Gittenberger (GiMaRIS); H. Verreycken (Instituut Natuur- en Bosonderzoek); B.H.J.M. Crombaghs, S. van de Koppel and N. van Kessel (Natuur Balans-Limes Divergens); J.W. Lammers and J. Leferink (The Netherlands Food and Consumer Product Safety Authority); J.L.C.H. van Valkenburg (Plant Protection Service); J. Kranenbarg, M. Schiphouwer and R. Zollinger (RAVON); L.P.M. Lamers, H.J.R. Lenders and L.N.H. Verbrugge (Radboud University, Nijmegen); R. Pot (Roelf Pot research and consultancy); H.H. van Kleef (Stichting Bargerveen); J.A. Vonk (University of Amsterdam); O.L.M. Haenen and L.A.J. Nagelkerke (Wageningen University). Additionally, we would like to thank the Invasive Alien Species Team (TIE) and the Office for Risk Assessment and Research programming (BuRO) of the Netherlands Food and Consumer Product Safety Authority (NVWA) who commissioned and provided financial support for the risk assessments included.

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5 Rapid range expansion of the invasive quagga mussel in relation to zebra mussel presence in the Netherlands and Western Europe

Jonathan Matthews Gerard van der Velde Abraham bij de Vaate Frank P.L. Collas K. Remon Koopman Rob S.E.W. Leuven

Biological invasions 16: 23-42. January 2014.

The invasive quagga mussel in the Netherlands and Western Europe

5.1 Abstract

Since its appearance in 2006 in a fresh- water section of the Rhine-Meuse (Rijn-Maas) estuary (Hollandsch Diep, the Netherlands), the non-indigenous quagga mussel has displayed a rapid range expansion in Western Europe. However, an overview characterising the spread and impacts of the quagga mussel in this area is currently lacking. A literature study, supplemented with field data, was performed to gather all available data and information relating to quagga mussel dispersal. Dispersal characteristics were analysed for rate and direction and in relation to hydrological connectivity and dispersal vectors. To determine ranges of conditions suitable for quagga mussel colonisation, physico-chemical characteristics of their habitats were analysed. After its initial arrival in the freshwater section of the Rhine-Meuse estuary and River Danube, the quagga mussel demonstrated a rapid and continued range expansion in Western Europe. Quagga mussels have extended their non- native range to the network of major waterways in the Netherlands and in an upstream direction in the River Rhine (Germany), its tributaries (rivers Main and Moselle) and the River Meuse (Belgium and France). The calculated average quagga mussel dispersal rate in Europe was 120 km year-1 (range 23- 383 km year-1). Hydrological connectivity is important in determining the speed with which colonisation occurs. Dispersal to water bodies disconnected from the freshwater network requires the presence of a suitable vector e.g. pleasure boats transferred over land. Upstream dispersal is primarily human mediated through the attachment of mussels to watercraft. The relative abundance of quagga mussel to zebra mussel has greatly increased in a number of areas sampled in the major Dutch rivers and lakes and the rivers Main and Rhine and the Rhine-Danube Canal leading to a dominance shift from zebra mussels to quagga mussels. However, evidence for displacement of the zebra mussel is limited due to the lack of temporal trends relating to the overall density of zebra and quagga mussels.

5.2 Introduction

The recent appearance and rapid range expansion of the non-indigenous quagga mussel, Dreissena rostriformis bugensis Andrusov, 1897 in Western Europe highlights the need for assessment of its colonisation and dispersal mechanisms in European freshwater ecosystems. Until the 1940s, the quagga mussel only occurred in the mouth of two rivers discharging into the Black Sea, viz. the Southern Bug and Dnieper in the Ukraine (Son 2007; Van der Velde et al. 2010b). Since then it has extended its range in Russia considerably, reaching the River Volga via the River Don and the Volga-Don Canal (Orlova et al. 2001, 2004; Son 2007; Zhulidov et al. 2004, 2005). The quagga mussel was first recorded in North America in 1991, possibly arriving simultaneously with the zebra mussel, Dreissena polymorpha (Pallas, 1771) around 1986. This introduction probably occurred at the larval stage via ballast water discharge (May & Marsden 1992). By 2004 the quagga mussel had expanded its range to the Romanian section of the River Danube,

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Eastern Europe (Micu & Telembici 2004; Popa & Popa 2006). The first observation in Western Europe was made in 2006 in the Hollandsch Diep, a dammed, freshwater section of the Rhine and Meuse estuary in the Netherlands (Molloy et al. 2007). Subsequently, the species spread rapidly and it is now classified as invasive in Western Europe (Matthews et al. 2012d). Certain records indicate that quagga mussel density is higher than that of the earlier invader, the zebra mussel (Bonhof et al. 2009; De Rooij et al. 2009; Bij de Vaate 2010a; Matthews et al. 2012d; Heiler et al. 2012).

Besides exhibiting a planktonic larval stage, unique for dreissenids among freshwater bivalves in Europe and North America, the adults secrete proteinaceous byssal threads allowing attachment to hard substrata (Bonner & Rockhill 1994; Clarke & McMahon 1996). Dreissenids byssally attach to vectors e.g. recreational boats or shipping and are subsequently transported upstream or overland to hydrologically disconnected, un-infested water bodies (Johnson & Carlton 1996).

A review exclusively based on North American studies, identified excretion rate and heavy metal accumulation as the only impacts that differed between quagga and zebra mussels (Kelly et al. 2010). Dreissenids negatively impact unionid mussels and certain fish populations, alter macroinvertebrate composition, reduce plankton abundance and increase macrophyte and benthic algal growth (Kelly et al. 2010). However, dreissenids provide food for certain water birds, fish, zoobenthic detritivores, crayfish and crabs. Moreover, increase in benthic invertebrate abundance after dreissenid invasion benefits invertebrate predators (Kelly et al. 2010; Mitchell et al. 1996; Mörtl et al. 2010; Van Eerden & De Leeuw 2010).

An overview characterising the spread and impacts of the quagga mussel in Western Europe is currently lacking. Additionally, insight into controllable environmental factors that influence quagga mussel distribution and population establishment is required. This paper characterises the current spread of the quagga mussel in the Netherlands and Western Europe and describes dispersal mechanisms and factors that influence quagga mussel distribution. Additionally, dominance shifts due to quagga mussel colonization within established zebra mussel habitats are analysed.

5.3 Methods

5.3.1 Data acquisition and analyses of current distribution

A literature study was carried out to derive information on the current distribution, colonisation vectors and dispersal mechanisms of dreissenids in the Netherlands and Western Europe. Search terms included the Latin names: Dreissena bugensis and D. rostriformis bugensis which produced search results written in languages other than English

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and Dutch. English terms and their Dutch equivalents included: quagga, zebra, exotic, dispersal, colonisation, abundance, invasive, connectivity, vectors, physiological, tolerance, replacement, demographics, population, dynamics, Netherlands. Scientific papers, (grey) publications and websites were searched systematically, ensuring the quality of the data obtained, using several academic search engines (the Zoological Record, Web of Knowledge, Scopus, Google Scholar, www.science.gov, SCIRUS, Grey Literature in the Netherlands (GLIN), The Dutch Central Catalogue (PiCarta) and Greynet). Records taken from grey literature were only included if they were associated with a named specialist and/or governmental or non-governmental organisation.

Field surveys were carried out in the Netherlands with supplementary data obtained from co-authors who collected samples in Belgium, Germany and France. In the Netherlands, previously unexamined water bodies were identified for sampling. Sites were chosen to increase the representativeness of different levels of hydrological connectivity and water types. Sites were added at previously sampled locations to explore trends in abundance and species replacement. Sampling in the Netherlands occurred between March and September in the years 2011 (59 samples) and 2012 (26 samples) at different locations, primarily in the littoral zones of large rivers, tributaries, canals and lakes (Fig. 5.1).

Figure 5.1: Locations sampled during field surveys in the Netherlands.

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The major substrata (soft substratum, stone and vegetation) were sampled. If stones were sampled, five were removed at random along a 10-20 m stretch from depths of up to 75 cm. Groynes were sampled from their downstream side. All mussels were removed from the most densely colonised side of each stone. Five samples were scraped from vegetation, soft substrates, ship hulls and immovable substrates using a sieve or dip net. Mussel samples were preserved in ethanol (70%), identified in the laboratory and analysed for relative abundance and density. The water-type (river, stream, canal etc.), colonisation vectors present (shipping, driftwood etc.) were recorded. Data on mussel abundance, sampling location, sampling methodology and sampling date derived from the literature study and field notes were entered into a database.

5.3.2 Analysis of dispersal rate and vectors

Dispersal rates (km year-1) were calculated using the shortest distance of continuous waterway located between the point of initial quagga mussel colonisation in the Netherlands (Molloy et al. 2007; Bij de Vaate 2006), and the farthest known Dutch records to the north, south and east and the time delay occurring between the establishment of these records. This method has been used previously by Voelz et al. (1998), Leuven et al. (2009) and Kappes & Haase (2012). Arrival time was determined using a calculation of maximum age, derived from shell length, where sample populations were divided into length classes that represented age classes: 0+, 1+ and 2+ years at an average length of 5.6, 16.1 and 25.6 mm, respectively (Bij de Vaate 2008). The rates obtained were averaged and compared to rates for Western Europe found in literature. Dispersal maps illustrating a different time period during quagga mussel colonisation in the Netherlands were created using the mapping programme Stipt (Frigge 2012). The yearly cumulative totals of Dutch quagga mussel locations were calculated. Vector types and modes of dispersal (active, passive, upstream, downstream, overland) used by quagga mussels were identified from literature and during field sampling.

5.3.3 Distribution in relation to connectivity and physico-chemical characteristics of water bodies

Dutch sampling locations were divided into hydrological connectivity classes (permanently connected, disconnected and seasonally connected water bodies) to analyse the effect of hydrological connectivity on quagga dispersal. Permanently connected water bodies are located downstream of quagga populations allowing colonisation by passive drift. A disconnected water body is defined where no quagga population exists upstream of the sampled location and where colonisation is facilitated by vectors (includes land-locked water bodies). Seasonally connected water bodies are connected discontinuously, depending on water level. Sampling locations were only included in the analysis when accurate information was available detailing the hydrological connectivity of the sampled water-body. Subsequently, the number of samples containing quagga mussels within each

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connectivity class was calculated and the results compared. The same methodology was applied in a separate analysis of water-type where sampling locations were assigned to the categories of river/stream, canal, lake and harbour. For the statistical analysis, categories were defined according to the abundance of individual mussel species. Potential pseudo- replication was avoided as data used were obtained from spatially differentiated sampling locations and sampling was not repeated at the same location. Differences between the categories were tested for statistical significance using the Kruskal-Wallis test (SPSS 17, IBM, version 17.0.0). The results were judged to be significant at a level of P = <0.05.

Physiological tolerance limits obtained from literature were validated by comparing them with water quality data collected from monitoring stations coinciding with quagga mussel presence in the Netherlands. Where tolerance limits for the quagga mussel were not available, data for the zebra mussel, indicative for quagga mussel tolerance, were used. Data on calcium, cadmium, copper, lead, nitrate, dissolved oxygen, pH, salinity, temperature, total phosphate and zinc were sourced from Waterbase.nl, a validated online database of the Dutch Ministry of Infrastructure and the Environment. Data was collected from the monitoring stations: River Meuse at Eijsden (N 48° 16’ 19.86”, E 5° 40’ 59.67”) and Belfeld (N 51° 19’ 6.97”, E 6° 6’ 47.97”), River Waal at Lobith (N 51° 51’ 15.14”, E 6° 5’ 28.24”), River IJssel at Kampen (N 52° 33’ 13.74”, E 5° 55’ 26.07”), Amsterdam- Rhine Canal at Nieuwegein (N 52° 1’ 21.83”, E 5° 6’ 46.83”), Twente Canal at Eefde (N 52°9037.5400, E 6°14015.8100), Lake IJsselmeer at Vrouwezand (N 52° 48’ 37.26”, E 5° 23’ 35.30”) and Lake Markermeer (N 52° 31’ 36.09”, E 5° 13’ 9.72”). Minimum and maximum values were calculated and compared with the established physiological tolerance data. Flow velocity data were obtained from 10 sampling points within quagga mussel microhabitats in the rivers Waal, Meuse, Nederrijn and IJssel using a TAD-micro flow meter (Van Vugt Instrumentation) that measures multi-directional flow rate with a small propeller. Reference and extreme values were measured during monitoring (e.g. before and directly following the passage of large vessels).

5.3.4 Analysis of species dominance shifts

Mussel abundances and total density data were derived for quagga and zebra mussels from Dutch sampling locations and literature. The criteria for inclusion were that both species were present and that sampling had occurred repeatedly at the same location. Species replacement was said to be occurring in the presence of a positive trend in quagga relative abundance in association with a positive or neutral trend in density of both species. A reduction in overall mussel density accompanied by an increase in quagga mussel density suggests that factors unrelated to inter-specific competition may have resulted in reductions in zebra mussel density. Therefore, situations where an overall reduction in mussel density occurred were not defined as species replacement. Density trends of both species derived

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from consistent sampling approaches are necessary to conduct a sound analysis of species replacement.

In a separate analysis, Dutch sampling locations were divided into three classes: (1) only quagga mussels found, (2), only zebra mussels found, (3), both quagga and zebra mussels found. This identified the level of coexistence of quagga and zebra mussels at the sampled sites. Coexistence of mussel species suggests that quagga and zebra mussels require similar habitat and introduces the possibility that species replacement may occur when quagga mussels colonise sites shared by zebra mussels.

5.4 Results

5.4.1 Current European distribution

The first observation of the quagga mussel in Western Europe was made in 2006 in the Hollandsch Diep, a former estuary of the rivers Rhine and Meuse in the Netherlands (Bij de Vaate 2006; Molloy et al. 2007; Bij de Vaate & Jansen 2007; Schonenberg & Gittenberger 2008). Figures 5.2 and 5.3 illustrate the range expansion of the quagga mussel in the Netherlands and Western Europe. Bij de Vaate et al. (2010b, 2013) argued that quagga mussel introduction into Western Europe was not the result of a continuous range expansion through the River Danube and the Main-Danube Canal and River Rhine, as previously suggested by Molloy et al. (2007). Ballast water transport and release in the Hollandsch Diep or transport after attachment to inland shipping have been suggested as potential dispersal mechanisms (Bij de Vaate 2010b; Bij de Vaate & Beisel 2011; Bij de Vaate et al. 2013; Heiler et al. 2013). Ballast water may have originated from the Black sea area or North America in the Port of Rotterdam or Hollandsch Diep (Van der Velde et al. 2010b).

In 2007 Van der Velde & Platvoet (2007) discovered quagga mussels in the River Main (Germany), a tributary of the River Rhine, but not in the Danube near the Main-Danube canal or the Main- Danube canal itself. This led to further observations in Germany. Martens et al. (2007) discovered the species in a series of Upper Rhine harbours, while Haybach & Christmann (2009) found them in the Lower Rhine in 2008 between Dormagen and Bimmen. In 2008 quagga mussels were found in the northern part of the Main-Danube Canal (Bij de Vaate 2010b). Mayer et al. (2009) found quagga mussels on ship’s hulls on the slipway of a shipyard at Speyer, along the Upper Rhine.

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Figure 5.2: Available records demonstrating the rapid range expansion of the quagga mussel in Europe; numbers refer to locations used to calculate dispersal rates in table 5.1.

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Figure 5.3: Rapid range expansion of the quagga mussel through the waterway network after its first record in the freshwater section of the Dutch Rhine–Meuse river estuary, Hollandsch Diep in 2006; numbers refer to locations used to calculate dispersal rates in table 5.1.

Quagga mussels found near Karlsruhe an Mannheim in the River Rhine were smaller than in the River Main (Imo et al. 2010). However, genetic analysis ruled out the presence of founder effects. Based on non-continuous distribution and shell size, it was concluded that range expansion in Germany involved at least two independent settling events. The first event occurred before 2005, probably caused by jump dispersal, while the second event was due to continuous range expansion. Heiler et al. (2012) observed range expansion in Germany in an easterly direction through canals of which the Mittelland Canal is most important, connecting the river catchments of the Weser, Elbe and Oder. Bij de Vaate & Beisel (2011) are responsible for the first French record in the River Moselle, another tributary of the River Rhine. In 2009 Sablon et al. (2010) recorded the quagga mussel for the first time in Belgium (Albert Canal, in the vicinity of Grobbendonk), while Marescaux et al. (2012) observed upstream migration of the species in the Belgian section of the River Meuse beginning in 2010. This pattern of records demonstrates the rapid range extension of the quagga mussel in Western Europe after its establishment in the Netherlands.

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The invasive quagga mussel in the Netherlands and Western Europe

5.4.2 Rate and direction of colonisation in the Netherlands

Initially, quagga mussel records were limited to the Hollandsch Diep, lakes IJsselmeer and Markermeer, and the rivers Meuse and Waal. Spatial analysis shows the rapid range expansion of the quagga mussel to all the major rivers and canals in the Netherlands since 2006 (Fig. 5.3). Records exist for the rivers Nederrijn, IJssel and Bovenrijn, and large canals and lakes connected to the European network of waterways e.g. the Frisian lakes, Lake Volkerak-Zoommeer, the Pannerdensch Canal, the Amsterdam-Rhine Canal, the Rhine-Scheldt Canal, the Wilhelmina Canal, the Bathse Spui Canal and the Meuse-Waal canal (Soes 2008; Bij de Vaate & Jansen 2009, 2011; Bij de Vaate 2009, 2010a; Raad 2010; Bij de Vaate et al. 2011; Matthews et al. 2012d). Recently, quagga mussel records have extended as far south as the River Meuse at the mouth of the Juliana Canal at the Dutch-Belgian border, South Limburg and as far North as the Van Harinxma Canal to the West of Leeuwarden in Friesland and the Van Starkenborgh Canal in the Province of Groningen.

The number of quagga mussel records has been increasing in the Netherlands since 2006 (Figs. 5.3, 5.4). The cumulative number of records sharply increased between 2006 and 2008. Since 2008 the rate of increase has reduced, however the overall number of records continues to rise.

Figure 5.4: Cumulative number of locations where quagga mussels were identified in the Netherlands.

The dispersal rate for the quagga mussel in Europe was found to range between 23 and 383 km year-1 with an average of 120 ± SE 53.8 km year-1 (Table 5.1). Dispersal rates calculated for major Dutch rivers, lakes and canals varied between 23 and 105 km year-1. Rates calculated within the Netherlands were for upstream dispersal in lotic water bodies apart from the lentic water bodies Lake IJsselmeer and the Prinses Margriet Canal (Province of Friesland).

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Table 5.1: Dispersal rate of the quagga mussel in the Netherlands and Western Europe. Numbers in brackets refer to coordinates of locations identified in Figures 5.2 and 5.3.

Dispersal route Coordinates Dispersal Reference

attributes

km)

)

1

-

Distance ( Duration (years) Dispersal rate y (km

From Hollandsch Diep to the N51° 41' 34.38" 116 5 23 upstream Matthews et Lake Kaliwaal, via the River E4° 25' 57.45" (1) al. (2012d) Waal (the Netherlands) to N51° 51' 18.82" E5° 59' 17.28" (2)

From Hollandsch Diep to the N51° 41' 34.38" 265 4 66 upstream Bij de Vaate, entrance of the Juliana Canal, via E4° 25' 57.45" (1) unpublished the River Meuse* (the to data Netherlands) N50° 52' 11.10" E5° 41' 57.95" (3)

River Meuse from Sambeek (the N 51° 38' 12" 260 3 87 upstream Bij de Vaate Netherlands) to Gives (Belgium)* E 5° 59' 24" (4) et al. (2013) to N 50° 30' 33" E 5° 08' 55" (5)

From Hollandsch Diep via N51° 41' 34.38" 211 2 105 upstream and via Bij de Vaate Dordtsche Kil, River Oude E4° 25' 57.45" (1) lentic water bodies (2009) Meuse, Noord, River Lek, to (lakes) and water Amsterdam-Rhine Canal to Oude N52° 52' 9.76" bodies with low Zeug in Lake IJsselmeer (the E5° 7' 38.33" (6) flow velocity Netherlands) (canals)

From Hollandsch Diep via N51° 41' 34.38" 339 6 57 upstream and via Bij de Vaate, Dordtsche Kil, River Oude E4° 25' 57.45" (1) lentic water bodies unpublished Meuse, Noord, River Lek, to (lakes) and water data Amsterdam-Rhine Canal, Lake N53° 16' 37.7", bodies with low IJsselmeer, Prinses Margriet E6° 45' 28.4" (7) flow velocity Canal to Eemskanaal (the (canals) Netherlands)

River Danube, from Cernavodă N 44° 21' 1450 4 383 upstream Bij de Vaate (Romania) to Komarum E 28° 01' (8) et al. (2013) (Hungary)* to N 47° 44' E 18° 08' (9) * Dispersal rate determined using sampled years

5.4.3 Distribution in relation to the connectivity of water bodies, substrates and water- type

While both quagga and zebra mussels occurred at a similar frequency in permanently connected water bodies, limited hydrological connectivity reduces quagga mussel occurrence to a greater degree (Fig. 5.5). In seasonally connected water bodies, quagga mussels occurred at 20% of locations whereas the zebra mussel occurred at 40% of locations. In disconnected water bodies the quagga mussel occurred at no locations whereas the zebra mussel occurred at 35% of locations. Hydrological connectivity was found to have a statistically significant effect on the distribution of the quagga mussel (χ2 = 24.667, df = 2, P = <0.001) but not the zebra mussel (χ2 = 5.702, df = 2, P = 0.058). These

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The invasive quagga mussel in the Netherlands and Western Europe

observation must be viewed with caution, however, as only five samples were included in the analysis of seasonal connectivity.

Figure 5.5: The influence of the level of hydrological connectivity on quagga and zebra mussel occurrence in the Netherlands.

The quagga mussel was found to attach to a number of different hard substrates. These included stone, metal and concrete (Table 5.2). Moreover, quagga mussels were observed to attach to the empty shells of dead native mussels. Of the 18 soft substrate samples taken at four locations, one sample contained seven quagga mussels that were attached to each other. At no other locations were quagga mussels observed on soft substrate. Moreover, of the 15 samples taken of vegetation (3 locations), none contained specimens of either quagga or zebra mussels.

Table 5.2: Presence of quagga mussels on sampled substrates. Substrate Number of quagga mussel locations Total number of locations sampled (n) (n) Stone 21 44 Concrete 1 8 Soft (e.g. sand) 1 4 Metal 3 5 Rubber 1 1 Wood 1 8 Vegetation 0 3

Quagga and zebra mussels were present in all sampled water-types: rivers/streams, canals, lakes and harbours (Fig. 5.6). Quagga mussels occurred most frequently in canals and harbours followed by rivers/ streams and lakes. Zebra mussels occurred most frequently in

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harbours followed by lakes, rivers and streams and canals, however this trend was not statistically significant.

Figure 5.6: Presence of dreissenids species per water-type.

The largest difference in occurrence between zebra and quagga mussels was between the harbour and lake water-types. However, this result should be treated with caution as only four harbour locations were sampled. Water-type was found to have a statistically significant effect on the distribution of the quagga mussel (χ2 = 19.677, df = 3, P = <0.001), but not on the distribution of the zebra mussel (χ2 = 2.190, df = 3, P = 0.534).

5.4.4 Identification of dispersal vectors

An overview of dispersal vectors utilised by dreissenid mussels is given in Table 5.3. The quagga mussel was observed attached to driftwood and boat hulls and it was present in high abundance inside a discarded car tyre on the banks of a floodplain lake. The literature survey revealed a number of human mediated vectors that facilitate dreissenid mussel dispersal. Attachment to watercraft was the most frequently cited mode of human mediated dispersal (Keevin et al. 1992; Allen & Ramcharan 2001; Johnson et al. 2001).

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The invasive quagga mussel in the Netherlands and Western Europe

Table 5.3: Overview of dispersal vectors relevant to the spread of dreissenids.

Vector / mechanism Mode of transport Examples and relevant References information

Water-flow Downstream, lateral In the presence of upstream 1, 2 populations Foot and shell movement Active up- and Triggered by change in habitat 3 downstream, lateral Wind driven water flow Multi-directional Most relevant in lentic water bodies Vegetation, wood*, Downstream, lateral Longitudinal with water-flow and 4, 5, 6, 7 detached floats, buoys, and overland lateral with flooding or attached plastic rubbish, rubber to other vectors e.g. watercraft tyres* Commercial shipping Up- and downstream Barges 4, 8, 9, 10, 11, 12

Watercraft* Up- and downstream Recreational vessels 7, 13

Ballast water Sea going Sea going vessels 14, 15, 16 Watercraft overland Overland Canoes, sailing / rowing / motor 4, 11, 13, 17, 18, 19, boats. dependent on resistance to 20, 21, 22, 23, 24 desiccation Other overland Overland Metal pontoon, floats, entangled 7, 21, 25 vegetation Several animal species Up- and Predatory animals such as aquatic 26, 27, 28 downstream, lateral birds, muskrats and brown rats, and overland and attached to Chinese mitten crabs, crayfish and turtles 1: Johnson & Padilla (1996); 2: Stoeckel et al. (1997); 3: Kappes & Haase (2012); 4: Carlton (1993); 5: Horvarth & Lamberti (1997); 6: Wilson et al. (1999); 7: Matthews et al. (2012d); 8: Keevin et al. (1992); 9: Nehring (2002); 10: Slynko et al. (2002); 11: Bij de Vaate et al. (2002); 12: Olenin (2002); 13: Minchin et al. (2003); 14: Carlton & Geller (1993); 15: Endresen et al. (2004); 16: Holeck et al. (2004); 17: Padilla et al. (1996); 18: Buchan & Padilla (1999); 19: Kraft & Johnson (2000); 20: Allen & Ramcharan (2001); 21: Johnson et al. (2001); 22 : Kraft et al. (2002); 23 : Minchin et al. (2002); 24 : Pollux et al. (2003); 25 : Bidwell (2010); 26 : Johnson & Carlton (1996); 27 : Bossenbroek et al. (2001); 28: Karatayev et al. (2003). *: Quagga mussel dispersal vectors observed during this study.

5.4.5 Analysis of physico-chemical characteristics of colonised water bodies

Water quality data indicate that most measurements lay within the physiological limits of the quagga mussel (Table 5.4). Temperature measurements did not exceed the maximum physiological tolerance for the quagga mussel. The overall minimum temperature was 1.1 °C, however, measurements were taken at the water surface and relatively warm, deeper water may provide refuges for mussels (Leuven et al. 2011). Data for pH tolerance were only available for the zebra mussel. pH remained within physiological tolerances in lake habitats but minimum pHs measured in canals and rivers were 7.2 and 7.3, respectively, and below the minimum tolerance of 7.4. However, this minimum represents one in 24 measurements recorded in the Twente canal (2010-2011) and one in 260 measurements recorded in the river Meuse (2007-2011). pH was measured at relatively high frequencies in

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the Twente canal and river Meuse, and could only have resulted in unsuitable conditions for relatively short periods. Dissolved oxygen was measured above the minimum tolerance for zebra mussels at all locations. Peak flow velocities measured in quagga mussel microhabitats far exceeded physiological tolerances. However, these measurements were transient, resulting from turbulence created by passing ships. Moreover, at the location of maximum water velocity, only a single live quagga mussel was found. When the effect of shipping was removed, the maximum flow velocity was far lower (Table 5.4).

Table 5.4: Ranges of physico-chemical properties per water-type where the quagga mussel has colonised in the Netherlands in comparison with physiological tolerances. Physico-chemical Large Canals Lakes Combined Physiological property rivers tolerances Calcium (mg l-1) 35.0 - 92.1 39.2 - 105.0 39.7 - 99.0 35.0 - >121 105.0 Cadmium (μg l-1) 0.05 - 2.27 0.05 - 0.45 0.05 - 0.20 0.05 - 2.27 <272,*,+ Copper (μg l-1) 1.38 - 39.4 1.89 - 20.1 1.34 - 6.63 1.34 - 20.1 <432,*,+ Lead (μg l-1) 0.17 - 37.0 0.1 - 10.0 0.1 - 4.7 0.1 - 37.0 <912,*,+ Zinc (μg l-1) 6.2 - 330 2.24 - 67 1 - 37 1 - 330 <1312,*,+ Nitrate (mg l-1) 1.09 - 4.4 0.789 - 8.72 0.01 - 0.48 0.01 - 8.72 No Data Total phosphate (mg l-1) 0.04 – 1.9 0.02 - 1.3 0.02 - 0.48 0.02 - 1.9 No Data Dissolved oxygen 3.2 - 14.8 5.18 - 12.8 8.7-15.9 3.2- 15.9 >1.8–2.43* content (mg l-1) pH 7.3 - 8.48 7.2 - 8.33 8.3 - 10.4 7.2 - 10.4 7.44,5 - (9.3-9.66)*

Salinity (psu) 0.0 - 0.4 0.2 - 0.6 0.3 - 0.44 0.0 - 0.6 <57,8 Temperature (oC) 1.1 - 26.8 0.6 - 24.1 0.2 - 2 0.2 - 26.8 9 for reproduction9 – (25-3410,7,11) Flow velocity (cm s-1) 1 - 2.2$ / 3 - 7 No Data 1 – 117# <(9-2012, 13) 117# $: maximum velocity under no influence of shipping; #: One live quagga mussel found at this location, maximum velocity caused by passing shipping; *: Data only available for zebra mussel; +: 50% effect concentration for filtration, 10 week exposure; 1: Jones & Ricciardi (2005); 2: Kraak et al. (1994); 3: Karatayev et al. (2007); 4: Ramcharan et al. (1992); 5: Neary & Leach (1992); 6: Bowman & Bailey (1998); 7: Karatayev et al. (1998); 8: Spidle et al. (1995); 9: Bij de Vaate (2008); 10: Mills et al. (1996); 11: Verbrugge et al. (2012b); 12: Ackerman (1999); 13: Eckman et al. (1989).

50% effect concentrations for filtration of metals (EC50 filt) were available for the zebra mussel only. EC50 values for lead, cadmium and copper were higher than environmental concentrations found in quagga mussel habitats. A single zinc measurement in the river Meuse at Eijsden exceeded the EC50. Due to the sampling frequency, this condition could only have persisted for a maximum of 14 days within the 5 year period that quagga mussels have been recorded in the river Meuse. All other measurements for metal concentrations lay below the maximum tolerance values.

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The invasive quagga mussel in the Netherlands and Western Europe

5.4.6 Shifts to quagga mussel dominance

When ordered chronologically, samples taken from Dutch and German water bodies, where both species occurred together, indicate that the quagga mussel increased in abundance relative to the zebra mussel (Fig. 5.7). Data from 2011 indicates that the quagga population represented 95% of the dreissenid population in Lake IJsselmeer and Hollandsch Diep. However, in recent years in the rivers IJssel and Nederrijn, the quagga mussel relative abundance has reduced by as much as 66%.

Figure 5.7: Percentage contribution of quagga mussel to the overall quagga and zebra mussel population in major European water bodies (Bij de Vaate 2008, 2010a; Bij de Vaate & Jansen 2012; Heiler et al. 2012; Matthews et al. 2012d; unpublished data A. bij de Vaate).

Twelve locations were identified where temporal trends could be used to establish if species replacement was occurring, some within the same water body (Table 5.5). Zebra mussels were present in the most recent samples at all locations. In total, seven out of 12 locations lacked mussel density data or data that could be used to calculate density. Of the remaining five locations, three demonstrated an increasing trend in total mussel density and two showed a decreasing trend in total mussel density. The density ratio of quagga to zebra mussels increased in four out of the five locations where density data were available. None of the locations identified in Table 5.5 showed a decreasing trend in quagga relative abundance. Species replacement could be happening at three locations, in the River Nederrijn in the side channel at Bakenhof, the Meuse-Waal canal at Dukenburg and in the Hollandsch Diep. Eighty percent of samples containing dreissenids in the years 2011 and 2012 included both mussel species. Only 15 and 5% of locations contained only zebra and quagga mussels, respectively.

88

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3 2 1 89

The invasive quagga mussel in the Netherlands and Western Europe

5.5 Discussion

The number of quagga mussel locations in the Netherlands has increased rapidly since it was first identified in the Rhine-Meuse estuary in 2006 (Fig. 5.3). Dispersal in an easterly and northerly direction rules out the possibility of naturally occurring modes of dispersal, e.g. downstream drift in the River Rhine, and supports the hypothesis that initial establishment of quagga in the Netherlands occurred in the Hollandsch Diep (Molloy et al. 2007; Bij de Vaate 2006). In the absence of upstream source populations, colonisation via jump dispersal facilitated primarily by watercraft must occur first, followed by passive ‘natural’ dispersal on the water current originating from the newly established upstream population. A number of authors have emphasised the primary importance of watercraft as a vector for dreissenid dispersal (Keevin et al. 1992; Allen & Ramcharan 2001; Johnson et al. 2001; Heiler et al. 2013). During our field research, quagga mussels were found attached to ship hulls. This combined with the presence of quagga mussels in navigable water bodies suggests that watercraft may be the primary mode of dispersal for quagga mussels in the Netherlands. Inland vessels do not feature ballast tanks, however, they often feature equipment for hull decontamination that contains residual water that may provide a vector for mussel larval transfer. Moreover, sludge that may contain mussels is transferred via inland waterways by dredgers. Passive ‘natural’ dispersal of larvae is largely accomplished by movement of water, including lake currents driven by wind and water density differences and/or unidirectional flow associated with rivers, streams and slow flowing canals (Johnson & Padilla 1996; Stoeckel et al. 1997). Moreover, mussels have the ability to move over short distances using foot and shell movement. Active upstream movement may be well below 0.1 km year-1 for bivalves, however (Kappes & Haase 2012). Attachment to floating debris such as plant material, plastic and wood facilitates the transport of adult mussels downstream (Pollux et al. 2010). In the Netherlands, various modes of dispersal allow the quagga mussel to disperse rapidly. Quagga mussels are able to attach to wood, rubber and ship hulls. In Western Europe, available records for the quagga mussel demonstrate its rapid expansion into the rivers Main and Rhine in Germany, the River Meuse and the Albert canal in Belgium and the French section of the River Moselle (Bij de Vaate & Beisel 2011; Imo et al. 2010; Sablon et al. 2010; Van der Velde & Platvoet 2007).

Calculated dispersal rates in the Netherlands were lower than those in the Danube (Table 5.1). Dispersal occurred in a primarily upstream direction and shipping is likely to be a major mode of dispersal. The Danube dispersal rate was calculated over a significantly longer distance (between Romania, Germany and Austria) than those calculated within the Netherlands. The length of the dispersal route and mode of dispersal may influence the calculated dispersal rate. Commercial shipping can traverse large distances within a few days. Theoretically, quagga mussels may begin colonisation 1,000s of kilometres away from the original colonisation source after a single ship passage. Moreover, dispersal rates may be underestimated due to the time lag between dispersal, establishment of populations

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and the first recorded observation of each non-indigenous species (Leuven et al. 2009). However, other authors have used a similar method to calculate dispersal rates (Voelz et al. 1998; Leuven et al. 2009; Kappes & Haase 2012) and the extensive governmental sampling network and large number of amateur observers in the Netherlands increases the probability that quagga mussels will be recorded close to their time of arrival.

Quagga and zebra mussels use similar vectors of dispersal (Table 5.3). Dispersal of both dreissenid mussels is hindered by a lack of hydrological connectivity (Fig. 5.5), and facilitated by dispersal vectors such as watercraft, moveable pontoons and floats. Samples taken from the tributaries of large rivers and hydrologically disconnected water bodies in the Netherlands contained no quagga mussels, however, zebra mussels were sometimes present. The delayed dispersal of the quagga mussel to disconnected water bodies may be explained by the time needed for source populations to develop in downstream reaches. The quagga mussel was first recorded in Western Europe in 2006 whereas the zebra mussel colonised at least 200 years ago also with delayed initial dispersal to more isolated water bodies (Bidwell 2010; Van der Velde et al. 2010a).

Transient recreational boating is commonly perceived as the primary means by which dreissenids are transported between disconnected water bodies (Johnson et al. 2001). However, a relative intolerance to drying (Ricciardi et al. 1995; Allen & Ramcharan 2001) may reduce the probability of quagga dispersal overland. Macrophytes, however, may provide a refuge against desiccation. Attachment to macrophytes has been implicated in the spread of zebra mussels by overland transport on boats (Wilson et al. 1999; Johnson et al. 2001). However, quagga mussels exhibit a lesser ability to attach to macrophytes than zebra mussels (Diggins et al. 2004) and no quagga mussels were found attached to macrophytes during our field surveys (Table 5.2). Moreover, higher byssal thread synthesis rate in zebra mussels will likely minimize their dislodgment in flow and increase short-term attachment strength (Peyer et al. 2009; Grutters et al. 2012). However, another study found no significant difference in attachment strength measured between dreissenid species (Ackerman et al. 1995). As both the quagga and zebra mussel use byssal threads to attach to dispersal vectors and despite issues relating to desiccation and attachment strength, it may only be a matter of time before quagga mussels colonise hydrologically isolated water bodies.

The analysis of environmental conditions indicates that physiological tolerance ranges obtained from literature are representative of quagga mussel populations found in the Netherlands (Table 5.4). Tolerances relating to nutrients and feeding conditions were not found in the literature, however, the quagga mussel is generally more tolerant of high silt levels than the zebra mussel and prefers more oligotrophic conditions (Karatayev et al. 1998; Orlova et al. 2005). Total phosphorus and nitrate data ranges give examples of conditions suitable for quagga mussel colonisation as nutrient levels will impact on

91

The invasive quagga mussel in the Netherlands and Western Europe

plankton and seston food sources and mussel filter feeding. In experiments examining the impact of temperature, salinity and light on substrate attachment, it was found that both species were similar in their initial re-attachment ability (Grutters et al. 2012). Byssal threads were not produced by either species at salinities of 4 psu or higher. Byssogenesis was similar in both species below 25 °C, a value that was not exceeded at our sampling locations. The maximum flow velocity in the littoral zone of major rivers and canals lay outside the physiological tolerances of both dreissenid species. However, if peak values created by passing shipping are removed, ambient flow velocities are not limiting. The presence of mussels at sites with a pH of between 7.2 and 8 supports the view that pHs below 8 while not ideal, do not rule out dreissenid invasion (Mackie 2005). pH has a greater impact on survival at lower calcium concentrations (Claudi et al. 2012). In general, calcium concentrations in Dutch freshwaters are relatively high and do not limit the colonisation of the quagga mussel (Matthews et al. 2012d). Acidified bogs, moorland pools in the southern, central and eastern parts of the Netherlands and locations with seepage of

CO2-rich groundwater are examples of habitats where pH levels exclude quagga and zebra mussels (Leuven et al. 1992). Dreissenids are relatively salinity tolerant and salinity levels at sampling sites lay well within the maximum tolerance of the quagga mussel. Estuaries, coastal lagoons and brackish inland waters featuring salinities above 5 psu are the only areas where the quagga mussel may be limited (Matthews et al. 2012d). Temperature will probably present no barrier to quagga mussel colonisation as maximum water temperatures in Dutch freshwaters rarely exceed 25 °C (Matthews et al. 2012d). Moreover, mussels may acclimatise to adverse pH, temperature and salinity conditions (Bowman & Bailey 1998; Thorp et al. 1998).

Core samples taken from soft substrates established the presence of quagga mussels in one location only (Table 5.2). Dreissenids exhibit a preference for the colonisation of hard substrates (Wilson et al. 2006). In areas where hard substrates such as groyne stones are absent, colonisation is restricted to unionid mussel shells and in the case of the IJsselmeer and Markermeer to old empty sea shells such as Mya arenaria (Noordhuis et al. 2010) and zebra mussel beds. However, there is some evidence to suggest that a profundal variety of the quagga mussel can establish on soft substrates. The profundal variety of the quagga mussel has recently been recorded in the Cheboksary Reservoir situated in the midstream of the Volga River (Russian Federation) (Pavlova 2012). Moreover, quagga mussels have been recorded in the profundal zones of the Great Lakes of North America and the Volga river catchment (Orlova et al. 2005). However, the relatively shallow depth of Dutch water bodies may prevent colonisation by the profundal variety of the quagga mussel in the Netherlands.

A number of factors may explain the more recent range expansion of the quagga mussel relative to the zebra mussel. In general, the quagga mussel originated from a much smaller indigenous range resulting in a much smaller chance of dispersal (Van der Velde et al.

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2010b). The quagga mussel has a lower respiration rate, a greater body mass and shell growth than the zebra mussel (Stoeckmann 2003). At 25 °C and a salinity of 0.2 psu the zebra mussel exhibits a significantly higher byssogenesis rate compared to the quagga mussel increasing the likelihood of substrate attachment (Grutters et al. 2012). Moreover, the quagga mussel produces fewer byssal threads in flowing water (Peyer et al. 2009). The quagga mussel may suffer from selective predation as its thinner shell allows even large individuals to be crushed and digested by fish (Zhulidov et al. 2006). The recent increasing abundance of predatory Ponto-Caspian gobies in the Netherlands may have intensified this effect (Matthews et al. 2012d). However, the most efficient European dreissenid predator, the roach (Rutilus rutilus), is limited by prey size and not by crushing force (Nagelkerke & Sibbing 1996). This suggests that shell thickness will play a less significant role in predation by this species. During reproduction, zebra mussels release more eggs and gametes than quaggas in general (Stoeckmann 2003). As a result, the zebra mussel may demonstrate greater reproductive success and be more likely to colonise new habitats than the quagga mussel. Orlova et al. (2005) suggested that a preference to lentic conditions may have resulted in the quagga mussels delayed range expansion. Quagga mussel invasions started only when newly created ecosystems (i.e. reservoirs) changed from riverine to lacustrine (Orlova et al. 2005). Moreover, quagga mussels prefer more oligotrophic conditions compared to the zebra mussel (Therriault & Orlova 2010). High densities of filter feeding zebra mussels may encourage oligotrophication and so facilitate the colonisation of the quagga mussel.

It is expected that a number of changes in environmental parameters will occur that may affect the establishment of the quagga mussel in Dutch fresh-water bodies. A quantitative prediction of the effects of these changes on establishment success of quagga mussels is currently not possible due to a lack of data and predictive models. However, overall it is expected that in large rivers an increase in the abundance of quagga mussels will be limited by bank rehabilitation (e.g. removal of hard substrates) and an increase in peak discharges, wave stress due to expanding shipping activities, water depth fluctuations (desiccation) and increasing abundance of predatory Ponto-Caspian gobies (Matthews et al. 2012d). Rehabilitation of floodplain lakes and side channels may, however, increase the likelihood of quagga mussel establishment due to their preference for lentic environments (Zhulidov et al. 2010). Increases in macrophyte populations may lead to an increase in the relative abundance of the zebra mussel over the quagga mussel. The quagga mussel is found in lesser abundance in habitats with a high macrophyte coverage in comparison to the zebra mussel (Diggins et al. 2004; Zhulidov et al. 2010; Karatayev et al. 2011b).

It is commonly suggested that the zebra mussel is gradually replaced by the quagga mussel following its establishment, particularly in deeper more oligotrophic habitats (Ricciardi & Whoriskey 2004; Orlova et al. 2005; Zhulidov et al. 2010). Species replacement may have happened in the River Nederrijn at the Bakenhof side channel, the Meuse-Waal canal at

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Dukenburg and in the Hollandsch Diep in the Netherlands (Table 5.5). The analysis of coexistence demonstrates the rapidity with which the quagga mussel has made inroads into zebra mussel habitat since its discovery in the Netherlands in 2006. Quagga and zebra mussels have overlapping niches increasing the chance that they will colonise the same water bodies. Once established, increasing mussel density will play an important role in determining possible inter-specific interactions leading to the dominance of the quagga mussel. Evidence of increasing relative abundances was identified (Fig. 5.7). Dominance shifts from zebra mussels to quagga mussels were reported for the river estuary freshwater remnants Haringvliet and Lake Volkerakmeer, with relative abundances of quagga mussels reaching >95 and 99%, respectively (Bij de Vaate et al. 2010a 2011). However, analysis of the causes of species dominance is hampered by a lack of consecutive measurements of overall dreissenid density at sampling locations. A consistent approach and a focus on mussel density as-well as relative abundance is required to determine whether the quagga mussel is replacing the zebra mussel in Western European freshwaters or is additional to the zebra mussel, increasing the total numbers of dreissenid mussels. Dominance shifts from zebra mussels to quagga mussels may be caused by interaction (facilitation, competition) or factors not influenced by mussel interaction. Relative abundances give no insight into the mechanism leading to a dominance shift nor do they indicate actual species replacement. Consistent series of density data will be required to unravel the mechanism of species dominance (e.g. by species replacement).

This information becomes important if the impacts of the quagga mussel are different to that of the zebra mussel. Evidence of differing impacts is limited and conflicting. The condition of diving ducks that feed on dreissenids in the Great Lakes of North America, such as the Lesser and Greater Scaup (Aythya affinis and Aythya marila) is affected by the higher concentrations of selenium found in the quagga mussel (Schummer et al. 2010). A thinner shell allows larger quagga individuals to be crushed and digested leading to selective predation by certain fish species (Zhulidov et al. 2006). Zebra mussels are more likely than quagga mussels to colonise native unionid mussel shells inferring a greater impact from zebra mussel presence (Conn & Conn 1993). Heiler et al. (2011), however, suggest that the impact on European unionids will probably be similar to that of the zebra mussel. Increased levels of toxic cyanobacteria have been linked to quagga mussel invasion alone (Sarnelle et al. 2010) and together with the zebra mussel (Makarewicz et al. 1999; Zhang et al. 2011). However, during a laboratory experiment, Dionisio Pires et al. (2007) observed that zebra mussel grazing reduced the levels of the cyanobacterium Planktothrix agardhii.

Coexistence of both species and an increase in the overall dreissenid density will increase the overall impact of dreissenids on the aquatic ecosystem. An increase in overall dreissenid abundance as a result of coexistence may occur due to the quagga’s wider depth tolerance and lower respiration rate that decreases metabolic cost and increases tolerance of

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oligotrophic conditions (Mills et al. 1996; Baldwin et al. 2002; Stoeckmann 2003). Moreover, dreissenids display differences in substrate preferences (Mills et al. 1996; Karatayev et al. 1998; Diggins 2004), exhibit different temperature and silt tolerances and nutrient requirements (Roe & MacIsaac 1997; Karatayev et al. 1998; Orlova et al. 2005).

Acknowledgements

This paper has been presented by RSEW Leuven at the 18th International Conference on Aquatic Invasive Species (ICAIS) in Niagara Falls (Canada), April 21–25, 2013. The authors would like to acknowledge the Schure-Beijerinck-Popping Fund (SBP2012/54) and the Invasive Alien Species Team (TIE) of the Netherlands Food and Consumer Product Safety Authority, Ministry of Economy, Agriculture and Innovation (TRCPD/2010/3092) for their financial support. We would also like to thank Jeroen Driessen, Yen Le and Saskia Smits for their assistance with field work and data acquisition, and two anonymous reviewers and associate editor Jamie Dick who gave valuable advice resulting in improvements to the article.

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6 A dominance shift from the zebra mussel to the invasive quagga mussel may alter the trophic transfer of metals

Jonathan Matthews Aafke M. Schipper A. Jan Hendriks T. Yen Le Abraham bij de Vaate Gerard van der Velde Rob S.E.W. Leuven

Environmental Pollution 203: 183-190. August 2015.

Invasive quagga mussel may alter the trophic transfer of metals

6.1 Abstract

Bio-invasions are a major cause of biodiversity and ecosystem changes. The rapid range expansion of the invasive quagga mussel (Dreissena rostriformis bugensis) causing a dominance shift from zebra mussels (Dreissena polymorpha) to quagga mussels, may alter the risk of secondary poisoning to predators. Mussel samples were collected from various water bodies in the Netherlands, divided into size classes, and analysed for metal concentrations. Concentrations of nickel and copper in quagga mussels were significantly lower than in zebra mussels overall. In lakes, quagga mussels contained significantly higher concentrations of aluminium, iron and lead yet significantly lower concentrations of zinc66, cadmium111, copper, nickel, cobalt and molybdenum than zebra mussels. In the river water type quagga mussel soft tissues contained significantly lower concentrations of zinc66. Our results suggest that a dominance shift from zebra to quagga mussels may reduce metal exposure of predator species.

6.2 Introduction

Bio-invasions are one of the major, and growing, causes of biodiversity loss (European Commission 2013). The EU Biodiversity Strategy (European Commission 2011), and the International Union for Conservation of Nature (IUCN) guidelines for the prevention of biodiversity loss caused by invasive alien species (IAS), stress the need to identify the most harmful invaders (ISSG 2000; Katsanevakis et al. 2013). Currently, the rapid range expansion of the quagga mussel (Dreissena rostriformis bugensis) is resulting in a dominance shift from the established zebra mussel (Dreissena polymorpha) to the quagga mussel (Diggins 2001; Bonhof et al. 2009; De Rooij et al. 2009; Bij de Vaate 2010a; Heiler et al. 2012; Matthews et al. 2014b; Bij de Vaate et al. 2013). Both dreissenid freshwater bivalves appear to be invasive in Western Europe and North America (Neumann & Jenner 1992; Mitchell et al. 1996; Watkins et al. 2007; Gonzalez & Downing 1999; Ward & Ricciardi 2010; Matthews et al. 2014b), and are an important source of food for native water birds, fish, crayfish and crabs (Kelly et al. 2010; Mörtl et al. 2010; Van Eerden & De Leeuw 2010; Bij de Vaate 2010a; Noordhuis et al. 2010; Matthews et al. 2014b). Some waterfowl species have been reported to alter their dietary intake and migration patterns in response to the ready availability of zebra mussels (Petrie & Knapton 1999).

The ability of both these mussel species to filter large quantities of water allows them to accumulate toxicants, which may lead to the secondary poisoning of native predator species (Rutzke et al. 2000; Kwon et al. 2006; Hogan et al. 2007; Mueting & Gerstenberger 2010). Accumulation of toxicants may lead to mortality and sub-lethal effects such as altered growth, reproduction, and behaviour (Flemming & Trevors 1989; Custer & Custer 2000; Santore et al. 2002; Custer et al. 2003; Petrie et al. 2007). Metal accumulation has been implicated in many of these effects. For example, cadmium transfer from zebra mussels to

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the tufted duck (Aythya fuligula) has resulted in behavioural disturbances in adults, growth retardation, and embryonic mortality (De Kock & Bowmer 1993). Selenium toxicity impacts staging, winter body condition and health of lesser and greater scaup (Aythya affinis and Aythya marila), diving ducks that feed primarily on dreissenids (Custer & Custer 2000; Custer et al. 2003; Petrie et al. 2007). Accumulation of copper may cause mortality and sub lethal effects such as altered growth, reproduction, and behaviour in fish and macroinvertebrate species (Flemming & Trevors 1989). High concentrations of zinc may result in calcium uptake inhibition in certain fish species (Santore et al. 2002). Moreover, lead has long been considered one of the most significant metals from the standpoint of environmental contamination and toxicology (Scheuhammer 1987).

However, metal concentrations in mussel soft tissues may vary, depending on species- specific factors such as reproduction cycle, filtration rate, and ventilation rate (Kraak et al. 1991; Veltman et al. 2008). Therefore, a shift in dominance from established to newly invading mussel species may alter the trophic transfer of metals to native predators. Peer reviewed literature focussing on potential differences in accumulation of metals between the soft tissues of the quagga and zebra mussel is scarce, often inconclusive and limited to North America (Johns & Timmerman 1998; Rutzke et al. 2000; Richman & Somers 2005; Le et al. 2011). Moreover, studies reporting metal concentrations in quagga mussels are particularly rare. This article aims to (1) identify potential inter-species differences in metal concentrations between the invasive quagga and established zebra mussel; (2) identify potential intra-species differences in metal concentrations in the soft tissues of quagga mussels in relation to shell size and water type (i.e., river or lake); (3) discuss the possible implications of these inter- and intra-species differences in relation to a dominance shift from zebra mussels to invading quagga mussels for the trophic transfer of metals.

6.3 Methods

6.3.1 Field survey and chemical analyses

Zebra and quagga mussels were collected by hand from groyne stones from four river locations and with a trawl net from two lake locations in the Netherlands (Fig. 6.1, Table 6.1). These sites were selected based on available evidence on the co-existence of the two species. Mussels were separated according to species and size class (small: <15 mm, medium: 15-22 mm, and large: >22 mm). The mussels were not depurated prior to extraction from their shells as this more accurately reflects metal exposure to predators. Mussel predators consume the entire mussel and are therefore exposed to both stomach contents and mussel tissue. Metal concentrations in dreissenid shells have been found to be orders of magnitude lower than in mussel soft tissue (Van der Velde et al. 1992). Therefore, metal concentrations in mussel shells were considered negligible and shells were not included in the analysis. The soft tissue was extracted from mussel shells and subsequently

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dried at 70 oC for 24 h. Dried samples were then weighed using a Sartorius LA310s micro balance (Sartorius AG, Göttingen, Germany) to produce replicates of 0.2 g dry weight. The dried samples were digested with 4 ml HNO3 65% and 0.5 ml H2O2 in a Milestone Ethos D microwave. Following digestion, 100 ml of high quality deionized water was added to each sample. In addition, blanks were prepared to allow for corrections to metal concentrations determined from mussel samples. Analysis of metal concentrations was undertaken using inductively coupled plasma mass spectroscopy (ICP MS) for aluminium (Al), chromium (Cr), manganese (Mn), iron (Fe), cobalt (Co), nickel (Ni), copper (Cu), zinc (Zn66 and Zn68), arsenic (As), selenium (Se), molybdenum (Mo), cadmium (Cd111 and Cd112), tin (Sn), mercury (Hg), and lead (Pb). We considered the complete range of metals measured by ICP-MS as this gives the most complete insight into possible changes in metal exposure of predator species resulting from a dominance shift in prey species.

Figure 6.1: Locations sampled during field surveys.

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6.3.2 Data analysis

A number of comparisons were made to identify inter-species and intra-species differences in soft tissue metal concentrations (Tables 6.2 and 6.3). Data on metal concentrations were aggregated and compared according to mussel species, size class and origin of sampled material (water type and water body). Water bodies included in the analysis were the lakes IJsselmeer and Markermeer, and rivers Waal, Nederrijn, and Meuse (Fig. 6.1).

Table 6.1: Sampling locations and number of individuals collected. Sampling location Water Coordinates Sampling period Number of Number of type zebra mussels quagga mussels Ewijkse Plaat, River 51o 52’ 50.7” N; October 2008 287 35 River Waal 23o 27’ 30.7” E Lexkesveer, River River 51o 57’ 33.9” N; April 2010 9 415 Nederrijn 5o 41’ 22.7” E Wageningen River 51o 57’ 41.7” N; April 2010 162 64 harbour, River 5o 39’ 31.4” E Nederrijn Near Belfeld, River 51o 19’ 13.6” N; April 2010 90 598 River Meuse 6o 6’ 55.2” E Lake IJsselmeer Lake 52o 47’ 52.2” N; June, July, September 1156 2471 5o 19’ 24.9” E and November 2009 Lake Markermeer Lake 52o 32’ 18.0” N; July, November 2009 870 308 5o 15’ 20.1” E

In the inter-species comparisons, only paired samples were used (same location and time period). The number of samples for each species included in these analyses was, therefore, the same. Inter-species comparisons consisted of analyses (1) combining all size classes and locations, (2) per water type (river or lake) and combining size classes, and (3) per size class and combining all water types (Table 6.2). Intra-specific comparisons were made for the quagga mussel. In order to reduce possible bias due to differences in sampling period, only samples taken in 2009 and 2010 were used for this comparison.

Table 6.2: Number of paired samples used in the analyses of inter-species differences in soft tissue metal concentrations. Data aggregation per species Number of paired samples analysed per metal Overall 13 Large size class 3 Medium size class 5 Small size class 4 Lake water type 6 River water type 7

To obtain sufficient statistical power, concentrations were aggregated either by size class or by sampling location. This was done (1) per size class combining all locations, (2) for lake and river water types combining all size classes and aggregating data for all water bodies

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within each water type, (3) individual rivers combining all size classes, and (4) individual lakes combining all size classes (Table 6.3).

Table 6.3: Number of samples used in the analyses of intra-species differences of soft tissue metal concentrations in quagga mussels (Dreissena rostriformis bugensis). Data aggregation Locations and number of samples analysed per metal Lakes IJsselmeer (3); Markermeer (3) Rivers Nederrijn (6); Meuse (3) Size class Large (5); medium (5); small (5) Water type Lakes (6); rivers (9)

6.3.3 Statistical analysis

Potential differences between groups were tested for statistical significance using ANOVA or the non-parametric Mann Whitney test (IBM SPSS Statistics, release 20.0.0). Non parametric tests were applied if the data was not normally distributed and/or if non homogenous variance between groups was determined following log transformation. The Shapiro Wilk and Levene tests were applied to assess normality and equality of variance, respectively.

6.4 Results

6.4.1 Inter-species differences in metal concentrations in soft tissues

The overall comparison of the two species revealed significantly higher concentrations of nickel and copper in the zebra mussel than in the quagga mussel (Fig. 6.2A, Table 6.4). In the lake water type, zebra mussel soft tissues contained significantly higher concentrations of zinc66, cadmium111, copper, nickel, cobalt and molybdenum than quagga mussel soft tissue by a factor of 1.5-2.3. However, concentrations of aluminium, iron and lead in quagga mussel soft tissue were significantly higher than those in zebra mussel soft tissue, by factors of 1.7-1.9, respectively (Fig. 6.2B, Table 6.4). In the river water type zebra mussel soft tissues contained significantly higher concentrations of zinc66 (Fig. 6.2C, Table 6.4). For size classes, no significant differences between the two species were found (Fig. 6.2D-F).

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Table 6.4: Statistically significant inter-specific differences in metal concentrations found between zebra mussel (Dreissena polymorpha) and quagga mussel (Dreissena rostriformis bugensis) tissues sampled in different water types, lakes and rivers. Variable Metal ANOVA Mann Whitney Overall1 Nickel Na U = 44.00, z = -2.08, P < 0.05 Overall1 Copper Na U = 28.00, z = -2.90, P < 0.05 Lake water type2 Aluminium Na U = 4.00, z = -2.24, P < 0.05 Lake water type2 Copper Na U = 0.00, z = -2.88, P < 0.05 Lake water type2 Zinc66 Na U = 0.00, z = -2.88, P < 0.05 Lake water type2 Cobalt Na U = 2.00, z = -2.56, P < 0.05 Lake water type2 Iron Na U = 3.00, z = -2.40, P < 0.05 Lake water type2 Nickel Na U = 0.00, z = -2.88, P < 0.05 Lake water type2 Molybdenum Na U = 4.00, z = -2.24, P < 0.05 Lake water type2 Lead Na U = 4.00, z = -2.24, P < 0.05 Lake water type2 Cadmium111 F(1, 10) = 13.54, P < 0.05 Na River water type3 Zinc66 Na U = 0.00, z = -3.13, P < 0.05 Na: statistical test not applicable; 1all subgroups combined; 2lakes combined; 3rivers combined.

6.4.2 Intra-species differences in metal concentrations in soft tissue

Iron, zinc66 and zinc68 were accumulated at significantly higher concentrations in quagga mussel soft tissue obtained from rivers than from lakes, by factors of 2.5, 1.5 and 1.8, respectively (Fig. 6.3, Table 6.5). No metals were present in significantly higher concentrations in quagga tissue obtained from lakes compared to rivers.

Table 6.5: Statistically significant differences in metal concentrations in the soft tissues of quagga mussels (Dreissena rostriformis bugensis) sampled in different water types: lakes and rivers (intra- species differences). Variable Metal ANOVA Mann Whitney Water type1 Iron F(1, 13) = 19.57, P < 0.05 Na Water type1 Zinc66 Na U = 3.00, z = -2.83, P < 0.05 Water type1 Zinc68 Na U = 3.00, z = -2.83, P < 0.05 Rivers2 Cadmium111 Na U = 0.00, z = -2.32, P < 0.05 Rivers2 Cadmium112 Na U = 0.00, z = -2.32, P < 0.05 Rivers2 Iron Na U = 0.00, z = -2.32, P < 0.05 Rivers2 Manganese Na U = 1.00, z = -2.07, P < 0.05 Rivers2 Cobalt Na U = 0.00, z = -2.32, P < 0.05 Rivers2 Nickel Na U = 0.00, z = -2.32, P < 0.05 Rivers2 Mercury Na U = 0.00, z = -2.32, P < 0.05 Rivers2 Tin Na U = 1.00, z = -2.07, P < 0.05 Rivers2 Zinc66 Na U = 0.00, z = -2.32, P < 0.05 Rivers2 Zinc68 Na U = 0.00, z = -2.32, P < 0.05 Rivers2 Arsenic Na U = 0.00, z = -2.32, P < 0.05 Rivers2 Selenium Na U = 0.00, z = -2.32, P < 0.05 Na: statistical test not applicable; 1lakes combined, rivers combined; 2Nederrijn, Meuse.

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16 16 Fe 14 14 Fe Al Al Mn Mn 12 12

drywt) Zn

1 1 Zn - 10 dry wt) Cu Cu 1 10 As - Ni Se Ni Pb As 8 CrSe 8 Pb Cd Cr Co Cd Sn Sn 6 6 Co Mo Mo 4 4

Hg Hg Quagga mussel Quaggamussel (Ln µg.kg Quagga mussel µg.kg (Ln mussel Quagga 2 2

0 0 0 2 4 6 8 10 12 14 0 2 4 6 8 10 12 14 Zebra mussel (Ln µg.kg-1 dry wt) Zebra mussel (Ln µg.kg-1 dry wt) 16 16 Fe Fe Al 14 Al 14 Mn 12 12 Mn

dry dry wt) Zn

Zn 1 - dry dry wt) 10 Cu

1 1 10 - Cu Se As Pb Ni Pb As Ni 8 8 Cd Cr Se Cr Co Sn Cd Co 6 6 Mo Sn Mo

4 4 Quagga mussel Quaggamussel (Ln µg.kg Hg Hg Quagga mussel Quaggamussel (Ln µg.kg 2 2

0 0 0 2 4 6 8 10 12 14 0 2 4 6 8 10 12 14 Zebra mussel (Ln µg.kg-1 dry wt) Zebra mussel (Ln µg.kg-1 dry wt) 16 16 Fe 14 Al 14 Fe Al Mn

12 12 Mn

dry dry wt)

dry wt) dry

1 1

- 1 1

- Zn Zn 10 Cu 10 Ni As As Pb Ni Cu Cr 8 Se 8 Cd Cr Cd Se Pb Co 6 6 Sn Co Sn Hg Mo Mo

4 4

Quagga mussel Quaggamussel (Ln µg.kg Quagga mussel Quagga mussel µg.kg(Ln Hg 2 2

0 0 0 2 4 6 8 10 12 14 0 2 4 6 8 10 12 14 Zebra mussel (Ln µg.kg-1 dry wt) Zebra mussel (Ln µg.kg-1 dry wt) Figure 6.2: Comparison of average metal concentrations in quagga mussel (Dreissena rostriformis bugensis) and zebra mussel (Dreissena polymorpha) soft tissues in A) all subgroups combined; B) lake water type; C) river water type; D) <15 mm shell length; E) 15e22 mm shell length; F) >22 mm shell length. Error bars represent Ln standard deviation.

The analysis of individual water bodies revealed no significant differences in metal concentration between quagga mussel soft tissues sampled from lake Markermeer and lake IJsselmeer (Table 6.5). Tissues of quagga mussels sampled from the river Meuse showed significantly higher concentrations of mercury, cadmium111, cadmium112, iron, manganese, cobalt, nickel, zinc66, zinc68, arsenic, selenium and tin than samples taken from the river Nederrijn (Table 6.5).

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16

14 Fe 12 Al

Mn

dry dry wt) 1 1 - 10 Ni Zn As Cu Cr Se 8 Cd Sn Pb 6 Co Hg Mo Lakes Lakes (Ln µg.kg 4

2

0 0 2 4 6 8 10 12 14 Rivers (Ln µg.kg-1 dry wt) Figure 6.3: Comparison of dry weight metal concentrations in the soft tissues of quagga mussels (Dreissena rostriformis bugensis) from river and lake water types.

6.5 Discussion

In the present study we investigated potential differences in soft tissue metal concentration between quagga and zebra mussels, as well as intra-species differences in quagga mussel metal concentrations in relation to water type and shell size. To that end, we collected and analysed the soft tissue of zebra and quagga mussels of different size classes originating from various water bodies in the Netherlands. In order to minimise bias resulting from temporal and spatial variation in sampling intensity, paired samples (i.e., same location and sampling period) were used for the species and size class comparisons. In order to reduce possible bias due to differences in sampling period in the intra-species comparisons of size and water body, only samples from the years 2009 and 2010 were used. Therefore, spatial and temporal variations in metal concentrations are unlikely to have influenced the results.

6.5.1 Differences in soft tissue metal concentrations between quagga and zebra mussels

A number of mainly American and Canadian studies on metal accumulation in dreissenid mussels support our findings. Richman & Somers (2005) found generally higher concentrations of zinc, nickel, copper, and cadmium in soft tissues of zebra mussels than in quagga mussels from the Niagara river, Canada. Our results also agree with the estimates of the dynamic bioaccumulation model developed by Le et al. (2011), as well as the results of a field study in Lake Ontario, Canada, where higher copper and zinc concentrations were found in zebra mussels (Johns & Timmerman 1998). Moreover, mercury concentrations were similar in the two dreissenid species in lakes Mead, Mohave and Havasu, the United States, consistent with our results (Mueting & Gerstenberger 2010). However, in a study

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examining seasonal and inter-annual variation, higher concentrations of cadmium were found in quagga mussels than zebra mussels sampled from the outflow of lake Ontario, Canada, contrasting with the results of this study (Johns & Timmerman 1998). Moreover, Rutzke et al. (2000) found no statistical differences in a large set of soft tissue metal concentrations between quagga and zebra mussels sampled in June 1997 from lakes Erie and Ontario, Canada.

Different habitat characteristics may explain the contrasting results shown in these studies as compared to our results. Dutch lakes are relatively shallow compared to those present in North America such as the Great Lakes. For example, the largest lake in the Netherlands, Lake IJsselmeer, has a mean depth of 4.5 m (Berger & Sweers 1998). Different depth and temperature regimes may play a role in the exposure and absorption of metals in mussels (Wiesner et al. 2001; Veltman et al. 2008). When zebra and quagga mussels occur in the same lake, zebra mussels generally reach their highest densities in warm, shallow water, whereas quagga mussels occur in colder, deeper water (Dermott & Munawar 1993; Mills et al. 1996). Clearance rate and temperature relationships are reported in dreissenid mussels and dreissenid filtering activity may be positively influenced by temperature (Reeders & Bij de Vaate 1990; Borcherding 1992). Therefore, the quagga mussels preference for colder, deeper water than the zebra mussel and a possible relationship between temperature and filtration rate may explain the differences in metal accumulation measured in these studies.

Apart from differences in habitat preference and characteristics, the inter-species difference in metal accumulation may be related to differences in physiological characteristics, e.g. energy expenditure, filtration and growth rate. For example, pollutant uptake in dreissenids has been linked to filtration rate (Kraak et al. 1991; Bervoets et al. 2004; Veltman et al. 2008; Le et al. 2011). Veltman et al. (2008) attributed high uptake rate constants in mussels to their high filtration rate. However, contrasting observations of inter-species differences in filtration rate in quagga and zebra mussels have been reported. Ackerman (1999) was unable to demonstrate any inter-species difference in filtration rate. Yet, quagga mussels have been found to contain greater soft tissue dry mass than zebra mussels per unit shell length (Baldwin et al. 2002). Ackerman's (1999) observations were based on mussel length only and not corrected for inter-species differences in soft tissue mass. Diggins (2001) compared individual mussel and ash-free dry weight filtration rates and found that, in both cases, quagga mussels filtered significantly faster than zebra mussels. Baldwin et al. (2002), on the other hand, observed that when samples were corrected for soft tissue dry mass, zebra mussel clearance rates (volume of water cleared of particles per unit time as a result of filtration) were two to seven times higher than those of the quagga mussel. Higher filtration rates in zebra mussels may account for the higher metal concentrations we observed in this species. However, inter-species differences in filtration rate may depend on local conditions, mussel size class and the definition applied to mussel size.

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The quagga mussel expends less energy on reproduction, shell development and respiration than the zebra mussel, allowing it to invest more energy in soft tissue growth (Stoeckmann 2003; Casper & Johnson 2010). Consequently, the quagga mussel is able to grow quicker than the zebra mussel (Baldwin et al. 2002; Karatayev et al. 2011a). Higher growth rates in quagga mussels may, in turn, result in higher growth dilution when compared to zebra mussels, leading to a decrease in body metal concentration with age and size (Carrasco et al. 2008; Le et al. 2011). Moreover, differences in metal concentrations between other mussel species have been related to differences in growth rate rather than to any direct difference in metabolism (Lobel et al. 1990). However, in a study of lake Erie, North America, growth rates were shown to be very similar in quagga and zebra mussels, which corresponds with the lack of statistically significant inter-species differences in metal concentrations observed in this lake (Macisaac 1994; Rutzke et al. 2000).

6.5.2 Differences in mussel soft tissue concentrations with mussel size

Many Eurasian predatory fish species, including the roach (Rutilus rutilus), are highly selective of prey size (Nagelkerke & Sibbing 1996). A change in dominance from zebra mussels to invading quagga mussels may result in changes in the trophic transfer of metals that differs between mussel size classes. However, in this study, no inter- or intra-species differences in metal concentration within size classes were found. In literature, some relationships have been reported between shell size and metal concentrations. A study of the Niagara River in Canada indicated that larger quagga and zebra mussels (i.e., 16-25 mm) generally had higher tissue concentrations than smaller mussels (<15 mm length) (Richman & Somers 2005). However, differences in tissue concentrations between size classes were present only for cadmium, copper and manganese, making these results inconclusive. Mills et al. (1993) examined zebra and quagga mussels in lake Ontario, Canada, and found significant differences in manganese concentrations between large and small quagga mussels. In both lakes Erie and Ontario, Canada, Rutzke et al. (2000) found no significant differences between quagga and zebra mussel size classes for a wide range of metals. A possible explanation for the lack of statistically significant results between quagga size classes may be that continued metal accumulation and higher filtration rates in larger mussels are counter balanced by growth dilution. Based on these results, no conclusions can be drawn regarding changes in trophic transfer of metals based on prey size following a change in mussel dominance to the invasive quagga mussel.

6.5.3 Differences in metal concentration in the soft tissues of mussels between water types

Invasive mussels may accumulate metals to a higher degree in certain water types compared to others, leading to possible differences in trophic transfer of metals between water types in the event of invasion and selection of quagga mussels by predators. Lower

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metal concentrations observed in the soft tissues of mussels sampled from lakes compared to those from rivers may be explained by lower metal contamination/pollution levels in lake water compared to river water at the sampling locations. Higher concentrations of metals have been observed near river mouths in lakes compared to other locations within the same lakes suggesting that river water may carry higher concentrations of pollutants than lake water (Rosales-Hoz et al. 2000; Kishe & Machiwa 2003). Current velocity and turbulent flow may be higher in rivers than lakes, for example because of intensive shipping. Resulting sediment re-suspension may increase metal concentrations in the water column (Vuori 1995; Rosales-Hoz et al. 2000; Ji et al. 2002). Sediment re-suspension may increase metal availability and lead to higher mussel soft tissue metal concentrations in rivers.

Significantly higher concentrations of the majority of metals tested in quagga mussels taken from the river Meuse compared to the river Nederrijn may be related to higher levels of metal pollution in the river Meuse. This may be applicable for the period from 2008 to 2010 when zinc and cadmium were present in higher average concentrations in the river Meuse at Eijsden than in the river Rhine at Lobith (Rijkswaterstaat 2014). However, this explanation does not hold for copper. Average copper concentrations were similar in the Rhine and the river Meuse for the same time period (Rijkswaterstaat 2014; ICPR 2014). Local variations in environmental metal concentration and differences in metal accumulation mechanisms may explain the disagreement between metal concentrations in water and mussel soft tissue concentration (Marie et al. 2006; Veltman et al. 2008). Intra- species differences in tissue metal concentrations between individual rivers, and river and lake water types, indicate that predator habitat preference may be an important consideration when assessing the consequences of aquatic bio-invasions for the trophic transfer of metals.

6.5.4 Implications for secondary poisoning

A change in species dominance from zebra to the invasive quagga mussel may alter the trophic transfer of metals as a result of inter-species differences in metal concentrations. The similarity in inter-species selenium concentration, and the significantly lower concentrations of copper (overall) and cadmium111, copper and zinc66 (in lakes) found in the quagga mussel suggest that, in general, a dominance shift and resulting dietary switch from zebra to quagga mussels will reduce exposure of predator species to a number of metals of concern at our sampling sites. Moreover, further reductions in exposure may occur if predator species exhibit a dietary preference for quagga mussels following a dominance shift from zebra to quagga mussels. The quagga mussel often has a larger soft tissue mass relative to shell size and a thinner shell than the zebra mussel (Baldwin et al. 2002; Zhulidov et al. 2006), which may be favoured by dreissenid predators. Eurasian cyprinids, some species of goby and whitefish are highly selective of molluscan prey (Starobogatov 1994), and selective fish predation may explain the widespread decline of the

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quagga mussel relative to the zebra mussel in the Don and Manych River systems in Russia (Zhulidov et al. 2006). In conclusion, in the event of a switch in dominance from the zebra to the invasive quagga mussel, lower concentrations of copper, cadmium and zinc measured in our analysis of quagga mussels may contribute to the overall reduction in trophic transfer of these metals, adding to metal exposure reduction resulting from general improvements in water quality (Durance & Ormerod 2009).

Acknowledgements

We would like to thank the Invasive Alien Species Team (TIE) of the Netherlands Food and Consumer Product Safety Authority (NVWA), Ministry of Economic Affairs (TRCPD/2010/3092) and the Schure Beijerinck-Popping fund (SBP2012/54) for financial support, Jelle Eygensteyn for his assistance in carrying out the ICP MS analysis, Aryan Ransijn for his help during field work and metal analyses and two anonymous reviewers who provided many constructive comments.

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7 Synthesis

Synthesis

7.1 Introduction

An initial aim of this thesis was to ascertain which indicator groups are most appropriate for the measurement of short term success in river rehabilitation. A conclusion derived from this is that, until very recently, invasive alien species (IAS) were not explicitly or systematically considered during river rehabilitation planning and assessment. This conclusion led to the formulation of the central aim of constructing a framework for the assessment of potential risks that alien species pose to river rehabilitation success. The ultimate goal of the thesis is to contribute to the body of knowledge that aims to ensure that river rehabilitation planning leads to practical interventions with respect to IAS that will increase the ecological integrity of the target ecosystem with or without alien species. To that end, different approaches to the prioritisation and assessment of the risk posed by potential IAS were presented.

The current chapter addresses the central aim of this study by selecting, synthesising, contrasting and comparing relevant research results from the preceding chapters and discussing their implications for the planning process in ecological rehabilitation with special reference to large rivers. Firstly, the results regarding the current state of IAS assessment in river rehabilitation planning (chapter two) will be examined in Section 7.2. Secondly, the implications of the various approaches to risk prioritisation and assessment of IAS, described in chapters three and four, to ecological rehabilitation planning will be considered (Section 7.3). Finally, the implications of example assessments carried out for the quagga mussel (Dreissena rostriformis bugensis) for river rehabilitation projects in Western Europe (chapters five and six), and the bridging of knowledge gaps will be considered (Section 7.4).

7.2 Invasive species risk assessment in river rehabilitation planning

The European Union’s Water Framework Directive (WFD) aims to improve the ecological status of the Union’s freshwater bodies, but does not refer to the impacts and management of alien species. A survey carried out by the Joint Research Centre of the European Commission showed that the majority of member states do not take alien species explicitly into account for the classification of freshwater quality under the WFD (European Commission 2000; Vandekerkhove & Cardoso 2010). Instead, member states use methods to measure ecosystem degradation or improvement while not necessarily identifying which drivers may be attributed to which pressures, including those exerted by IAS. The questionnaire investigating river manager’s interpretation of rehabilitation success, (Chapter 2), revealed a lack of awareness among managers of the potential impacts of IAS on river rehabilitation outcomes. Three open questions in particular gave opportunities for nature managers to describe how IAS were considered in river rehabilitation planning and project evaluation. However, none of the 54 respondents who originated from Germany, the

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Netherlands and the United Kingdom referred to alien species in any of the responses given. Even though alien species were not directly referred to in the questionnaire, it is remarkable that an issue that has been observed repeatedly to limit the success of river rehabilitation was neglected by all respondents, suggesting that it was considered unimportant. This supports observations made in the literature that suggest that IAS, to date, have not been considered in the ecological rehabilitation planning and assessment process (Woolsey et al. 2007; O’Donnell & Galat 2008; section 1.1.4), and may explain why species invasions often occur following rehabilitation interventions (Strayer et al. 2005; Bij de Vaate et al. 2006; Van der Velde et al. 2006a; section 1.1.3).

Ironically, alien species may be purposefully introduced to water-bodies where no earlier record of the species was made in an effort to achieve an improvement in water quality. For example, in the Netherlands the invasive bivalve D. rostriformis bugensis has purposefully been introduced to ponds and lakes in an effort to reduce the effects of eutrophication on water clarity and suppress cyanobacterial and algal blooms (Brabantse Delta 2013; De Hoop et al. 2015; Dutch Water Tech 2015; Waajen et al. 2016). Moreover, previous research has been undertaken that considered the closely related zebra mussel (Dreissena polymorpha) for use as a biological filter in Europe (Reeders 1990; Reeders & Bij de Vaate 1990; Noordhuis et al. 1992; McLaughlan & Aldridge 2013). Recent guidelines have been proposed that aim to reduce the risks associated with the introduction of alien Dreissena species for water management purposes (De Hoop et al. 2015).

The lack of explicit reference to alien species in European regulations aimed at improving water quality and ecological status, a failure to consider alien species as a separate pressure on aquatic ecosystems at member state level, the purposeful introduction of alien species to improve certain aspects of water quality, and a lack of awareness of the potential threat of IAS by the managers of rehabilitation projects, underlines the need to incorporate alien species prioritisation and assessments in the planning of ecological rehabilitation of river catchments.

7.3 Implications of invasive species assessment to rehabilitation planning

7.3.1 Past and present approaches to river catchment management and rehabilitation planning

Past and present approaches to river rehabilitation planning have tended to frame project aims around existing problems within limited spatial boundaries, have ignored transboundary issues, and failed to predict future change that alter the invasiveness of riverine habitats and lead to the introduction of alien species. In the latter half of the 20th century, the river Rhine was virtually devoid of species due to high levels of pollution (Van der Velde et al. 1991). Pollution also acted as a barrier that hindered the dispersal of alien

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species between river catchments via artificial connections such as canals. However, this barrier was alleviated in the rivers Rhine and Meuse as a result of measures introduced by the Rhine Action Programme that ran from 1987 to 2000 and led to water quality improvements (ICPR 2016). The Rhine Action Programme and the project ‘Ecological Rehabilitation of the rivers Rhine and Meuse’ were initiated following the Sandoz disaster of 1986, and were two of the first initiatives to incorporate regional and transboundary ecological rehabilitation goals. An increase in the presence of alien species in these rivers following improvements in water quality is a prime example of the unintended effects of ecological rehabilitation (Bij de Vaate 2003). A similar situation occurred following the introduction of the original US Clean Water Act of 1972 that improved water quality, thereby removing pollution barriers to species invasion, but failed to address ballast water discharge and IAS (Lovett 2012), thereby allowing the introduction and spread of alien species such as Asian clams (Corbicula sp.) and the zebra mussel (Dreissena polymorpha). This issue was addressed in 2016 by an executive order ‘safeguarding the nation from the impacts of invasive species’ signed by President Barack Obama (Office of the Press Secretary 2016).

In January 2001, the Rhine 2020 programme on the sustainable development of the Rhine was put into action, succeeding the previous Rhine Action Programme (ICPR 2016). The targets of Rhine 2020 concentrate on the following fields of action: ecosystem improvement, flood prevention and protection, protection of water quality, and groundwater protection (ICPR 2016). There have been major successes in the area of water and habitat quality improvement in the river Rhine and a number of characteristic species returned as a result of these programmes, along with the introduction and establishment of a number of alien species. Following the opening of the German Main-Danube Canal that connected the river Rhine to the river Danube in 1992, the number and abundance of alien species in the Rhine has accelerated, often at the expense of native species (Leuven et al. 2009; Schöll 2015). For example, four alien goby species have established in the Rhine since the canals were establishment and now threaten the protected river bullhead (Cottus perifretum): the western tubenose goby (Proterorhinus semilunaris), bighead goby (Neogobius kessleri), round goby (Neogobius melanostomus), and the monkey goby (Neogobius fluviatilis) (Van Kessel et al. 2013, 2014, 2016). The racer goby (Babka gymnotrachelus) (completing all five species of Gobiidae) and the Amur sleeper (Perccottus glenii) (Odontobutidae) are both expected in the river Rhine in the near future (ICPR 2013). N. melanostomus now makes up, on average, 28% of fish species sampled by the ICPR, and in the German-French Upper Rhine there is, at certain locations, a relative frequency of N. melanostomus to other fish species of more than 90% (Korte et al. 2015). Native species have been displaced at these locations. For example, the previously regularly occurring Eurasian ruffe (Gymnocephalus cernuus) does not establish well in rocky habitats (e.g., river groynes), structures that are ideal for goby species allowing them to develop high population densities (Korte et al. 2015; Van Kessel et al. 2016).

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Similar approaches to water quality improvement have been made in the river Meuse by the international commission for the Meuse (IMC) that was created in 2002 and includes Belgium, the Netherlands, France, Germany, and Luxembourg. Reports of the IMC measurement network indicate that physico-chemical measures of water quality have improved, however, measures of biological quality highlighted an increase in alien taxa within the macro-benthos species group that by 2007 numbered 25, and were especially abundant in the lower reaches of the river (IMC 2011). Southward dispersal of alien species from the Rhine catchment is made possible by the many canals that connect the Rhine catchment and the Meuse (Ketelaars 2004). For example, the invasive gammarid Dikerogammarus villosus is thought to have moved south through the Meuse-Waal canal (the Netherlands) to the Meuse catchment, where it was first recorded in late 1996 (Liefveld et al. 2001; Ketelaars 2004). Similarly, P. semilunaris, N. fluviatilis and P. kessleri are thought to have accessed the river Rhine from the river Danube via the Main-Danube canal, and subsequently the river Meuse from the Rhine via numerous canals (Leuven et al. 2009; Van Kessel et al. 2016). However, measures of water quality status often do not include assessments of alien species (e.g., IMC 2015; Reeze et al. 2017).

Currently, the EU is focussing its attention on the development of River Basin Management Plans (RMBPs) that provide a road map for meeting the requirements of the WFD. The integrated research project Restoring rivers FOR effective catchment Management (REFORM) provides guidance for the design and implementation of the 2nd and future RMBPs for the WFD (REFORM 2016). REFORM provides a blueprint for the future ecological rehabilitation and management of river catchments within the EU. However, the project formulation section of the planning protocol for river rehabilitation provided by REFORM directs practitioners to identify issues affecting the water body, meaning that interventions and endpoints are based on a comparison of issues affecting the current state of the river section, potentially ignoring wider transboundary problems that may, in combination with rehabilitation measures, result in future changes that limit the success of rehabilitation measures, such as species invasion. Therefore, if the waterbody is free of IAS prior to rehabilitation, benchmarks relating to IAS may not even be included in post project evaluations of success.

7.3.1 Integration of assessments for alien species in river rehabilitation planning

The inclusion of an analysis of catchment scale, transboundary problems and the setting of related benchmarks during project formulation would assist in the identification of source populations of potential IAS that may establish following the implementation of measures and threaten rehabilitation success. Tools such as horizon-scanning, rapid assessment protocols and risk assessments may be used to identify alien species that may establish as a result of rehabilitation interventions (Chapters 3 and 4).

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A river catchment scale horizon-scan may be used to identify alien species present within the boundaries of a rehabilitation project amenable to eradication, and identify and risk prioritise alien species present within the river catchment that have access to the area earmarked for rehabilitation due to the presence of vectors and pathways of introduction. Rapid assessment protocols and risk assessment techniques may then be used to assess if planned rehabilitation interventions will potentially increase the likelihood that alien species will establish and become invasive. For example, interventions that reduce water flow velocity and introduce hard substrates would improve habitat suitability for alien dreissenid mussels and increase the likelihood of species invasion in the presence of source populations.

Traditionally, river rehabilitation planning has used a historical or pristine reference to create benchmarks that provide the basis for interventions and evaluation. However, the use of a pristine reference condition ignores the existence of socio-economic constraints such as flood safety requirements, cooling water discharge, and the need for unhindered commercial navigation in large river systems. To allow for these socio-economic limitations, Kern (1992) and Muhar (1996) introduced the ‘leitbild’ concept that is based on experience with rehabilitation projects in Austria and Germany. A pristine reference condition or ‘visionary leitbild’ is compared to the existing condition to pinpoint current deficits. Current deficits are then assessed to allow for existing constraints resulting in the development of an achievable set of rehabilitation objectives for the river system or ‘operational leitbild’ (Jungwirth et al. 2002; Lenders et al. 2003). Benchmarks to measure project success may then be created allowing for existing constraints.

Jungwirth et al. (2002) give the example of land use as an existing constraint, however, it may be suggested that the presence of potential IAS in the river catchment may also be considered as an existing constraint. The presence of potential IAS may be measured in a number of ways. For example, Panov et al. (2009) refers to the presence of alien species as biopollution defined as the introduction of alien species with noticeable effects on individuals, populations and communities of native species and/or resulting in adverse socioeconomic consequences (Elliott 2003; Panov et al. 2009). The pathway-specific biological contamination rate (PBCR) reflects the propagule pressure of alien species per introduction pathway, regardless of their potential negative impact (Panov et al. 2009). The vulnerability of an ecosystem to alien species establishment and damage may be defined by the biological contamination level (BCL) and integrated biological pollution risk (IBPR) respectively, using an assessment of already established alien species (Arbačiauskas et al. 2008; Olenin et al. 2007; Panov et al. 2009). However, this method cannot be applied to assess the future vulnerability of locations that will be subject to rehabilitation interventions. Panov et al. (2009) also defines a biological pollution risk (SBPR) index that estimates invasiveness based on the potential establishment, spread and impacts in a new environment. A measure of future vulnerability to species invasion could be defined using a

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combination of the PBCR and SBPR, however, this avenue was not explored by the authors. If potential IAS are acknowledged as an existing constraint, the ‘operational leitbild’ concept allows the potential threat of alien species to be recognised and incorporated into river rehabilitation planning. The proposed adaptations highlight the importance of transboundary issues such as the presence of IAS in the river catchment and the presence of pathways and vectors of introduction of new IAS, together with diffuse pollution and eutrophication that may influence the course of ecosystem development following rehabilitation interventions. The recognition of transboundary issues during the planning process requires that assessments are carried out to ascertain the level of risk they pose to the success of river rehabilitation interventions. Risk analyses will contribute to the development of an ‘operational leitbild’ which replaces the reference condition and encourages the development of achievable benchmarks that acknowledge the presence of socio-economic constraints.

7.4 Assessing the potential for alien species establishment as a result of rehabilitation interventions

7.4.1 River catchment scale horizon-scanning

The horizon-scanning method may be applied on the river catchment scale to risk prioritise alien species that may benefit from rehabilitation interventions. D. rostriformis bugensis has been chosen to illustrate how horizon-scanning and risk assessments can be applied on a river catchment scale to develop an ‘operational leitbild’ during river rehabilitation planning. D. rostriformis bugensis was not included in the risk prioritisation of the horizon- scanning for potential IAS (Chapter 3) as the species is already widespread in Western Europe and eradication measures are, therefore, not cost-effective . In general, technical issues that could potentially hinder interventions for the management of IAS in the Netherlands relate to access to private land, methods of elimination, fishing rights, sediment or soil removal, unintentional collection of other (native) species during sampling, the conservation status of areas colonised by IAS and protected species recorded in the same region as IAS (De Hoop et al. 2015). It is expected that if the horizon-scanning methodology is applied on a river catchment scale at temperate and subtropical locations where D. rostriformis bugensis does not already occur, the species would be classified as a high priority species. This is because of this species’ widespread introduced range in Western Europe (Section 5.4.1, Fig. 5.2), the presence of multiple vectors and pathways of introduction (Section 5.4.4, Fig. 5.3), and its identification as a high risk species in other regions. For example, relevant high risk classifications exist for the USA (from the temperate east to the subtropical south and west) and Canada (e.g., Roy et al. 2014a;Adams 2013; Province of British Columbia 2015).

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7.4.2 Risk prioritisation and assessment of the quagga mussel (Dreissena rostriformis bugensis)

Following risk prioritisation, a full risk analysis or rapid risk assessment may be implemented for high priority species if an existing risk categorisation is unavailable for the country within which the rehabilitation project is to be carried out. Risk assessments of D. rostriformis bugensis using the Belgian ISEIA and Harmonia+ risk assessment protocols (Chapter 4) indicated that the species poses a high ecological and socio-economic risk in the Netherlands (De Hoop et al. 2015). The impacts of D. rostriformis bugensis are similar to that of the zebra mussel (D. polymorpha) that has colonised all suitable habitats since it was first recorded in 1824. However, physiological differences between the two species, such as tolerance to silt and oligotrophic conditions (Matthews et al. 2014b), may result in the establishment of D. rostriformis bugensis in locations where D. polymorpha is currently absent, leading to an increase in the severity and geographical extent of impacts. Risk prioritisation and assessment of D. rostriformis bugensis revealed several major knowledge gaps that could influence the course of river rehabilitation interventions depending on project goals, such as (1) current distribution and dispersal mechanism in Western Europe and (2) the trophic transfer of metals. Our analysis of current distribution and dispersal mechanisms in Western Europe indicated that D. rostriformis bugensis has spread rapidly, utilizes a wide range of dispersal vectors and is already widespread in the Netherlands (Chapter 5). D. polymorpha is a species that is closely related to D. rostriformis bugensis and currently widespread in Europe. Therefore, rehabilitation interventions that increase the suitability of habitats may increase the risk of D. rostriformis bugensis establishment in Dutch rivers due to the presence of source populations and a high dispersal potential. The lack of knowledge concerning the consequences of a dominance shift between D. polymorpha and D. rostriformis bugensis on the trophic transfer of metals may introduce uncertainty if the goals of river rehabilitation are to encourage the return of predator species that are vulnerable to metal toxicity. Our analysis of metal concentrations in both mussel species indicated that a dominance shift to D. rostriformis bugensis may reduce the trophic transfer of nickel and copper in all water bodies and increase the trophic transfer of zinc in river water (Section 6.4.1). Therefore, river rehabilitation interventions that encourage the establishment of D. rostriformis bugensis may positively impact native species that are vulnerable to the effects of nickel and copper accumulation, and may be detrimental for native species that are vulnerable to the effects of zinc accumulation.

7.4.3 Implications of rehabilitation interventions for ecosystem invasibility

Following the identification of D. rostriformis bugensis as a high risk IAS that could negatively impact rehabilitation goals, river rehabilitation planners should examine proposed river rehabilitation interventions, and possible alternatives, to determine which interventions are least likely to result in D. rostriformis bugensis establishment. These considerations should take into account transboundary factors that will influence the risk of

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introduction of the species such as the location of source populations that are, or may become, hydrologically connected to the project area, and the presence of dispersal vectors such as inland shipping (Chapter 5). For example, the creation of a river side channel aimed at increasing habitat diversity may involve the reduction of current velocity and shipping related turbulence, and the introduction of different types of hard substrate (wood, stone etc.). D. rostriformis bugensis is a bivalve filter feeder with a wide environmental tolerance that uses byssal threads to attach and colonise hard substrates (Matthews et al. 2012d). Therefore, it is reasonable to assume that low current velocity, preferred by filter feeders, and the presence of hard substrate that allows byssal attachment, will increase the vulnerability of the river section to D. rostriformis bugensis invasion in the presence of source populations, introduction vectors, and limited competition from established species with similar habitat requirements.

7.4.4 Uncertainty

Inconsistencies in risk classifications between countries identified during the application of horizon-scanning and risk assessment methods such as ISEIA and Harmonia+ may be due to uncertainty intrinsic to the application of risk assessment protocols. Uncertainty relating to qualitative and semi-quantitative risk assessments were defined by Leung et al. (2012): (1) linguistic uncertainty, (2) stochasticity (also referred to as irreducible uncertainty or natural variation) and (3) epistemic uncertainty (also referred to as reducible uncertainty or incertitude). Linguistic uncertainty occurs because verbal and written communication is frequently open to interpretation, and even exact language may be deciphered in different ways by different people. Linguistic uncertainty may be increased when risk assessments are applied internationally due to the requirement of a common language that may be non- native to many contributors. Stochasticity results from temporal and spatial variability, and the probabilistic mechanisms that originate from this variability. Epistemic uncertainty reflects the level of available knowledge and its reliability relating to, e.g., sample size, surrogate measurements and observer error. Epistemic uncertainty may be sub-divided into parameter uncertainty, model (structural) uncertainty and data/observation error (Leung et al. 2012).

During risk assessment workshops using the ISEIA protocol, inter-assessor variability in risk scoring was often the result of imprecise language that increased linguistic uncertainty (Leuven et al. 2016). For example, vague terminology such as ‘remote places’, ‘rarely exceeds’, ‘highly fecund’ and able to ‘easily disperse’ requires interpretation that can vary between assessors. The Harmonia+ protocol reduces the influence of linguistic uncertainty by posing a single question that can be answered by choosing one out of five or six options on a Likert type scale with no associated criteria: inapplicable, very low, low, medium, high, very high.

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The scale of the area under assessment may introduce an element of stochasticity to the assessment process. Biological processes occurring on the habitat scale may be obscured or ignored due to the chosen assessment scale leading to greater predictive error. Moreover, the use of cases that are similar in biology or geography when direct evidence appears lacking, as advocated in Harmonia+ (D’hondt et al. 2015), may be problematic. This is because several species are known to expand to other habitat types once outside their native range (Wittenberg & Cock 2001; Verbrugge et al. 2012a), there is no universal explanation of successful exotic invasion into native communities (Dawson et al. 2015), and only limited research is available that links traits of particular species groups to potential invasiveness (e.g., Moravcová et al. 2015; Pyšek et al. 2015). Moreover, local context influences the local categorisation of risk. For example, in the Netherlands the coypu (Myocastor coypus) poses a high risk of socio-economic impact due to its burrowing behaviour that can damage dikes and therefore increase the risk of dike breaches and flooding. This risk is not relevant in other countries where dikes are not relied upon to maintain public safety.

A great source of epistemic uncertainty in semi-quantitative alien species risk assessments comes from observation and data error, in particular relating to data gaps, the intensity of data mining, and different methodological approaches (e.g. Gassó et al. 2010; Verbrugge et al. 2012a). Of the more than 12,000 European alien species registered in the DAISIE database, ecological impacts are only known for 1094 species (11%) and economic impacts for only 1347 species (13%) (Vilà et al. 2009; Hulme 2012), though other impacts may occur but have not yet been investigated or published. Since this research was conducted, the DAISIE database has been incorporated into the European Alien Species Information Network (EASIN), but not yet updated. Current risk assessments often suffer from insufficient data availability and may rely heavily on expert opinions and interpretations that carry their own set of uncertainties (Maguire 2004; Strubbe et al. 2011; Verbrugge et al. 2012a; Leung et al. 2012). The ISEIA protocol applies strict criteria to define risk categories which may result in greater instances of data deficiency. For example, criteria such as ‘natural dispersal rarely exceeds more than 1 km per year’ and ‘the alien species is known to cause local changes (< 80%) in population abundance, growth or distribution of one or several native species’ require specific information that is rarely included in literature resulting in the frequent application of expert judgement with a resulting increase in epistemic uncertainty. Expert opinion may be vulnerable to (1) focusing bias, where too much importance may be placed on certain preconceived attributes associated with invasiveness; (2) framing, where depending on how it is presented, different conclusions are drawn from the same information; (3) anchoring, where an initially formed opinion shapes subsequent judgement, even in the face of evidence to the contrary; and (4) confirmation bias, where information is searched for, or interpreted in a way that supports an initial preconception (Hastie & Dawes 2010; Hulme 2012).

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The Harmonia+ protocol does not allow a response that recognises data deficiency or the application of expert opinion, choosing instead to apply a certainty score to answers provided. Tull & Hawkins (1993) state that omitting the ‘don’t know’ category forces assessors to make a choice, potentially encouraging the occurrence of epistemic bias relating to expert opinion. Moreover, assessors may be inclined to select a ‘medium’ rating which will result in two misleading outcomes: (1) the impression will be given that more assessors have opinions than actually do, and (2) the mean and median will be shifted toward the middle of the scale (Friedman & Amoo 1999). Even the inclusion of a low confidence score to accompany the ‘medium’ rating may be misleading as it suggests that (weak) supporting evidence exists.

Model uncertainty is a sub-division of epistemic uncertainty that occurs as a result of methodological structure. The ISEIA and Harmonia+ approaches may introduce model uncertainty by (1) applying the precautionary principle, and (2) using a semi-quantitative approach. The precautionary principle, or taking the worst-case scenario when different scenarios are possible, is advocated in Harmonia+ (D’hondt et al. 2015) and ISEIA (Branquart et al. 2009), reflecting the requirements of the CBD (UNEP 2014a). Ironically, the application of the precautionary principle together with potential linguistic biases in relation to source information may exacerbate focussing and confirmation biases, anchoring, or framing by encouraging assessors to select information that portrays alien species in the worst possible light. Furthermore, D’hondt et al. (2015) suggest that choosing the maximum score from each category instead of the mean, while satisfying the precautionary principle, reduces the protocol’s discriminatory power by skewing the results of each module towards 1.0 as it is highly unlikely that a maximum score would not be applied to any single question. A reduction in discriminatory power and an accompanying increase in species classified as invasive may result in reduced management effectiveness if management budgets remain unchanged. In addition, application of the precautionary principle will have a greater effect for alien species that lack regional spread and impact data as knowledge gaps increase the need for reliance on expert knowledge. On the other hand, choosing the maximum score as part of a precautionary approach maintains the visibility of potential risks. Therefore, it is vital that the application of the precautionary principle is accompanied by a discussion of its potential implications for certainty of risk categorisations and protocol discriminatory power. Additionally, both the mean and the maximum scores can be made visible during the reporting stage to maximise transparency.

Both the ISEIA and Harmonia+ protocols are semi-quantitative methods that convert what is frequently qualitative data to a quantitative value to enable the calculation of a final risk score. The use of Likert type scales featuring normative cut-off thresholds means that small changes in the assessment (e.g., slightly different judgements of available data) may lead to changes in risk classification (Verbrugge et al. 2012a). Furthermore, this semi-quantitative approach hides the complexity and potentially variable quality of the supporting

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information, may obscure the application of expert opinion, and may give a false impression of legitimacy. However, the inclusion of normative aspects in non-native species risk assessments is unavoidable (Verbrugge et al. 2012a). Therefore, approaches that increase transparency by highlighting uncertainty are vital to the legitimacy of any risk assessment method.

The ISEIA protocol and Harmonia+ protocols use different approaches to express uncertainty. ISEIA does not require a quantification of uncertainty, but does allow for the transparent application of expert knowledge. Harmonia+ quantifies uncertainty but does not require that the level of influence of expert opinion on the final risk classification is reported. Inter-assessor variability and uncertainty may be reduced by applying a consensus method that involves the opinions of multiple experts. However, neither the ISEIA nor Harmonia+ protocols require that a consensus method is followed or that the area or level of expertise, or number of assessors that contributed to the risk assessment are reported. The consensus approach involves two stages where multiple assessors firstly complete a risk assessment individually by reviewing available evidence, and secondly discuss disagreements to determine whether they are based on linguistic or epistemic uncertainty and can be solved, or if they are genuine and unbiased differences in opinion (Leung et al. 2012; D’hondt et al. 2015). In order to increase the perceived validity of the risk classification, the discussions that led to a consensus should be reported, including information on the number and professional background of the assessors involved, and their experience with the assessment protocol. Moreover, choices with regards to non-standard weighting factors in Harmonia+ should be clearly reported as these are locally specific and may reduce the generalisability of results to other climatically similar regions.

7.4.5 Adherence to standards for the risk assessment of IAS of EU concern

To allow standardisation and reduce uncertainty in relation to alien species risk assessment in the EU, 14 minimum methodological standards for risk assessment methods were defined by Roy et al. (2014b) in a project funded by the European Commission. However, ISEIA and a former version of Harmonia+ did not meet all the requirements of these standards. The ISEIA protocol does not comply due to a lack of (1) questions relating to taxonomy, invasion history, distribution range (native and introduced), geographic scope and socio- economic benefits; (2) assessment of entry and establishment likelihoods; (3) distinction between intentional and unintentional spread; (4) direct assessment of ecosystem services; (5) attention to the effects of climate change; (6) assessment of adverse socio- economic impacts; (7) an explicit measure of uncertainty (Roy et al. 2014b). The updated Harmonia+ protocol improves on the ISEIA methodology and also includes socio-economic factors (agriculture, animal husbandry, human health), climate change in an assessment of future conditions, direct assessment of ecosystem services, and confidence levels. Unfortunately, the assessment of ecosystem services does not contribute to the overall risk score. However, the Harmonia+ protocol does not comply with the minimum standards set

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out by Roy et al. (2014b) due to a lack of (1) questions relating to invasion history, distribution range (native and introduced) and socio-economic benefits; (2) fixed procedure because users may organize the current scoring process as they wish (Roy et al. 2014b). This second point may be addressed by reporting user choices as part of the assessment results. Additionally, it is arguable whether the cost of possible management interventions should be considered without contrasting this with the financial cost resulting from the negative impacts of IAS establishment.

Of all the elements required for compliance, the incorporation of ecosystem services into the assessment presents potentially the greatest challenge due to their inclusion of both the ecological and socio-economic domains, and incompatibility with current methodological frameworks that assess impact. However, the application of ecosystem services to assess impact is advantageous as it allows invasions to be assessed at the ecosystem rather than species level, facilitates communication with interest groups, helps incorporate multiple perspectives, and highlights the multi-dimensional character of impacts (Binimelis et al. 2007; Simberloff et al. 2013; McLaughlan et al. 2014). Ecosystem services may be defined as the benefits natural ecosystems provide to human society, or, in other words, the ecosystem processes by which human life is maintained (Charles & Dukes 2007). They can be split into three types, (1) provisioning services, such as food (crops, livestock, fisheries, etc.), fibre (timber, cotton, silk, etc.), fuel, genetic resources, bio- chemicals/pharmaceuticals/natural medicines, and ornamental resources; (2) regulating services, such as pollination, climate regulation, water purification, soil stabilisation, disease regulation and flood mitigation; (3) cultural services, such as recreation and tourism, aesthetic beauty, spirituality, religion, ceremony and tradition; and (4) supporting services which are necessary for the production of all other ecosystem services, such as photosynthesis, primary production and nutrient cycling (Millennium Ecosystem Assessment Board 2005; Charles & Dukes 2007; Pejchar & Mooney 2009).

Luck et al. (2003) present a conceptual framework that categorises ecosystem services within “service-providing units” that are linked to species populations. A service-providing unit may be defined as ‘the component of biodiversity (collection of individuals) necessary to deliver a given ecosystem service at the level required by service beneficiaries’ (Luck et al. 2003). Service-providing units are sub-divided into functional components characterized by morphological, phenological or physiological traits (functional traits) that influence reproduction, survival and growth (Violle et al. 2007). IAS modify functional traits within one or more service-providing units leading to ecosystem impacts (Gilioli et al. 2014). Ecosystem services are regulated and may be defined by combinations of functional traits across trophic levels (trait-service clusters) (Díaz et al. 2007; Kremen et al. 2007; de Bello et al. 2010; Gilioli et al. 2014). Gilioli et al. (2014) applied the service-providing unit concept by determining the potential for IAS to interfere with functional traits categorised within service providing units at the individual species, population or community level

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using literature describing previous species invasions. Gilioli et al. (2014) then applied a standard index of five classes expressed in terms of percentage of losses to ecosystem service provision to characterise the intensity of impact of trait alteration on each service- providing unit.

Despite the existence of examples where the impacts of IAS have been assessed using the ecosystem services concept, many obstacles remain. Difficulties associated with quantification pose a major challenge (Pejchar & Mooney 2009). Research has focused on the regulating services that particularly plant species and functional diversity exert on biological processes such as nitrogen retention, primary production, stability and decomposition (Huston 1997; Schwartz et al. 2000; Díaz & Cabido 2001; Loreau et al. 2001; Tilman et al. 2001; Duffy 2002; Srivastava & Vellend 2005; Tilman et al. 2006; Gilioli et al. 2014). However, impacts on regulating services are rarely quantified but likely to be substantial, and impacts on cultural services, while resonating highly with stakeholders, are potentially the most difficult to evaluate (Pejchar & Mooney 2009). The ecosystem services concept encompasses both the ecological and socio-economic aspects of ecosystems (Binimelis et al. 2007). However, the majority of risk assessment systems categorise economic, environmental, and social impacts separately and will require significant methodological revision to allow the incorporation of ecosystem services. Moreover, current approaches to risk assessment involve the wide application of expert judgement, whereas impacts on ecosystem services should be assessed by a much broader group of assessors, including stakeholders (Binimelis et al. 2007).

7.4.6 Implications of IAS establishment for rehabilitation planning

The recognition that the risk of IAS establishment may change as a result of ecological rehabilitation poses a problem for the nature managers of river systems. Moreover, though not examined in this thesis, ecological rehabilitation of terrestrial, marine and lake habitats may also lead to an increase in the risk of alien species invasion. If a rapid assessment or risk assessment indicates that interventions may increase the likelihood of IAS establishment, should those interventions be abandoned altogether? IAS establishment may result in the development of novel ecosystems, i.e., ecosystems that develop when, as a result of human agency, new species occur in combinations and relative abundances that have not occurred previously (Hobbs et al. 2006). For example, some alien plant species exhibit a higher tolerance of eutrophic conditions and high temperatures than their native counterparts (Grutters et al. 2017). Moreover, the absence of source populations, socio- economic constraints, or the absence of hosts for certain life stages, may result in a lack of propagule pressure and interspecific competition from native species that would otherwise prevent the establishment of alien species. The establishment of alien species and the formation of novel ecosystems may occur if a lack of financial or technical resources prevents these types of transboundary problems from being addressed. Hobbs et al. (2006) argues that novel ecosystems should be accepted as it may be very difficult or costly to

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return such systems to their previous state. In this case, novel ecosystems should be managed in an effort to maximise beneficial changes and reduce the less beneficial aspects (Hobbs et al. 2006).

The acceptance and management of novel ecosystems has implications for the definition of rehabilitation success. In order to define success, it is important to understand which system characteristics are important in determining ecosystem recovery, and to what extent rehabilitation measures need to overcome barriers that prevent this recovery (Hobbs 2007). In other words, it is important to determine whether the presence of certain alien species in a novel ecosystem will hinder recovery. According to Scherer (1995), a restored environment can contain different species to the original, as long as the new species fill the niches and perform the functions of the original species. If this interpretation is accepted, functional ecological equilibrium, or functional equivalence, may become a criterion for rehabilitation success. Therefore, assuming suitable conditions for establishment, all alien species that are assessed as invasive or non-invasive may fill niches and perform functions of species representative of the pristine system that are unlikely to return. In this way, functional equilibrium may be achieved and non-invasive alien species may also contribute to the success of rehabilitation projects. Hobbs et al. (2009) acknowledges this view, stating that in some cases rehabilitation of the system to its original state may be possible but that for some more highly degraded systems, rehabilitation of ecosystem structure and / or function may be the only feasible alternative. Katz (1996) states that the notion of functional equivalence eliminates any direct argument for the preservation of natural entities within the system. However, if transboundary problems are intractable, a functional equivalence approach may be the only option that will allow the rehabilitation of ecosystem functioning, which may in turn support the return of elements of the reference species assemblage.

In ecosystems degraded to a degree where rehabilitation to a historical reference is no longer possible, beneficial outcomes that enhance ecosystem services, biodiversity conservation and ecological integrity may still be achievable (Chazdon 2008; Seastedt et al. 2008). Ewel and Putz (2004) describe alien species as potential allies in the rehabilitation of some landscapes if they provide essential ecological or socioeconomic services. For example, in the Netherlands a number of alien Impatiens species are recognised as important providers of nectar for various insect species (Matthews et al. 2015). The challenge for rehabilitation managers lies in judging whether or not certain IAS may be accepted as part of the rehabilitated or novel ecosystem (Ewel & Putz 2004), or whether the (re)introduction of native species could provide the functions lost. This is made particularly difficult in view of the potential lag time between introduction and impact of alien species that may extend to decades or more (Kowarik et al. 1995). Moreover, abundances of alien species are known to cycle from high to low which may give a distorted image of alien species establishment and spread during short-term monitoring efforts. Herein lie further

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opportunities for the application of a risk analysis approach on a spatial rather than individual species basis to assess the potential positive structural and / or functional contribution of alien species to the rehabilitation of ecosystems.

7.5 Conclusions

 An analysis of 41 river rehabilitation projects revealed that IAS were not explicitly considered during river rehabilitation planning and assessment.  Previous rehabilitation attempts such as those that have been undertaken in the Lower Rhine river have led to the facilitation of a number of alien species and IAS.  The use of horizon-scanning and risk assessment may contribute to the identification of potential IAS that may be introduced in river catchments and facilitated as a result of rehabilitation measures.  The results of the horizon-scanning for the Netherlands showed that freshwater and terrestrial animals, and terrestrial plants were most frequently classified as high risk, followed by marine animals and freshwater plants. The escape from confinement pathway was associated with the highest number of ecological-impact types attributed to high risk species, followed by the transport stowaway and transport contaminant pathways.  Horizon-scans for potential IAS have, to date, largely neglected the prioritisation of symbionts, parasites and commensals.  The quagga mussel (Dreissena rostriformis bugensis) was assessed and identified as a high risk species in the river catchments of the Netherlands and EU.  Four additional species were classified as high risk: Prussian carp (Carassius gibelio), common carp (Cyprinus carpio), pike-perch (Sander lucioperca), and fanwort (Cabomba caroliniana).  All but two out of the 12 species assessed with existing risk classifications for neighbouring countries were classified inconsistently (C. gibelio and D. rostriformis bugensis). Reasons for these inconsistencies could be the application of different risk assessment schemes, methodological application on a national rather than biogeographical scale, differences in the definition, application of criteria and habitat availability, and uncertainties that are intrinsic to risk assessment methodologies.  Risk assessment of D. rostriformis bugensis revealed knowledge gaps relating to dispersal factors and ecological effects that reduced the certainty of its ecological risk classification.  D. rostriformis bugensis has demonstrated a rapid and continued range expansion in Western Europe. A dominance shift from zebra mussels (Dreissena polymorpha) to D. rostriformis bugensis may reduce the trophic transfer of nickel and copper in all water bodies and increase the trophic transfer of zinc in river water.

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7.6 Recommendations

 The potential for the unintentional introduction and establishment of alien species as a result of the rehabilitation of aquatic habitats should be assessed as an integral part of rehabilitation planning.  The species identified as high priority during the horizon-scanning described in this thesis may be subject to a full ecological risk assessment which will contribute to the design of rehabilitation measures linked to a set of achievable goals (‘operational leitbild’).  Risk prioritisations and assessments should be spatially delineated on a biogeographical rather than political or administrative scale due to the need for a consistent climate and habitat match of IAS to support risk classifications.  Future research should focus on the potential contribution that alien species make towards the functioning of (novel) ecosystems. Functional equivalence may form the basis for acceptance of certain alien species in an ‘operational leitbild’ that defines the goals of habitat rehabilitation and determines rehabilitation interventions.  Future horizon-scans for potential IAS should include prioritisations of symbionts, parasites and commensals.  Future research should be focussed on the feasibility of undertaking area specific assessments of multiple species and pathways, rather than alien species specific risk assessments. Assessments of the vulnerability of specific areas to species invasion highlight invasion pathways that increase the risk of invasion of one or more IAS. This in turn may facilitate the design of area specific prevention measures, early warning and rapid response, and management approaches that target multiple IAS.  The reporting of potential sources of inconsistency between available risk classifications should be integrated into risk assessment methodology. This approach is vital to the legitimacy of any risk assessment method and will increase acceptance among decision makers, nature managers and stakeholders.

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Appendices

Appendix 1: Categorisation of pathways for the introduction of alien species (adapted from UNEP 2014b).

Category Subcategory

RELEASE IN Biological control NATURE Erosion control/ dune stabilization (windbreaks, hedges, ...) Fishery in the wild (including game fishing) Hunting Landscape/flora/fauna “improvement” in the wild Introduction for conservation purposes or wildlife management Release in nature for use (other than above, e.g., fur, transport, medical use) Other intentional release ESCAPE Agriculture (including Biofuel feedstocks) Aquaculture / mariculture FROM Botanical garden/zoo/aquaria (excluding domestic aquaria) CONFINEMENT Pet/aquarium/terrarium species (including live food for such species ) Farmed animals (including animals left under limited control) Forestry (including afforestation or reforestation) Fur farms Horticulture Ornamental purpose other than horticulture Research and ex-situ breeding (in facilities) Live food and live bait Other escape from confinement

TRANSPORT – Contaminant nursery material Contaminated bait CONTAMINANT Food contaminant (including of live food) Contaminant on animals (except parasites, species transported by host/vector) Parasites on animals (including species transported by host and vector) Contaminant on plants (except parasites, species transported by host/vector) Parasites on plants (including species transported by host and vector) Seed contaminant Timber trade Transportation of habitat material (soil, vegetation,…) TRANSPORT - Angling/fishing equipment STOWAWAY Container/bulk Hitchhikers in or on airplane Hitchhikers on ship/boat (excluding ballast water and hull fouling) Machinery/equipment People and their luggage/equipment (in particular tourism) Organic packing material, in particular wood packaging Ship/boat ballast water Ship/boat hull fouling Vehicles (car, train, …) Other means of transport CORRIDOR Interconnected waterways/basins/seas Tunnels and land bridges

UNAIDED Natural dispersal across borders of invasive alien species that have been introduced outside borders via other pathway categories

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Appendix 2: List of potential IAS amenable to prevention and early eradication measures in the Netherlands.

Scientific name Species group Order Common name(s)

Acacia dealbata plants Fabales Silver wattle, blue wattle Acaena novae-zealandia plants Rosales Piri-piri-bur Acipenser baerii fish Acipenseriformes Siberian sturgeon Agrilus planipennis insects Coleoptera Emerald ash borer Akebia quinata plants Ranunculales Five-leaf Alopex lagopus mammals Carnivora Arctic fox Ameiurus melas fish Siluriformes Black bullhead Amorpha fruticosa plants Fabales False indigo Aponogeton distachyos plants Hydrocharitales Cape-pondweed Arthurdendyus triangulatus worm Turbellaria New Zealand flatworm Arundo donax plants Angiospermae Giant reed Asterias amurensis sea star Forcipulatida Japanese seastar, northern Pacific seastar Baccharis halimifolia plants Asterales Salt bush, eastern baccharis Bellamya chinensis* molluscs Architaenioglossa Chinese mystery snail Bursaphelenchus xylophilus worm Eudicotyledoneae Pinewood nematode Callosciurus erythraeus mammals Rodentia Pallas's squirrel, red-bellied tree squirrel Callosciurus finlaysonii mammals Rodentia Finlayson's squirrel Carpobrotus edulis plants Caryophyllales Hottentot fig Castor canadensis mammals Rodentia Canadian beaver Cercopagis pengoi crustacean Diplostraca Fish-hook waterflea Cervus nippon mammals Cetartiodactyla Sika deer Coptotermes formosanus insects Blattodea Formosan subterranean termite Cortaderia selloana plants Poales Pampas grass

Corvus splendens birds Passeriformes Indian house crow Cotoneaster horizontalis plants Rosales Wall cotoneaster, rockspray Cotoneaster dammeri plants Rosales Bearberry cotoneaster Crocosmia x crocosmiiflora plants Asparagales Montbretia Cynomys ludovicianus mammals Rodentia Black-tailed prairie dog Cyprinus carpio x Carassius sp.* fish Cypriniformes Crosscarp Cytisus striatus plants Fabales Hairy-fruited broom, Portuguese broom Echinocystis lobata plants Cucurbitales Wild cucumber, wild balsam apple Eichhornia crassipes plants Commelinales Water hyacinth Elaphe shrenckii* reptile Squamata Amur ratsnake, Siberian ratsnake Elaphe spp.* reptile Squamata Asian ratsnakes Elodea callitrichoides plants Alismatales South American waterweed Felis bengalensis mammals Carnivora Leopard cat Gaultheria mucronata plants Ericales Prickly heath Gunnera tinctoria plants Gunnerales Giant-rhubarb, Chilean rhubarb, Chilean gunnera Gyrodactylus salaris worm Gyrodactylidea Salmon fluke Heracleum persicum* plants Apiales Golpar, Persian hogweed Heracleum sosnowskyi plants Apiales Sosnowski's hogweed Homarus americanus crustacean Decapoda American lobster Hydrochoerus hydrochaeris mammals Rodentia Capybara *Species addition based on expert judgement

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Appendix 2: List of potential IAS amenable to prevention and early eradication measures in the Netherlands (cont.).

Scientific name Species group Order Common name(s) Hydropotes inermis mammals Artiodactyla Chinese water deer Lepomis cyanellus fish Perciformes Green sunfish Lepomis macrochirus fish Perciformes Bluegill Lithobates catesbeianus amphibian Anura American bullfrog Lonicera japonica plants Dipsacales Japanese honeysuckle Ludwigia peploides plants Myrtales Floating water-primrose Lysichiton americanus plants Alismatales American skunk cabbage Mephitis mephitis* mammals Carnivora Striped skunk Micropterus dolomieu fish Perciformes Smallmouth bass Micropterus salmoides fish Perciformes Largemouth black bass Morone americana fish Perciformes White bass Muntiacus reevesi mammals Artiodactyla Chinese muntjac Neogobius gymnotrachelus fish Perciformes Racer goby Obesogammarus obesus crustacean Amphipoda Scud Orconectes rusticus crustacean Decapoda Rusty crayfish Oxyura jamaicensis birds Anseriformes Ruddy duck Paralithodes camtschaticus crustacean Decapoda Red king crab Paspalum distichum plants Poales Knotgrass, water finger-grass Perccottus glenii fish Perciformes Rotan, amur sleeper Pileolaria berkeleyana worm Sabellida Polychaete tubeworm Pimephales promelas fish Cypriniformes Fathead minnow Pomoxis annularis fish Perciformes White crappie Pomoxis nigromaculatus fish Perciformes Black crappie Pontogammarus robustoides crustacean Amphipoda Ponto-Caspian shrimp, scud Procyon lotor mammals Carnivora Raccoon Rapana venosa molluscs Neogastropoda Rapa whelk Asian wild raspberry, cheeseberry, yellow Himalayan Rubus ellipticus plants Rosales raspberry, yellow raspberry Sarracenia purpurea plants Ericales Pitcherplant Sasa palmata plants Poales Broad-leaved bamboo Sciurus carolinensis mammals Rodentia Grey squirrel Sciurus lis* mammals Rodentia Japanese squirrel Sciurus niger* mammals Rodentia Fox squirrel Sinanodonta woodiana* molluscs Unionoida Swan mussel Solidago nemoralis plants Asterales Gray goldenrod Spiraea alba plants Rosales Pale bridewort, meadowsweet Tamias sibiricus mammals Rodentia Siberian chipmunk Tamiasciurus hudsonicus* mammals Rodentia American red squirrel Threskiornis aethiopicus birds Ciconiiformes Sacred ibis Trachemys scripta elegans reptile Chelonii Red-eared terrapin Triturus carnifex* amphibian Urodela Italian crested newt Triturus marmoratus* amphibian Urodela Marbled newt Vespa velutina insects Hymenoptera Asian hornet Watersipora subtorquata bryozoan Cheilostomatida Encrusting bryozoan Xenopus laevis amphibian Anura African clawed toad Heracleum persicum* plants Apiales Golpar, Persian hogweed *Species addition based on expert judgement

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Appendix 3: Overview of information sources.

Organisation / database Web address Invasive Species Compendium http://www.cabi.org/isc/ Global Invasive Species Database http://www.issg.org/database/welcome/ Web of Science http://apps.isiknowledge.com DAISIE (European Alien Species www.europe-aliens.org/ Information) European and Mediterranean Plant http://www.eppo.int/INVASIVE_PLANTS/ias_plants.htm Protection Agency (EPPO) NOBANIS http://www.nobanis.org/ Fishbase www.fishbase.org National Databank of Flora and Fauna https://ndff-ecogrid.nl/ Dutch Register of Catches of Sports Fish www.vangstenregistratie.nl Waarneming.nl www.waarneming.nl Netherlands Food and Consumer Product https://www.vwa.nl/onderwerpen/ongewenste-uitheemse-planten/dossier/invasieve- Safety Authority (NVWA) exoten/risicobeoordelingen-reactieperiode/risicobeoordelingsrapporten Macrofauna Newsletter http://www.helpdeskwater.nl/onderwerpen/monitoring/ecologie/macrofaunanieuws/ Werkgroep Exoten http://www.werkgroepexoten.nl Nederlands Soortenregister www.nederlandsesoorten.nl GB Non-Native Species Secretariat http://www.nonnativespecies.org/home/index.cfm

Invasive Species in Belgium http://ias.biodiversity.be/species/all French Museum of Natural History http://inpn.mnhn.fr Danish Nature Agency http://www.naturstyrelsen.dk/Naturbeskyttelse/invasivearter/Arter/Sortlisten/ Great Lakes Aquatic Nonindigenous http://www.glerl.noaa.gov/res/Programs/glansis/glansis.html Species Information System (GLANSIS) Integrated Taxonomic Information System http://www.itis.gov/ (ITIS) Google www.google.com

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Appendix 4: Overview of alien macrophyte species identified for risk analysis.

Latin name1 Common Native range Alien range name1 Cabomba Fanwort South America and eastern Austria, Belgium, France, Germany, Hungary, caroliniana North America2 Netherlands, Serbia, Sweden, UK, China, India, Japan, Papua New Guinea, Mexico, Australia, New Caledonia, New Zealand2 Egeria densa Brazilian Argentina, Brazil, Uruguay2 Algeria, Kenya, South Africa, Japan, Indonesia, Austria, waterweed Azores, Belgium, France, Germany, Hungary, Italy, Netherlands, Spain, Switzerland, UK, Canada Costa Rica, Cuba, El Salvador, Guadeloupe, Jamaica, Martinique, Nicaragua, Puerto Rico. Argentina, Bolivia, Brazil, Chile, Australia, New Zealand, Polynesia2 Lagarosiphon Curly South Africa2 New Zealand, Australia and in Europe: Austria, Belgium, major waterweed France, Germany, Hungary, Ireland, Italy, Netherlands, Spain, Switzerland, United Kingdom2 Vallisneria Tapegrass Likely native to tropical and Albania, Austria, Belgium, Bulgaria, Croatia, Cuba, spiralis sub-tropical Asia, Southern Czech Republic, Denmark, England, France, Germany, Europe and Northern Africa3 Greece, Jamaica, Luxembourg, Macedonia, Moldova, Montenegro, Netherlands, Poland, Romania, Russia, Serbia, Spain, USA (Hawaii and Texas)4 1According to species 2000; 2Q-bank (2016); 3Hussner & Lösch (2005), Les et al. (2008), Hussner (2012); 4Les et al. (2008); Hussner (2012); Lowden (1982); Dutrarte et al. (1997); Hussner & Lösch (2005); Van Ooststroom & Reichgelt (1961); Hutorowicz & Hutorowicz (2008); Katsman & Kuchkina (2010); Staples et al. (2003).

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Appendix 5: Overview of alien mollusc species identified for risk analysis.

Latin name1 Common name1 Native range Alien range Dreissena Quagga mussel Southern Bug and Dnieper rivers in North America, Romania, the Netherlands, rostriformis the Ukraine2 Germany, France, Belgium, England3 bugensis Bellamya Chinese mystery China, Taiwan, Korea, Japan, North America, the Netherlands and Belgium5 chinensis snail Indonesia, Burma, the Philippines, Russia, Thailand and Vietnam4 1According to species 2000; 2Son (2007), Van der Velde et al. (2010); 3May & Marsden (1992), Micu & Telembici (2004), Popa & Popa (2006), Matthews et al. (2014b), Molloy et al. (2007), Van der Velde & Platvoet (2007), Bij de Vaate & Beisel (2011), Sablon et al. (2010), Aldridge et al. (2014); 4Chiu et al. (2002), Global Invasive Species Database (2011); 5Collas et al. (2017); Karatayev et al. (2009); Global Invasive Species Database (2011), Van den Neucker et al. (2017).

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Appendix 6: Overview of alien fish species identified for risk analysis.

Scientific name1 Common Native range Alien range name1 Carassius gibelio Prussian carp Difficult to define due to long history of France, UK, Germany, Denmark, trade and taxonomic ambiguity2 Belgium, the Netherlands3 Coregonus albula Vendace Baltic catchment, lakes of the upper Northern and central Volga, White Sea basin, North Sea Germany and Poland, Finland, area basin east of the Elbe drainage, Gulf of bordering Norway and Russia, the Finland, Gulf of Bothnia (northernmost Netherlands5 part),the United Kingdom (four lakes)4 Ctenopharyngodon Grass carp Middle and lower sections of East Asian Introduced to over 50 idella rivers and ponds countries for aquaculture or weed control < 1000 metres above sea level and 23° e.g. western and south-eastern Europe, to 53° N latitude6 India, USA, Mexico, Japan7 Cyprinus carpio Common carp Eastern Europe to central Asia in the Introduced throughout the world. In basins of the Black, Caspian and Aral Europe the species has been domesticated Seas8 since the Middle Ages9 Cyprinus carpio X Hybrid cross Not applicable Farmed across Eurasia in aquaculture. Carassius spp. carp Common carp and Carassius spp. are reported to hybridize naturally in Europe, Asia, North America and Australia10 Leuciscus aspius Asp Ponto-Caspian region towards central Germany, the Netherlands, Belgium, Europe, southern Scandinavia, Danube Northern Russia, Kazakhstan11 drainage catchment, north western Turkey5 Oncorhynchus Rainbow trout United States pacific coast from Alaska Introduced to over a hundred different mykiss to Mexico, the Pacific Ocean, the countries in Europe, Asia, Africa, eastern coast of Asia12 Oceania, North, Central and South America12 Romanogobio Northern River catchments that flow into the The Netherlands, Germany13 belingi whitefin Baltic , Black and Caspian Seas. gudgeon Possibly the Elbe, Rhine and Meuse river catchments13 Salvelinus alpinus Arctic charr Scandinavia, Canada, France, Serbia, Kerguelen Islands15 Russia, Iceland, Greenland, the USA, UK, and Ireland. Pre-alpine and high- altitude lakes in the Alps14 Salvelinus fontinalis Brook charr Eastern Canada from Newfoundland to Widely introduced in North and South the western side of Hudson Bay; south America, Europe, Asia, in the Atlantic, Great Lakes, and Oceania and Africa12 Mississippi River catchments to Minnesota and northern Georgia16 Sander lucioperca Pike-perch Western Asia and northern and eastern Western Europe, including the UK, Europe including the Danube river Germany, Iberian Peninsula, Italy17 catchment5 Umbra pygmaea Eastern mud- USA18 Germany, Belgium, the Netherlands, minnow France, Poland, Denmark19 1According to species 2000; 2Szczerbowski (2001), Kottelat & Freyhof (2007), Kalous & Knytl (2011), FAO (2013); 3FAO (2013); 4Kottelat & Freyhof (2007), Elliot & Bell (2011) , Soes (2009); 5Kottelat & Freyhof (2007), Amundsen et al. (2013); Martijn Schiphouwer (unpublished data); 6Bíró (1999); 7Bíró (1999), FAO (2013), Jankovic (1998); 8Barus et al. (2002), Kottelat & Freyhof (2007); 9De Wilt & Van Emmerik (2008); 10Crunkilton (1977), Barus et al. (2002), Hänfling et al. (2005), Maes et al. (2007), Haynes et al. (2012); 11Kottelat & Freyhof (2007), FAO (2013), De Nie (1996), NDFF/RAVON (2013), Pawlowski et al. (2012); 12Soes & Broeckx (2010); 13Naseka et al. (1999), Freyhof et al. (2000); 14Klemetsen et al. (2003), Maitland et al. (2007), Achleitner et al. (2009); 15Jamet (1995), Machino (1995), Klemetsen et al. (2003), Lenhardt et al. (2010), Lecomte et al. (2013); Verreycken et al. (2007); 16Page & Burr (1991); 17Kottelat & Freyhof (2007), FAO (2013); 18Froese & Pauly (2009); 19Verreycken et al. (2010).

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Summary Samenvatting Acknowledgements Curriculum vitae List of publications

Summary

Summary

Invasive alien species (IAS) are an important threat to global ecosystems, having major ecological and socio-economic impacts. In general, after loss of habitat, invasive species are the second leading cause of global biodiversity loss. Both the Convention on Biological Diversity (2014) and the new European Union (EU) Regulation on Invasive Alien Species advocate the prevention of IAS introduction, early warning and rapid response in case of newly recorded IAS, and the management of established IAS. Therefore, the identification of effective methods for risk prioritisation and assessment of alien species in order to target preventative, early eradication and control measures on national, international and global scales has become increasingly important.

River rehabilitation often involves re-shaping the riverine landscape with the aim of reflecting natural conditions. However, large scale morphological change may reduce ecological integrity by encouraging the establishment of alien species. Moreover, transboundary factors such as hydrological connectivity between river catchments and the presence of source populations, occurring within a river catchment, but outside the spatial boundaries of rehabilitation projects, may play a significant role in the introduction and establishment of IAS following rehabilitation interventions. For example, if diffuse transboundary pollution and eutrophication remain following rehabilitation efforts, increases in colonization by IAS may occur as a result of local habitat quality improvements. Indeed, IAS establishment as a result of rehabilitation interventions has been observed in practice. In south-eastern Australia the removal of Salix sp. from riparian areas increased the likelihood of spread of the invasive aquatic grass Glyceria maxima, and rehabilitation efforts in the Hudson River in North America led to widespread IAS establishment. However, past and present approaches to river rehabilitation planning have tended to frame project aims around existing problems within limited spatial boundaries, have ignored transboundary issues, and failed to predict future change that may lead to the introduction of alien species. Current European guidelines on river rehabilitation planning largely ignore the potential for the establishment of IAS as a result of interventions aimed at improving hydro-morphological quality, and refer only briefly to the management of alien species present prior to rehabilitation interventions. Moreover, the European Union’s Water Framework Directive (WFD) that aims to improve the ecological quality of the Union’s freshwater bodies, omits any reference to alien species.

The central aim of this study is firstly to reaffirm the observation that invasive species are not considered during river rehabilitation planning and assessment, and secondly to provide a framework for the assessment of potential risks that alien species pose to the achievement of river rehabilitation goals. Chapter 2 comprises a literature analysis of 41 river rehabilitation projects to assess the short-term (5 years) ability of indicator groups to demonstrate progress towards river rehabilitation goals. A questionnaire was developed to investigate river manager’s interpretation of rehabilitation success. Three open questions in particular gave opportunities for nature managers to describe how invasive species were

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considered in river rehabilitation planning and project evaluation. However, none of the 54 respondents who originated from Germany, the Netherlands and the United Kingdom referred to alien species in any of the responses given, suggesting a lack of awareness among practitioners of the implications of IAS to the goals of river rehabilitation projects.

The recognition of transboundary issues during river rehabilitation planning requires that assessments are carried out to ascertain the risk they pose to success. Horizon-scanning identifies and prioritises potential IAS, identifying the most important transboundary pathways and vectors of IAS introduction to a target region (project area). It therefore provides a starting point for the design of preventative, early identification and rapid action measures for effective management of IAS during river rehabilitation planning. Application of the method to the Netherlands (Chapter 3) produced a list of potential IAS revealing that the international trade in plants and animals is the most important pathway for the introduction of IAS nationally. D. rostriformis bugensis was chosen to illustrate how horizon-scanning and risk assessments can be applied on a river catchment scale to contribute to the development of an achievable set of rehabilitation objectives during river rehabilitation planning which acknowledges existing socio-economic constraints (‘operational leitbild’). It is expected that D. rostriformis bugensis would be identified as a high priority species during a horizon-scan due to the limited spatial scale of most rehabilitation projects, the presence of multiple vectors and pathways of introduction, and the existence of many high risk categorisations for this species.

Following risk prioritisation, a full risk analysis or rapid risk assessment may be carried out for high priority species. Risk analyses should be based on a balanced assessment between the potential contribution of alien species to ecosystem functioning in novel ecosystems and potential threats to the achievement of rehabilitation goals. Chapter 4 presents the risk classifications of 18 aquatic alien species for the Netherlands, compares these with risk classifications made for neighbouring, biogeographically similar countries within the EU, and provides explanations for inconsistencies between different risk classifications. Risk assessments of D. rostriformis bugensis using the Belgian ISEIA and Harmonia+ risk assessment protocols indicated that the species poses a high ecological and socio-economic risk in the Netherlands. Four additional species were classified as high risk: Prussian carp (Carassius gibelio), common carp (Cyprinus carpio), pike-perch (Sander lucioperca), and fanwort (Cabomba caroliniana). Of the 12 species with existing risk classifications for neighbouring countries, all but two of the assessed species (C. gibelio and D. rostriformis bugensis) were classified inconsistently. Reasons for these inconsistencies could be the application of different risk assessment schemes, application on a national rather than biogeographical scale, differences in the definition, application of criteria and habitat availability, and uncertainties that are intrinsic to risk assessment methodologies.

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Risk prioritisation and assessment of D. rostriformis bugensis revealed several major knowledge gaps that could influence the course of river rehabilitation interventions depending on project goals, such as (1) the current distribution and dispersal mechanisms of D. rostriformis bugensis in Western Europe and (2) the trophic transfer of metals from D. rostriformis bugensis to predator species. Our analysis of current distribution and dispersal mechanisms in Western Europe indicated that D. rostriformis bugensis has spread rapidly, utilizes a wide range of dispersal vectors, and is widespread in the Netherlands and many parts of Western Europe (Chapter 5). Due to the wide availability of source populations and high dispersal potential, rehabilitation interventions that increase the suitability of habitats for D. rostriformis bugensis will likely increase the risk of D. rostriformis bugensis spread. The zebra mussel (Dreissena polymorpha) is a species that is closely related to D. rostriformis bugensis and currently widespread in the Netherlands. The lack of knowledge concerning the consequences of a dominance shift between D. polymorpha and D. rostriformis bugensis on the trophic transfer of metals may introduce uncertainty if the goals of river rehabilitation target the return of predator species that are vulnerable to metal toxicity. Our analysis of metal concentrations in both mussel species indicated that a dominance shift to D. rostriformis bugensis may reduce the trophic transfer of nickel and copper in all water bodies and increase the trophic transfer of zinc in river water in the Netherlands (Section 6.4.1). Therefore, river rehabilitation interventions that encourage the establishment of D. rostriformis bugensis may benefit target species that are vulnerable to nickel and copper toxicity and may be detrimental for target species that are vulnerable to zinc.

Following risk analyses, the risk of establishment of IAS as a result of rehabilitation interventions should be incorporated into the formulation of an ‘operational leitbild’ or achievable set of rehabilitation objectives. River rehabilitation planners should examine proposed river rehabilitation interventions, and possible alternatives, to determine which interventions are least likely to result in the establishment of IAS. For example, a reduction in current velocity and the introduction of hard substrate as part of rehabilitation measures may increase the vulnerability of a river section to D. rostriformis bugensis invasion in the presence of source populations, introduction pathways and transport vectors. Following the implementation of rehabilitation measures based on the ‘operational leitbild’, monitoring to measure the level of rehabilitation success should take into account lag times and fluctuating abundance that are characteristic of IAS establishment and spread.

Future research should focus on the potential contribution that alien species make towards the functioning of novel ecosystems. Functional equivalence may form the basis for acceptance of certain alien species in an ‘operational leitbild’ that defines the goals of habitat rehabilitation and determines rehabilitation interventions. It is recommended that horizon-scanning and risk assessments are spatially delineated on a biogeographical rather than political or administrative scale due to the need for a consistent habitat and climate

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match to support risk classifications. Moreover, future horizon-scans for potential IAS should include prioritisations of symbionts, parasites and commensals as these groups have been largely neglected. Future approaches could include area specific, rather than species specific risk assessments. Area specific assessments of the vulnerability of an area to species invasion highlight invasion pathways that increase the risk of invasion of one or more alien species. This in turn may facilitate the design of area specific prevention, early warning and rapid response, and management measures that target multiple IAS. In addition, the reporting of potential sources of inconsistency between risk classifications should be integrated into risk assessment methodology. This approach is vital to the legitimacy of any risk assessment method and will increase acceptance among decision makers, nature managers and stakeholders.

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Samenvatting

Samenvatting

Invasieve exoten vormen op wereldschaal een belangrijke bedreiging voor biodiversiteit en ecosysteemfuncties met grote veranderingen in ecosysteemdiensten, socio-economische problemen en risico’s voor volksgezondheid als gevolg. Zowel het biodiversiteitverdrag van de Verenigde Naties als de nieuwe Europese Unie (EU) verordening met betrekking tot invasieve exoten benadrukken het belang van preventie van introductie van exoten met vroegtijdige signalering en snelle interventie als belangrijke maatregelen in het geval van vestiging van nieuwe invasieve soorten. Daarom wordt het steeds belangrijker om effectieve preventie, vroegtijdige uitroeiing en beheersmaatregelen op nationaal, internationaal en wereldniveau voor risicovolle soorten vast te stellen. Daarvoor zijn nieuwe prioriterings- en risicobeoordelingsmethoden nodig.

De geomorfologie van rivieren wordt vaak gewijzigd als onderdeel van rivierherstelprojecten waarbij natuurlijke referenties worden gebruikt om oorspronkelijke biotopen terug te brengen. Grootschalige veranderingen in de riviermorfologie kunnen de vestiging van invasieve exoten faciliteren en daardoor de ecologische integriteit beïnvloeden. Wijzigingen in de hydrologische connectiviteit tussen stroomgebieden en populatiebronnen of de aanpak van grensoverschrijdende watervervuiling en eutrofiëring als gevolg van rivierherstelmaatregelen spelen een belangrijke rol bij de introductie en dispersie van exoten. Invasieve exoten kunnen zich soms bijvoorbeeld massaal vestigen na verbetering van de lokale habitatkwaliteit. Deze onbedoelde vestiging van exoten na bepaalde herstelmaatregelen is in de praktijk al waargenomen. In Zuid-Oost Australië is de toenemende verspreiding van het invasieve aquatische liesgras Glyceria maxima waarschijnlijk mede veroorzaakt door het verwijderen van wilgen. Herstelmaatregelen in het Noord-Amerikaanse rivierstroomgebied van de Hudson leidden tot het wijdverbreid vestigen van invasieve exoten. Met de huidige rivierherstelmethoden worden milieuproblemen meestal binnen bepaalde ruimtelijke grenzen aangepakt maar worden grensoverschrijdende problemen nog vaak genegeerd. Daarbij wordt de vestiging van invasieve exoten als resultaat van herstelmaateregelen niet voorspeld. De huidige Europese richtlijnen met betrekking tot rivierherstel besteden weinig aandacht aan de mogelijke vestiging van exoten als gevolg van herstelmaatregelen. Naar het beheer van al gevestigde exoten in het herstelgebied wordt slechts kort verwezen. De EU Kaderrichtlijn Water (KRW) biedt geen expliciete verwijzing naar invasieve exoten in verband met verbetering van de waterkwaliteit of ecologische status.

Het voorliggende proefschrift bevestigt dat tijdens de planning van rivierherstelprojecten weinig aandacht aan exoten is besteed. Daarom worden richtlijnen voorgesteld voor het beoordelen van potentiële risico’s die exoten vormen voor het behalen van natuurdoelen. Hoofdstuk 2 bestaat uit een literatuuranalyse van 41 rivierherstelprojecten om indicatoren voor ecologisch herstel binnen korte termijn (5 jaar) te identificeren. Er is een vragenlijst ontwikkeld om de definities en interpretaties van de term ‘herstelsucces’ door natuurbeheerders te onderzoeken. Hiervoor zijn drie open vragen aan 54 natuurbeheerders

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uit Duitsland, Nederland en het Verenigd Koninkrijk gesteld. Geen van de ondervraagden verwees naar exoten met betrekking tot het behalen van projectdoelen.

Erkenning van grensoverschrijdende problemen, zoals de introductie van invasieve exoten, vereist een beoordeling van de risico’s van deze soorten voor het succes van ecologische herstelmaatregelen. Horizonscanning is een beoordelingsmethode om potentiële invasieve exoten in kaart te brengen en te prioriteren. Bovendien worden met deze methode de belangrijkste introductieroutes en vectoren voor een bepaalde regio of projectgebied vastgesteld. Horizonscanning biedt uitgangspunten voor het effectieve beheer van invasieve exoten (preventie, vroege signalering en snel ingrijpen) als onderdeel van de planning van rivierherstel. Na toepassing van de horizonscanningmethode in Nederland (Hoofdstuk 3) is een lijst van potentiële invasieve exoten samengesteld. Uit nadere analyse is gebleken dat voor Nederland de internationale planten- en dierenhandel de belangrijkste introductieroute vormt. De quaggamossel (Dreissena rostriformis bugensis) is vervolgens gebruikt als voorbeeld om te laten zien hoe in het rivierherstel exoten binnen een stroomgebied kunnen worden beoordeeld met horizonscanningmethoden en risicoanalyses. De resultaten van betreffende analyses zijn bruikbaar om haalbare doelen voor rivierherstel te formuleren, waarbij bestaande socio-economische randvoorwaarden worden meegenomen (‘operationeel leitbild’). In een horizonscanning zou de quaggamossel waarschijnlijk als een soort met een hoog milieurisico zijn beoordeeld vanwege de beperkte ruimtelijke schaal van de meeste herstelprojecten, de diversiteit van bestaande vectoren en introductieroutes en het hoge aantal ‘hoog risico’ classificaties uit andere regio’s.

Soorten met een hoog risico worden vervolgens uitgebreider beoordeeld via een snelle risicobeoordeling of een volledige risicoanalyse. Tijdens een volledige risicoanalyse is een weging nodig van de mogelijke bijdrage van de exoot aan het functioneren van ecosystemen en de haalbaarheid van hersteldoelen. Hoofdstuk 4 geeft een overzicht van risicoclassificaties van 18 soorten exoten voor Nederland. Een vergelijking wordt gemaakt tussen risicoclassificaties voor Nederland en qua klimaat en biogeografie vergelijkbare gebieden van omringende EU landen. Uit risicoanalyses met de ISEIA en Harmonia+ protocollen blijkt dat de quaggamossel een hoog ecologisch en socio-economisch risico vormt voor Nederland. Tevens zijn vier andere soorten beoordeeld als exoten met een hoog risico: giebel (Carassius gibelio), karper (Cyprinus carpio), snoekbaars (Sander lucioperca) en waterwaaier (Cabomba caroliniana). Twee van de 12 exoten met bestaande risicoclassificaties uit andere landen zijn met Nederlandse risicoanalyses consistent beoordeeld (C. gibelio en D. rostriformis bugensis). Mogelijke verklaringen voor inconsistentie tussen risicoclassificaties voor de overige soorten zijn het toepassen van verschillende risicobeoordelingsprotocollen, gebruik daarvan op een nationale in plaats van biogeografische schaal, verschillen in de definitie en toepassing van beoordelingscriteria, verschillen in habitatbeschikbaarheid en kennisleemten.

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Literatuuronderzoek onthulde voor quaggamossel hiaten in kennis die, afhankelijk van de projectdoelstellingen, relevant kunnen voor het voorspellen van rivierherstel. Voorbeelden daarvan zijn (1) de huidige verspreiding en verspreidingsmechanismen van de quaggamossel in West-Europa, en (2) de trofische overdracht van metalen via deze soort naar predatoren. Onze analyse heeft aangetoond dat de quaggamossel zich snel verspreidt, een breed scala aan dispersievectoren gebruikt, en al wijd is verspreid in Nederland en vele andere delen van West-Europa (Hoofdstuk 5). Herstelmaatregelen die de habitatgeschiktheid verbeteren zullen waarschijnlijk ook het verspreidingsrisico van de quaggamossel verhogen. De driehoeksmossel (Dreissena polymorpha) is nauw verwant aan de quaggamossel en komt wijdverspreid in Nederland voor. Kennisgebrek over de gevolgen van een dominantieverschuiving van de driehoeksmossel naar de quaggamossel wat betreft de trofische overdracht van metalen kan onzekerheid veroorzaken als de doelstellingen van rivierherstel zich richten op het terugkeren van (top)predatoren die gevoelig zijn voor metaaltoxiciteit (Hoofdstuk 6). Uit de analyses blijkt dat in alle wateren een dominantieverschuiving naar de quaggamossel de trofische overdracht van nikkel en koper mogelijk vermindert en de trofische overdracht van zink in rivierwater mogelijk verhoogt. Rivierherstelmaatregelen die de vestiging van de quaggamossel bevorderen kunnen ten goede komen aan doelsoorten die gevoelig zijn voor nikkel- en kopertoxiciteit. Aan de andere kant kunnen zulke maatregelen nadelige effecten hebben op doelsoorten die gevoelig zijn voor zink.

De uitkomsten van risicoanalyses kunnen een waardevolle bijdrage leveren aan de beoordeling van herstelmaatregelen. Het risico van vestiging van invasieve exoten als gevolg van herstelmaatregelen kan bijvoorbeeld worden meegenomen bij het opstellen van een 'operationeel leitbild' of lijst van haalbare hersteldoelstellingen. Aanbevolen wordt dat natuurbeheerders ook alternatieve herstelmaatregelen overwegen om het risico van de vestiging van invasieve exoten te verminderen. Zo kunnen maatregelen gericht op een lage stroomsnelheid en toename in hard substraat het risico van een invasie van de quaggamossel verhogen als bronpopulaties, introductieroutes en transportvectoren aanwezig zijn. Bij het voorspellen van het succes van het project moet rekening worden gehouden met een mogelijke vertraging en fluctuerende abundantie die kenmerkend zijn voor de introductie en vestiging van exoten.

Nader onderzoek moet zich richten op de bijdragen die exoten leveren aan het functioneren van nieuwe of veranderde ecosystemen. Functionele gelijkwaardigheid kan misschien een uitgangspunt vormen voor het meenemen van bepaalde exoten in een 'operationeel leitbild'. Dit betekent dat exoten de functies van referentiesoorten overnemen als de betreffende referentiesoorten niet terugkeren, bijvoorbeeld door afwezigheid van bronpopulaties. Het wordt aanbevolen dat horizonscans en risicobeoordelingen worden toegepast op een biogeografische in plaats van een politieke of administratieve schaal. Deze maatregel zou de consistentie van habitat- en klimaatmatches verbeteren en daardoor de zekerheid van de

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risicoclassificaties verhogen. Bovendien moeten toekomstige horizonscans voor potentieel invasieve exoten ook symbionten, parasieten en commensalen meenemen, omdat deze groepen tot nu toe grotendeels verwaarloosd zijn.

Toekomstige risicobeoordelingen van exoten kunnen ook gebiedspecifieke in plaats van soortspecifieke benaderingen omvatten. Gebiedspecifieke risicobeoordelingen richten zich op de kwetsbaarheid van ecosystemen voor biologische invasies en zijn gebaseerd op bestaande introductieroutes van exoten. De resultaten daarvan kunnen bijdragen aan het ontwerp van gebiedspecifieke preventie, vroegtijdige signalering en snelle responsmaatregelen gericht op meerdere invasieve soorten. Daarnaast moet rekening worden gehouden met mogelijke bronnen van inconsistentie tussen risicoclassificaties met beschikbare risicobeoordelingsprotocollen. De voorgestelde aanpak is essentieel voor de legitimiteit van elke risicobeoordeling van exoten en zal de acceptatie daarvan door beslissers, natuurbeheerders en andere belanghebbenden vergroten.

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Acknowledgements

Acknowledgements

More than 20 years after graduating with a Bachelor degree in Environmental Science from the Manchester Metropolitan University in the UK, and nearly 10 years after completing my MSc in Transnational Ecosystem Based Water Management at Radboud University and the University of Duisburg Essen, I am defending my doctoral thesis in the Netherlands. It’s taken a while, and I’ve been diverted by various other challenges along the way, not least by a short career in the UK’s National Health Service working as a nurse in coronary care. But here I am, and I wouldn’t be standing here if it weren’t for the help of some important people.

First and foremost, I’d like to thank my promoters and co-promoters who have consistently supported me throughout the production of this thesis. Their indestructible positivity, optimism and enthusiasm led to solutions to what I often thought were insurmountable problems.

My promoter Rob Leuven provided consistent support and guidance which began during a placement in the Department of Environmental Science in 2009, and was sustained right up to the completion of the thesis. I have huge respect for his analytical and critical skills, intellect, passion for science and approach to team-working. Unfortunately, I didn’t receive funding that could have led to a continuous four year contract for the research done here. Instead a number of individual contracts were funded by a series of individual projects, each won by Rob. Each time a new project was won, Rob committed to creating a new contract or extending an existing one so that I could continue on the path that has led me to where I am today. Rob is also very modest. During a speech that celebrated him receiving a professorship, Rob gave credit to virtually every other member of the permanent staff in the Department of Environmental Science apart from himself for his achievement! When I asked him why he hadn’t taken just a little bit of credit for his professorship he replied ‘zo ben ik niet’ or ‘that’s just not me’.

My co-promotor Gerard van der Velde and promoter Jan Hendriks have provided continued advice and support during this process. Gerard’s critical thinking and consideration of different perspectives always added polish to the finished product, whether it was an article or a report. His superior taxonomic knowledge was a vital addition to many of the projects we undertook. Officially Gerard is retired, but whoever said that retirement should lead to the end of a remarkable scientific career? Since retirement Gerard has decided not to ‘sit behind the geraniums’, an English translation of a Dutch phrase that means taking it easy, but has continued his scientific efforts. Rather than sitting behind the geraniums, I imagine that Gerard would rather study them and publish the results in a renowned scientific journal! I am also grateful to Jan Hendriks who set me on this scientific pathway by supervising the publication of my first article, and provided valuable support during a work placement when things got difficult. His application of Socratic questioning always

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stimulates creative thinking, providing a fertile basis for innovative approaches and learning in his students and colleagues.

Then there are my other colleagues from the Institute for Water and Wetland Research (IWWR) who I’ve worked with on various project and articles on alien species. Aafke Schipper provided important guidance particularly during the writing of chapter 6. Her detailed experimental knowledge with regards to the analysis of metal transfer in food webs was vital to the success of this publication. Lisette de Hoop, Frank Collas and Kevin (Remon) Koopman, always provided useful comments and additional information. The discussions, swapping of ideas and constructive criticism only improved the results of my work, and I wish you all lots of luck in your future professional careers. Chapter 5 and 6 involved the identification of aquatic macroinvertebrate species and analysis of metal concentrations in their tissues. At the time, I didn’t have much experience in this type of laboratory work, so the assistance received from the lab technicians Germa Verheggen, Marij Orbons and Jelle Eygensteyn was gratefully received. Many other persons and organisations made valuable contributions to the projects on which much of this thesis is based, including the members of the Netherlands Centre of Expertise on Exotic Species (NEC-E), the Netherlands Food and Consumer Product Safety Authority (NVWA) and the Zuiderzeeland Regional Water Authority. If not mentioned in these acknowledgements, contributors are thanked at the end of each thesis chapter.

A big ‘thank you’ goes to all the permanent and temporary staff of the Department of Environmental Science, both past and present, for the lively discussions and positive atmosphere. Specifically thanks to Laura, Daan, Anastasia, Alessandra, Ligia, Rik, Thomas, Ana and Zoran for the fun times, PhD dinners, and adventures. You provided me with a welcome distraction from the challenges of living away from home and integrating into a foreign culture (and showed me that board games really are not my strong point!). I’m also really pleased that Rik and Ana are both able to take on the role of para-nymph during my defence, thanks guys.

Friendships are important and since moving to the Netherlands I have made many new ones, and cemented existing ones. These friendships have helped enormously when I needed to let of some steam and share the tough challenges that I faced not only from academic work, but also from settling in a new and unfamiliar country. So, thank you to Robert and Maikel, Frans, Heiko and all the members of Cantus Obliquus, Simone and Edwin, Petra and Paul for all the good times.

To my family in the UK, this thesis is for you. We’ve been through some tough times recently but, despite these, we’ve stuck together. If there’s anything I’ve learnt over the past years it’s that nothing stays the same! Thanks also go to my family in law in the Netherlands for all the support they’ve given since we arrived in the Netherlands in 2006:

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‘Bedankt voor jullie steun’. Last but not least, thank you to Sander, my partner of over 17 years, who has stood by me during all the ups and downs that life has thrown at us. Life keeps getting better, long may it continue!

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Curriculum vitae

Curriculum vitae

Jonathan Matthews was born on the 23rd of June 1973 in Epsom, the United Kingdom. After completing his secondary school education at Glyn School in Epsom, he moved to Manchester where he studied Environmental Science at Manchester Metropolitan University graduating in 1994. After graduation, he worked as a nursing assistant and eventually retrained, becoming a qualified nurse in 2000. In 2006, after working for five years in the UK’s National Health Service, he moved to the Netherlands with his Dutch partner and took the chance to return to the environmental sector by obtaining an Environmental Science Master degree (Bene Meritum) from Radboud University, Nijmegen in 2010. During his Master internship he worked at Arcadis Environmental Consultancy in Apeldoorn where he undertook research into the monitoring and assessment of river rehabilitation projects leading to the publication of the article that comprises chapter one of this thesis. In 2010, after a short period working as a junior researcher at the Alterra Research Institute in Wageningen, examining the evidence base supporting methods applied in the rehabilitation of rivers, he returned to the Radboud University where his research has focussed on the assessment of factors determining the introduction, establishment, spread, impacts and management of invasive, alien species in the Netherlands. During his time at the Radboud University, he has authored a number of reports on alien species for the Netherlands Food and Consumer Product Authority (NVWA) and Water-board Zuiderzeeland including a horizon-scanning project assessing the potential ecological risks of alien biomass crops to the natural environment in the Netherlands, a rapid assessment protocol for newly recorded alien species in the Netherlands, and an assessment of the influence of climate change on environmental stressors and species richness in the Province of Flevoland. Additionally, he has supervised students and presented during the courses Biodiversity and Ecological Assessment, Orientation in Biology and Environmental Sciences, and Humans and the Environment, provided by the Department of Environmental Science. On completion of his PhD, Jon began work as an adviser for the engineering consultancy SWECO Nederland.

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List of publications

List of publications

Peer-reviewed journal articles

Collas FPL, Breedveld SKD, Matthews J, Van der Velde G, Leuven RSEW. 2017. Invasion biology and risk assessment of the recently introduced Chinese mystery snail, Bellamya (Cipangopaludina) chinensis (Gray, 1834) in the Rhine and Meuse River basins in Western Europe. Aquatic Invasions 12: (in press). Leuven RSEW, Hendriks AJ, Huijbregts M, Lenders HJR, Matthews J, Van der Velde G. 2011. Differences in sensitivity of native and exotic fish species to changes in river temperature. Current Zoology 57:852-862. Leuven RSEW, Collas FPL, Koopman KR, Matthews J, Van der Velde G. 2014. Mass mortality of invasive zebra and quagga mussels by desiccation during severe winter conditions. Aquatic Invasions 9:243-252. Matthews J, Beringen R, Creemers R, Hollander H, Van Kessel N, Van Kleef H, Van de Koppel S, Lemaire AJJ, Odé B, Verbrugge LNH, Hendriks AJ, Schipper AM, Van der Velde G, Leuven RSEW. 2017. A new approach to horizon-scanning: identifying potentially invasive alien species and their introduction pathways. Management of Biological Invasions 8:37-52. Matthews J, Van der Velde G, Collas FPL, De Hoop L, Koopman KR, Hendriks AJ, Leuven RSEW. 2017. Inconsistencies in the risk classification of alien species and implications for risk assessment in the European Union. Ecosphere 8:e01832. 10.1002/ecs2.1832. Matthews J, Reeze B, Feld CK, Hendriks AJ. 2010. Lessons from practice: assessing early progress and success in river rehabilitation. Hydrobiologia 655:1-14. Matthews J, Schipper AM, Hendriks AJ, Le TY, Bij de Vaate A, Van der Velde G, Leuven RSEW. 2015. A dominance shift from the zebra mussel to the invasive quagga mussel may alter the trophic transfer of metals. Environmental Pollution 203:183- 190. Matthews J, Van der Velde G, Bij de Vaate A, Collas FPL, Koopman KR, Leuven RSEW. 2014. Rapid range expansion of the invasive quagga mussel in relation to zebra mussel presence in the Netherlands and Western Europe. Biological invasions 16:23-42.

Professional reports

Collas FPL, Beringen R, Koopman KR, Matthews J, Odé B, Pot R, Sparrius LB, Van Valkenburg JLCH, Verbrugge LNH, Leuven RSEW. 2012. Knowledge document for risk analysis of the non-native Tapegrass (Vallisneria spiralis) in the Netherlands. Reports Environmental Science 416. Radboud University Nijmegen, Institute for Water and Wetland Research, Department of Environmental Science. 38 pp. De Hoop L, Bruijs MCM, Collas FPL, Dionisio Pires LM, Dorenbosch M, Gittenberger A, Matthews J, Van Kleef HH, Van der Velde G, Vonk JA, Leuven RSEW. 2015. Risicobeoordeling en uitzetcriteria voor de uitheemse quaggamossel (Dreissena rostriformis bugensis) in Nederland. Verslagen Milieukunde 507. Radboud Universiteit Nijmegen, Instituut voor Water en Wetland Research, Afdeling Milieukunde. 84 pp. Koopman KR, Beringen R, Collas FPL, Matthews J, Odé B, Pot R, Sparrius LB, Van Valkenburg JLCH, Verbrugge LNH, Leuven RSEW. 2012. Knowledge document

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List of publications

for risk analysis of the non-native monkeyflower (Mimulus guttatus) in the Netherlands. Reports Environmental Science 415. Radboud University Nijmegen, Institute for Water and Wetland Research, Department of Environmental Science. 43 pp. Koopman KR, Matthews J, Beringen R, Odé B, Pot R, Van der Velde G, Van Valkenburg JLCH, Leuven RSEW. 2014. Risicoanalyse van de uitheemse Egeria (Egeria densa) in Nederland. Verslagen Milieukunde 469. Radboud Universiteit Nijmegen, Instituut voor Water en Wetland Research, Afdeling Milieukunde. 54 pp. Matthews J, Beringen R, Boer E, Duistermaat H, Odé B, Van Valkenburg JLCH, Van der Velde G, Leuven RSEW. 2015. Risks and management of non-native Impatiens species in the Netherlands. Reports Environmental Science 491. Radboud University Nijmegen, Institute for Water and Wetland Research, Department of Environmental Science. 178 pp. Matthews J, Beringen R, Collas FPL, Koopman KR, Odé B, Pot R, Sparrius LB, Van Valkenburg JLCH, Verbrugge LNH, Leuven RSEW. 2012. Risk analysis of the non-native native Tapegrass (Vallisneria spiralis) in the Netherlands. Reports Environmental Science 420. Radboud University Nijmegen, Institute for Water and Wetland Research, Department of Environmental Science. 32 pp. Matthews J, Beringen R, Collas FPL, Koopman KR, Odé B, Pot R, Sparrius LB, Van Valkenburg JLCH, Verbrugge LNH, Leuven RSEW. 2012. Risk analysis of the non-native native monkeyflower (Mimulus guttatus) in the Netherlands. Reports Environmental Science 419. Radboud University Nijmegen, Institute for Water and Wetland Research, Department of Environmental Science. 29 pp. Matthews J, Beringen R, Collas FPL, Koopman KR, Odé B, Pot R, Sparrius LB, Van Valkenburg JLCH, Verbrugge LNH, Leuven RSEW. 2012. Risk analysis of the non-native curly waterweed (Lagarosiphon major) in the Netherlands. Reports Environmental Science 418. Radboud University Nijmegen, Institute for Water and Wetland Research, Department of Environmental Science. 32 pp. Matthews J, Beringen R, Collas FPL, Koopman KR, Odé B, Pot R, Sparrius LB, Van Valkenburg JLCH, Verbrugge LNH, Leuven RSEW. 2012. Knowledge document for risk analysis of the non-native curly waterweed (Lagarosiphon major) in the Netherlands. Reports Environmental Science 414. Radboud University Nijmegen, Institute for Water and Wetland Research, Department of Environmental Science. 43 pp. Matthews J, Beringen R, Creemers R, Hollander H, Van Kessel N, Van Kleef H, Van de Koppel S, Lemaire AJJ, Odé B, Van der Velde G, Verbrugge LNH, Leuven RSEW. 2014. Horizonscanning for new invasive non-native species in the Netherlands. Reports Environmental Science 461. Radboud University Nijmegen, Institute for Water and Wetland Research, Department of Environmental Science. 114 pp. Matthews J, Beringen R, Huijbregts MAJ, Van der Mheen HJ, Odé B, Trindade L, Van Valkenburg JLCH, Van der Velde G, Leuven RSEW. 2015. Horizon-scanning and environmental risk analyses of non-native biomass crops in the Netherlands. Reports Environmental Science 506. Radboud University Nijmegen, Institute for Water and Wetland Research, Department of Environmental Science. 253 pp. Matthews J, Beringen R, Lamers LPM, Odé B, Pot R, Van der Velde G, Van Valkenburg JLCH, Verbrugge LNH, Leuven RSEW. 2013. Knowledge document for risk

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analysis of the non-native Fanwort (Cabomba caroliniana) in the Netherlands. Reports Environmental Science 443. Radboud University Nijmegen, Institute for Water and Wetland Research, Department of Environmental Science. 60 pp. Matthews J, Beringen R, Lamers LPM, Odé B, Pot R, Van der Velde G, Van Valkenburg JLCH, Verbrugge LNH, Leuven RSEW. 2013. Risk analysis of the non-native Fanwort (Cabomba caroliniana) in the Netherlands. Reports Environmental Science 442. Radboud University Nijmegen, Institute for Water and Wetland Research, Department of Environmental Science. 46 pp. Matthews J, Beringen R, Leuven RSEW, Van der Velde G, Van Valkenburg JLCH, Odé B. 2015. Knowledge document for risk analysis of the alien poison ivy (Toxicodendron radicans) in the Netherlands. Reports Environmental Science nr. 477. Radboud University Nijmegen, Institute for Water and Wetland Research, Department of Environmental Science. 57 pp. Matthews J, Collas FPL, De Hoop L, Van der Velde G, Leuven RSEW. 2017. Risk assessment of the alien Chinese mystery snail (Bellamya chinensis). Reports Environmental Science 557. Radboud University Nijmegen, Institute for Water and Wetland Research, Department of Environmental Science. 77 pp. Matthews J, Collas FPL, De Hoop L, Van der Velde G, Leuven RSEW. 2017. Management approaches for the alien Chinese mystery snail (Bellamya chinensis). Reports Environmental Science 558. Radboud University Nijmegen, Institute for Water and Wetland Research, Department of Environmental Science. 23 pp. Matthews J, Koopman KR, Beringen R, Odé B, Pot R, Van der Velde G, Van Valkenburg JLCH, Leuven RSEW. 2014. Knowledge document for risk analysis of the non- native Brazilian waterweed (Egeria densa) in the Netherlands. Reports Environmental Science 468. Radboud University Nijmegen, Institute for Water and Wetland Research, Department of Environmental Science. 58 pp. Matthews J, Oudendijk MJJM, Huijbregts MAJ, Schipper AM, Leuven RSEW. 2013. Gevoeligheid van aquatische soorten voor veranderingen in klimaatgerelateerde waterkwaliteitsparameters. Verslagen Milieukunde 428. Radboud Universiteit Nijmegen, Instituut voor Water en Wetland Research, Afdeling Milieukunde. 98 pp. Matthews J, Oudendijk MJJM, Leuven RSEW. 2014. Gevoeligheid van aquatische soorten voor veranderingen in klimaatgerelateerde waterkwaliteitsparameters 1: Ruimtelijke vertaling voor KRW doelsoorten. Verslagen Milieukunde 449. Radboud Universiteit Nijmegen, Instituut voor Water en Wetland Research, Afdeling Milieukunde. 98 pp. Matthews J, Oudendijk MJJM, Leuven RSEW. 2014. Gevoeligheid van aquatische soorten voor veranderingen in klimaatgerelateerde waterkwaliteitsparameters 2: KRW dominant negatieve macrofauna taxa. Verslagen Milieukunde 450. Radboud Universiteit Nijmegen, Instituut voor Water en Wetland Research, Afdeling Milieukunde. 51 pp. Matthews J, Oudendijk MJJM, Leuven RSEW. 2014. Gevoeligheid van aquatische soorten voor veranderingen in klimaatgerelateerde waterkwaliteitsparameters 3: Relaties tussen ecologische kwaliteitsratio’s en potentieel afwezige fracties van KRW doelsoorten. Verslagen Milieukunde 451. Radboud Universiteit Nijmegen, Instituut voor Water en Wetland Research, Afdeling Milieukunde. 37 pp. Matthews J, Van der Velde G, Bij de Vaate A, Leuven RSEW. 2012. Key factors for spread, impact and management of Quagga mussels in the Netherlands. Reports

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Environmental Science 404. Radboud University Nijmegen, Institute for Water and Wetland Research, Department of Environmental Science & Department of Animal Ecology and Ecophysiology & Waterfauna Hydrobiologisch Adviesbureau, Lelystad, Nijmegen, The Netherlands. 123 pp. Matthews J, Collas FPL, de Hoop L, Van der Velde G, Leuven RSEW. 2017. Rapid assessment protocol for new alien species in the Netherlands. Reports Environmental Science 566. Radboud University, Institute for Water and Wetland Research, Department of Environmental Science, (in prep.). Schiphouwer ME, Van Kessel N, Matthews J, Leuven RSEW, Van de Koppel S, Kranenbarg J, Haenen OLM, Lenders HJR, Nagelkerke LAJ, Van der Velde G, Crombaghs BHJM, Zollinger R. 2014. Risk analysis of exotic fish species included in the Dutch Fisheries Act and their hybrids. Nederlands Expertise Centrum Exoten (NEC-E), Nijmegen, The Netherlands, 207 pp Soes DM, Leuven RSEW, Matthews J, Broeckx P-B, Haenen OLM, Engelsma MY. 2011. A risk analysis of bigheaded carp (Hypophthalmichthys sp.) in the Netherlands. Bureau Waardenburg, Culemborg, The Netherlands. 118 pp.

Professional articles

De Hoop L, Van der Loop J, Matthews J, Van der Velde G, Leuven RSEW. 2017. Europese regelgeving voor beheer van invasieve exoten. De Levende Natuur 118 (4):112- 116. Leuven RSEW, Beringen R, Boer E, Duistermaat L, Van Kemenade L, Matthews J, Odé B, Simons B, Van Valkenburg J, Van der Velde G. 2017. Van horizonscanning naar kosteneffectief beheer van invasieve springzaden. De Levende Natuur 118 (4):139-142. Leuven RSEW, Matthews J, Van der Velde G. 2016. Exotische biomassagewassen: oplossing of probleem. Tijdschrift Milieu 4:34-35. Matthews JM, Van der Velde G, Leuven RSEW. 2016. Voorkomen beter dan bestrijden. Tijdschrift Milieu 4:30-31.

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