Recommended citation:
Oficialdegui, F. J. (2019). The invasion of the red swamp crayfish (Procambarus clarkii): introductions, impacts and management. PhD Thesis. Universidad de Sevilla, Seville,
Spain.
Cover images are original creations of Paula Martin Díaz (@paulamartinart), who remains her intellectual owner. Any form of reproduction, distribution, public communication or transformation of the same without prior authorization of the author is forbidden. THE INVASION OF THE RED SWAMP CRAYFISH
(PROCAMBARUS CLARKII):
Introductions, Impacts and Management
Francisco Javier Oficialdegui
PhD Thesis
Seville, 2019 Estación Biológica de Doñana, EBD - CSIC
Departamento de Ecología de Humedales
Universidad de Sevilla
Facultad de Biología
Departamento de Zoología
Doctorado en Biología Integrada
THE INVASION OF THE RED SWAMP CRAYFISH (PROCAMBARUS CLARKII):
Introductions, Impacts and Management
Memoria presentada por el Licenciado en Biología, Francisco Javier Oficialdegui, para
optar al título de Doctor por la Universidad de Sevilla
Fdo. Francisco Javier Oficialdegui Dra. Marta I. Sánchez, Profesora contratada doctora, INMAR (Instituto de
Investigación Marina), Universidad de Cádiz; Dr. Miguel Clavero, Científico Titular de la Estación Biológica de Doñana (EBD-CSIC); y Dra. Luz Boyero, Profesora de
Investigación Ikerbasque de la Universidad del País Vasco (UPV/EHU).
CERTIFICAN:
Que los trabajos de investigación desarrollados en la Memoria de Tesis Doctoral “The invasion of the red swamp crayfish (Procambarus clarkii): Introductions, Impacts and
Management”, son aptos para ser presentados por el Ldo. en Biología Francisco Javier
Oficialdegui ante el Tribunal que en su día se designe, para aspirar al grado de Doctor por la Universidad de Sevilla.
Y para que así conste, y en cumplimento de las disposiciones legales vigentes, firman el presente documento en Sevilla, a 10 de Abril de 2019.
Directores:
Fdo. Marta I. Sánchez Fdo. Miguel Clavero Fdo. Luz Boyero
Tutor: José Guerra
Universidad de Sevilla
This thesis was supported by projects P12-RNM-936 from the Junta de Andalucía and projects PIC2015FR4/PICS07360.
Crayfish samplings were carried out with all the necessary permits issued by the regional Department of the Environment (Consejería de Medio Ambiente, Junta de
Andalucía).
A ti,
a mi familia,
a Martina
“In the end we will conserve only what we love, we will love only what we understand, and
we will understand only what we are taught” (Baba Dioum)
INDEX
Abstract / Resumen 1
General introduction 9
Objectives 19
Chapter I 23
One century away from home: how the red swamp crayfish took over the
world
Chapter II 89
Unravelling the global invasion routes of a worldwide invader, the red
swamp crayfish (Procambarus clarkii)
Chapter III 133
Brought more than twice: the complex introduction history of the red
swamp crayfish into Europe
Chapter IV 145
The invasive red swamp crayfish (Procambarus clarkii) increases infection
of the amphibian chytrid fungus (Batrachochytrium dendrobatidis)
Chapter V 165
Rigid laws can set back invasive species management
General discussion 175
Conclusions 185
General references 187
Acknowledgements / Agradecimientos 195
ABSTRACT
Biological invasions are a significant component of current global environmental change, altering biodiversity, ecosystems and human well-being. Globalization has led to an exponential increase of large-distance movements of species through transport improvements and the increase in global connections. As consequence, many invasive species are currently widely distributed over the world. Understanding how invasive species are transported outside of their native ranges, which routes of introduction they follow and what impacts are they causing in the invaded area are some of the key questions on invasion science.
In this thesis, I use a multidisciplinary approach to shed light on the invasion process of one of the most invasive species worldwide, the red swamp crayfish
(Procambarus clarkii). It is native to southern U.S.A. and north-eastern Mexico and has been successfully introduced into all continents except Australia and Antarctica. Firstly,
I review the invasion history of the red swamp crayfish to know why, where and how it has translocated around the world. Secondly, I focus on the study of invasion routes at global scale using a population genetic approach. Thirdly, I centre my attention on indirect impacts in native fauna, focusing on amphibian communities in south-western
Spain and, in particular, the Doñana Natural Space, a highly protected area in Europe.
Finally, I analyse the conservation conflicts associated to the use of invasive species for commercial purposes, particularly, the crayfish industry in Isla Mayor (the third largest producer of the red swamp crayfish in the world).
By conducting an exhaustive literature search and using global databases of biodiversity (USGS, GBIF and iNaturalist), I made an update of the global invasion history of the red swamp crayfish in order to describe when it began to be introduced out of its native range, what has motivated to translocate this species worldwide and which invasion routes have been followed for this commercial species (Chapter 1).
Additionally, molecular tools were applied for reconstructing and genetically confirming the main global invasion routes to provide a comprehensive overview of its global invasion process (Chapters 2 and 3). Although the first introductions in the 1920s
1 Abstract occurred in western US and subsequently, Japan and China; there was not a boom of exportations from Louisiana until the 1960s and 1970s, related to commercial exploitation for human consumption but also associated to pet trade, mainly due to the sales of live specimens by commercial houses. Driven mostly by human-mediated introductions, commercial exploitation and pet trade probably led to a complex pattern of multiple introductions worldwide, in which, multiple scenarios of introduction took place. The resulting scenario depends on whether a potential bottleneck or founder effect exists or not. In turn, it is determined by the level of genetic admixture in the native range as consequence of local translocations for commercial purposes, where specimens come from (e.g., previously invaded areas can act as new source of introductions adding genetic diversity to recipient invasive population), and propagule pressure (i.e. amount of individuals and number of introduction events). Often, high haplotype diversity in populations of the red swamp crayfish was found in both native and some invaded areas. However, I also detected low levels of genetic diversity in some non-native areas
(e.g., in Asia). In addition, I identified, on the basis of genetic analysis, two unconfirmed invasion routes into Europe from Africa and Asia, independently of those from
Louisiana (U.S.A.) (Chapter 3). Thus, it demonstrates that the invasion process of the red swamp crayfish in Europe, and plausibly elsewhere, is more complex than generally assumed. Overall, these results allowed the identification of the likely geographic origin and main routes of invasion, helping us to understand how the invasion has happened over a long time scale.
Once the red swamp crayfish populations are established in an area, they cause severe impacts on invaded habitats and ecosystem functioning. Among these, freshwater crayfish are an important cause of amphibian declines worldwide through direct predation, but little is known about its indirect effects in relation with amphibian chytrid fungus. Chytridiomycosis is a fungus disease caused by Batrachochytrium dendrobatidis (Bd), which is considered one of the most important causes for the decline of amphibian populations worldwide. I studied the potential role of the red swamp crayfish as biological reservoir of Bd and the environmental persistence of this zoonotic
2 Abstract pathogen (Chapter 4). To do that, I conducted a survey across western Andalusian waterbodies in SW Spain, discovering the presence of zoosporangia of Bd in gastrointestinal tissue of crayfish. Subsequently, I carried out another survey in Doñana
Natural Space where the prevalence of Bd-infection in amphibians was significantly related to the presence of the red swamp crayfish in the pond, indicating that this crayfish species can be a suitable predictor of Bd-infection in co-occurring amphibians.
The results of this study suggest that the red swamp crayfish may act as potential reservoir of Bd pathogen, which have important implications for amphibian conservation worldwide.
Finally, I addressed the unavoidable social conflicts associated to the management of exploited invasive species and how rigid laws that force management options in spite of the lack of expectations of positive outcomes can set back invasive species management (Chapter 5). In particular, I focused on the conflict between conservationists and the crayfish industry in Isla Mayor, Seville, by showing the ups and downs in application of regulations for the use of the red swamp crayfish in Spain. I highlight that all stakeholders must feel involved in the planning of regulations affecting their interests and that these regulations support the common weal, in order to avoid misconception, opposition and setbacks.
The study of globally distributed invasive species, as the red swamp crayfish, allows identifying the main human-motivations to translocate invasive species and to understand the history of the invasion process, which may help to limit its spreading and avoid the entry of new emerging alien species. This thesis therefore provides a multidisciplinary framework for the prevention of emerging invaders as well as suggests further ecological and evolutionary questions for forthcoming studies on invasion biology. The results exposed here are utterly needed for biodiversity conservation in non-native areas.
3 RESUMEN
Las invasiones biológicas son un importante componente del actual cambio global, alterando la biodiversidad, los ecosistemas y el bienestar humano. La globalización ha conllevado a un aumento exponencial de los movimientos de especies a grandes distancias debido a las mejoras en el transporte y el aumento de las conexiones globales.
Como consecuencia, muchas especies invasoras actualmente están distribuidas en todo el mundo. Algunas de las preguntas clave sobre las invasiones biológicas son comprender cómo se transportan las especies invasoras fuera de sus rangos nativos, qué rutas de introducción siguen y qué impacto están causando en el área invadida.
En esta tesis, utilizo un enfoque multidisciplinar para arrojar luz sobre el proceso de invasión de una de las especies más invasoras del mundo, el cangrejo rojo
(Procambarus clarkii). Nativo del sur de los Estados Unidos y del noreste de México, se ha introducido con éxito en todos los continentes, excepto en Australia y la Antártida.
En primer lugar, reviso la historia de la invasión del cangrejo rojo para saber por qué, dónde y cómo se ha desplazado esta especie por todo el mundo. En segundo lugar, abordo las rutas de invasión a escala global utilizando un enfoque genético poblacional.
En tercer lugar, centro mi atención en los impactos indirectos en la fauna nativa, concretamente en las comunidades de anfibios del suroeste de España y, en particular, en el Espacio Natural de Doñana, un área altamente protegida en Europa. Finalmente, analizo los conflictos de conservación asociados con el uso de especies invasoras con fines comerciales, en particular, en la industria del cangrejo rojo en Isla Mayor (el tercer mayor productor de cangrejo rojo del mundo).
Mediante una búsqueda bibliográfica exhaustiva y utilizando bases de datos globales de biodiversidad (USGS, GBIF e iNaturalist), realicé una actualización de la historia de invasión global del cangrejo rojo con el fin de describir cuándo comenzó a introducirse fuera de su rango nativo, qué ha motivado su translocación por todo el mundo y qué rutas de invasión ha seguido (Capítulo 1). Además, se usaron técnicas de análisis molecular para reconstruir y confirmar genéticamente las principales rutas de invasión global, de esta manera, proporcionar una visión general de su proceso de
4 Resumen invasión global (Capítulos 2 y 3). Aunque las primeras introducciones ocurrieron hacia la década de los 20 en el oeste de los Estados Unidos y posteriormente, en Japón y China; no hubo un auge de las exportaciones desde Luisiana hasta las décadas de los 60 y 70, no solo relacionadas probablemente con la explotación comercial para el consumo humano, sino también asociadas al comercio de mascotas, principalmente debido a las ventas de especímenes vivos por parte de las casas comerciales. Impulsado principalmente por introducciones mediadas por humanos, la explotación comercial probablemente condujeron a un patrón complejo de múltiples introducciones por todo el mundo, en el que se llevaron a cabo múltiples introducciones. El escenario resultante depende de si existe o no un posible cuello de botella o un efecto fundador. A su vez, está determinado por el nivel de mezcla genética en el rango nativo como consecuencia de las translocaciones locales con fines comerciales, de donde provienen los especímenes
(por ejemplo, las áreas previamente invadidas pueden actuar como una nueva fuente de introducciones que agregan diversidad genética a la población invasora receptora) y presión de propágulo (es decir, cantidad de individuos y cantidad de eventos de introducción). A menudo, se encontró una gran diversidad haplotípica en las poblaciones del cangrejo rojo en el rango nativo y en algunas áreas invadidas. Sin embargo, también se encontraron bajos niveles de diversidad haplotípica en algunas
áreas no nativas (por ejemplo, en Asia). Además, en base al análisis genético, identifiqué dos rutas de invasión no confirmadas hacia Europa desde África y Asia, independientemente de las descritas desde Luisiana (EEUU) (Capítulo 3). Por lo tanto, estos resultados demuestran que el proceso de invasión del cangrejo rojo en Europa, y plausiblemente en otros lugares, es más complejo de lo que generalmente se supone. En general, estos resultados permitieron identificar el origen geográfico más probable y las principales rutas de invasión, lo que ayudó a comprender cómo ha ocurrido la invasión a lo largo del tiempo.
Una vez que las poblaciones de cangrejo rojo se han establecido en un área, causan graves impactos en los hábitats invadidos y en el funcionamiento del ecosistema. Entre
éstos, el cangrejo rojo es una importante causa del declive poblacional de anfibios en
5 Resumen todo el mundo a través de la depredación directa de larvas e individuos adultos; sin embargo, se sabe poco acerca de sus efectos indirectos en relación con el hongo quitridio.
La quitridiomicosis es una enfermedad causada por el hongo Batrachochytrium dendrobatidis (Bd), y se considera una de las causas más importantes del declive de las poblaciones de anfibios en todo el mundo. En este capítulo, estudié el papel potencial del cangrejo rojo como reservorio biológico de Bd y en la persistencia de este patógeno zoonótico en el ambiente donde vive (Capítulo 4). Para ello, realicé un estudio a través de arroyos y charcas en el oeste de Andalucía, suroeste de España, descubriendo la presencia de zoosporangios de Bd en tejido gastrointestinal del cangrejo rojo.
Posteriormente, realicé otro muestreo en el Espacio Natural de Doñana, donde la prevalencia de infección por Bd en anfibios se relacionó significativamente con la presencia del cangrejo rojo en la charca, lo que indica que esta especie de cangrejo de río puede ser un predictor adecuado de la infección por Bd en anfibios con los que conviven.
Los resultados de este estudio sugieren que el cangrejo rojo podría actuar como reservorio potencial del patógeno Bd, teniendo notables implicaciones para la conservación de anfibios en todo el mundo.
Finalmente, en el último capítulo abordé los inevitables conflictos sociales asociados con el manejo de las especies invasoras comercialmente explotadas por el ser humano, y como la rigidez de las leyes que potencian el manejo a pesar de la falta de expectativas positivas, pueden entorpecer y complicar el manejo de las especies invasoras (Capítulo 5). En particular, me centré en el conflicto que surgió entre conservacionistas y la industria del cangrejo rojo en Isla Mayor, Sevilla, mostrando los altibajos en la aplicación de las regulaciones para el uso del cangrejo rojo en España.
Aquí, señalo que todas las partes interesadas deben sentirse involucradas en la planificación de las regulaciones que afectan sus intereses y que estas regulaciones deben apoyar al bien común, a fin de evitar confusiones, la oposición y contratiempos.
El estudio de especies invasoras distribuidas globalmente, como el cangrejo rojo, sirve como ejemplo para identificar las principales motivaciones humanas a la translocación de las especies invasoras, entender la historia del proceso de invasión,
6 Resumen ayudar a limitar su propagación y evitar nuevas introducciones. Por lo tanto, esta tesis proporciona un marco multidisciplinar para la prevención de invasores emergentes y también sugiere preguntas ecológicas y evolutivas para los próximos estudios sobre la biología de la invasiones. Los resultados expuestos aquí son totalmente necesarios para la conservación de la biodiversidad en áreas no nativas.
7
GENERAL INTRODUCTION
Biological invasions are a leading driver of global environmental change (Sala et al.
2000; Simberloff et al. 2013), with devastating effects on biodiversity (Clavero & García-
Berthou 2005; Blackburn et al. 2019), ecosystems and human well-being (Vilà et al. 2010).
Alien species are those transported and introduced (accidentally or deliberately) by humans in an area outside of their native range (Gilroy et al. 2017), being able to establish, thrive, become abundant, and spread over non-native areas (named then as invasive species, Blackburn et al. 2011) and few of them cause severe impacts on the invaded ecosystems. Humans have transported species across biogeographical barriers and introduced them into new territories for millennia (Forcina et al. 2015). However, growing transport and global connections have led to an exponential increase of large- distance movements of alien species in recent decades (Hulme 2009), without showing a sign of saturation (Seebens et al. 2017) and causing the homogenization of biotas
(Capinha et al. 2015).
INVASION PROCESS
The invasion process is often described as a multi-step process with various stages separated by barriers, where introduced species can achieve or fail to overcome, usually being termed invasive if succeeding in making their way to the final stages (Blackburn et al. 2011; Lockwood et al. 2013) (Fig. 1). The first stage is the transport of individuals outside of their native range. Before overtaking the geographical barrier, there is an uptake of individuals in the native range where individuals have to be captured (1A in
Fig. 1). Often, this stage starts in the species’ native range, but not necessarily, another previously successfully invaded area can act as source of introductions (Fischer et al.
2017; van Boheemen et al. 2017). The more invaded areas there are, the more potential sources for future introductions are possible; thus favouring the rapid spreading by multiple events of introduction. Due to human-mediated dispersal, it is likely that transport stage is overcome several times before concluding the entire invasion process
(secondary introductions, Fig. 1). The second stage is the introduction of alien species
9 General introduction into a novel area outside of their native range. Due to the economic value of some alien species, individuals can be bred in captivity or cultured before being released into the wild (2A in Fig. 1). This situation can be given in many cases, for example, species for agriculture, ornamental horticulture, aquaculture or pet trade. However, many other alien species are introduced into the wild directly without overtaking this barrier (2B in
Fig. 1), what often happen intentionally with many game species or accidentally in the case of species that arrive through ballast water or transported goods (Mack 2003;
Lockwood et al. 2013). Regarding on how alien species are introduced, there are multiple pathways which are usually mechanism-, ecosystem- or taxon-dependent (e.g. stowaway for ballast water, escape for livestock, or releases for game species) (Hulme et al. 2008). Additionally, other factors such as artificial selection during the cultivation or captivity affect positively the invasiveness of species to overcome this barrier (Briski et al. 2018). The third stage is the establishment in a non-native area. This stage is probably the most crucial point within the entire invasion process because individuals have to survive (3A in Fig. 1) and after, be able to reproduce (3B in Fig. 1). In this sense, deliberately introduced species (e.g. aquaculture species or pets) may have a certain advantage in survival and reproduction because humans have taken care of their breeding and feeding (Pyšek et al. 2011) or have selected suitable individuals (i.e. high fitness) for the recipient environment (Briski et al. 2018). A key factor that alters the establishment success is the propagule pressure (Williamson & Fitter 1996) because the more individuals are introduced in a novel area, the higher their probability of establishment (Fig. 2) (Frankham 2005; Lockwood et al. 2009). Lastly, the fourth stage is the spread of naturalized or established species that are self-sustaining (3C in Fig. 1).
Here, individuals are able to thrive and expand over the non-native area facing to novel environmental conditions, to parasite acquisition and to compete for resources with native species, among others (4A in Fig. 1). While some naturalized species can expand naturally, the spread of many others consists of a set of continuous establishments of new individuals over the non-native area (i.e. secondary introduction events, Fig. 1).
When a naturalized species is able to live and be self-sustained far from where it has
10 General introduction been introduced, it called invasive species (5 in Fig. 1). Finally, through the entire invasion process, there are many casualties when individuals attempt to overcome each barrier, resulting in a reduction of individuals in each step what entails a decrease in the probability of establishing in a non-native area (Lockwood et al. 2005). Regarding on the strategies of management of alien species, the degree of impacts of alien species often increases through the whole invasion process whereas their management result to be progressively less efficient (Fig. 1).
Figure 1. The conceptual framework for the invasion process used throughout this thesis (adapted from Blackburn et al. 2011). This framework is useful for human-mediated invasions. Invasion process is a multi-step process with four stages (transport, introduction, establishment and spread) separated by certain barriers in each stage. An invasion is successful whenever individuals are able to cope to the various barriers and overcome them (white arrows). On contrary, it fails when individuals are unable to reach the following barrier (black arrows). Secondary introduction occurs when introduced, established or dispersal individuals are transported from an invaded area to a new one owing to overtake the transport barrier again. To the extent that invasion process moves forward across the several barriers, the number and strength of impacts increases; conversely, the management becomes progressively less efficient.
11 General introduction
GENETIC PARADOX?
Generally, only few captured individuals in the native range are able to cross the whole invasion process, be established and thrive into a novel environment (Williamson
1996). Overtaking all stages implies to have overcome a set of barriers where many individuals can fail (Fig. 1). This decrease in population size entails a demographic bottleneck giving rise to the reduction in genetic variation, inbreeding, genetic drift and a reduction in fitness of the introduced population (Allendorf & Lundquist 2003; Facon et al. 2006). This genetic impoverishment should hinder the ability of introduced populations to evolve (Fig. 2 and Blackburn et al. 2015; but see Kolbe et al. 2004); nevertheless, they are often able to flourish, adapt and become invasive. This phenomenon was termed as genetic paradox, but there is, however, a growing debate about how some invasive populations do not undergo such a reduction in genetic variability (Roman & Darling 2007; Selechnik et al. 2019). In a recent review, Estoup et al. (2016) pointed out that if such a genetic paradox exists, a series of characteristics must be accomplished in the invasive population (e.g., to be able to adapt to novel environment in spite of undergoing inbreeding and the effects of low genetic variation as consequence of a strong genetic bottleneck), otherwise, no genetic paradox could be considered. It is noteworthy that, although an initial introduced population does not undergo a true genetic paradox (Estoup et al. 2016), successive bottlenecks can take place along the invasion process with each secondary introduction event, leading to a progressive genetic impoverishment (Fig. 2). Whether this low genetic variability affects to the invasiveness of species is a key evolutionary question in invasion science considering that, for example, marmorkrebs is a parthenogenetic crayfish that produces genetically uniform offspring with a highly invasive potential (Andriantsoa et al. 2019).
However, multiple introductions (from same or distinct sources), the release of large numbers of individuals (i.e., high propagule pressure) and genetic admixture (i.e., interbreeding between individuals from genetically distinct source populations within a species) can help to maintain or increase the genetic diversity of a founder invasive population, thereby favouring their invasiveness (Sakai et al. 2001; Dlugosch & Parker
12 General introduction
2008; but see Rius & Darling 2014) (Fig. 2). Also, it is increasingly recognised that successful invasions distributed globally often involve previous successfully invasive populations, which act as sources of invaders to other new areas (i.e., the invasive bridgehead effect) (Lombaert et al. 2010). Additionally, other mechanisms can explain how invasive species overcome the genetic depression such as inbreeding x environment interactions (Schrieber & Lachmuth 2017) and epigenetic changes that can lead to rapid evolutionary changes in the individual phenotypes, enhancing their invasion success (Pérez et al. 2006). Therefore, multiple mechanisms and forces are involved within the invasion process that contribute to the resulting low or high genetic diversity in invasive populations, altering their invasiveness in greater or lesser extent.
Figure 2. Factors contributing to the genetic diversity of invasive populations, modified from Roman & Darling (2007). Source of individuals can come from the native range as well as one or several previous invaded areas. Admixture is a driver of genetic diversity where population structure hardly occurs, hence, the more admixture in the source, the highest diversity will have their populations in overall. Invasive bridgehead effect is a successful invasive population that act as source of other introductions, having probably positive effects on the invasiveness of new invaders. Propagule pressure represents another driver of introduction success because the more size of inoculum and number of events, the greatest is the probability of introduction success. All of these drivers have consequences on the extent of the bottleneck and, hence, on the resulting genetic diversity of the introduced population.
13 General introduction
Understanding the invasion patterns through genetics approaches has helped to respond key evolutionary issues in invasion science. In fact, molecular methods have been well recognised as powerful tools for investigating biological, ecological and evolutionary aspects of invasions (Miura 2007). Genetic characteristics of an invasive species can provide important opportunities to evaluate the potential for adaptation in the new habitat but also support which source populations are and reconstructions of invasion histories. This information can be crucial for preventing future invasions and conserving the native biodiversity (Estoup & Guillemaud 2010).
As invasive species are translocated by humans, invasive routes can sometimes be traceable by historical reports, grey literature, scientific studies or observational data but often result hard to reconstruct. In this regard, unravelling the introduction routes using genetic tools is extremely useful not only to understand invasion process but also to manage invasive species, examining the genetic variability in both the native and the invasive populations (Cristescu 2015; De Kort et al. 2016). Besides of identifying invasion routes (De Kort et al., 2016), genetic approaches are also vital to elaborate environmental
DNA protocols (Ficetola et al. 2008b), to assess propagule pressure in the invasion process (Ficetola et al. 2008a), and to analyse how invasive species are genetically structured in the native and invaded ecosystems (Dlugosch & Parker 2008). Thus, phylogeographic studies have been proposed as an integral tool of invasion science, thereby helping to biodiversity conservation planning (van de Crommenacker et al.
2015).
IMPACTS AND MANAGEMENT OF INVASIVE SPECIES
Freshwater ecosystems are among the most threatened ecosystems on Earth, due to the high intensity of human disturbances they suffer (Sala et al. 2000). There are many relevant drivers of freshwater diversity decline (Reid et al. 2018). Among all drivers, habitat degradation (e.g., dams, canalization and water pollution) and overexploitation of targeted fish and invertebrate species are the major threat categories for freshwater ecosystems (see Dudgeon et al. 2016; Reid et al. 2018). Because of the high connectivity
14 General introduction among inland water systems and human disturbance, they are particularly affected by biological invasions (Strayer 2010; Havel et al. 2015). Freshwater invasive species generate drastic impacts across aquatic food webs, decreasing the abundance and diversity of native species (Gallardo et al. 2016). While fish are the most commonly alien species translocated outside their native ranges, invertebrate species are also highly reported as introduced, having established self-sustaining populations (Strayer 2010). In particular, several freshwater crayfish species have been globally introduced (Lodge et al. 2012). They are the largest invertebrates living in freshwater ecosystems and often reach high density causing severe direct impacts on freshwater biodiversity
(Twardochleb et al. 2013) but also, indirect impacts through diseases such as the oomycete disease, Aphanomycosis (Aphanomyces astaci) (Diéguez-Uribeondo &
Söderhäll 1993), the White Spot Syndrome Virus (WSSV) (Yi et al. 2017) or to be a potential vector of amphibian chytrid fungus (Batrachochytrium dendrobatidis, Bd)
(McMahon et al. 2013).
Regarding on the impacts over non-native ecosystems, large amount of money has been invested in the mitigation of invasive species (Lovell et al. 2006). Indeed, there is a conservative estimate that invasive species cost €12 billion per year in damages to the
Member States of the European Union (Kettunen et al. 2009). Therefore, there is a need for narrowing the “knowing-doing” gap between invasion research and invasive species management (Esler et al. 2010), and expanding multidisciplinary role for invasion science is key to achieve it (Ricciardi et al. 2017).
With the aim of reducing negative impacts of invasive species, management interventions have been developed and implemented worldwide (Pyšek & Richardson
2010). However, even if invasive species often cause negative impacts (see paragraph above), simultaneously, they can provide benefits for human being, especially those species introduced purposefully for aquaculture, game species, pets, etc. Human interests on invasive species influence their management, often leading to conflicts among different groups of stakeholders (Estévez et al. 2015; Crowley et al. 2017; Vaz et al. 2017) and hampering invasive species management actions (Gárcia-Llorente et al.
15 General introduction
2008; Shackleton et al. 2019). To minimize the impact of conflicts related to management of invasive species, conceptual frameworks have been developed for engaging stakeholders (Novoa et al. 2018).
THE RED SWAMP CRAYFISH
The most common freshwater invasive species often are fish and several invertebrates (Gherardi et al. 2009; Strayer 2010), which usually are highly valued species, either for human consumption or as food for other cultured animals (e.g. shrimp or prawn for fishes, crayfish for bullfrogs, etc.), providing a high economic return (Resh
& Rosenberg 2015). In particular, freshwater crayfish are favoured for farming since they do not have a larval phase and are polytrophic, and they are relatively easy to rear compared with other cultured crustaceans (Holdich 1993). In the wild, freshwater crayfish are keystone species that can exert many pressures on aquatic ecosystems at all trophic levels (Power & Tilman, 1996).
The red swamp crayfish, Procambarus clarkii (Girard 1852), native to southern
United States and north-eastern Mexico, has been successfully introduced into all continents except Australia and Antarctica mainly due to its economic value (Hobbs et al. 1989). Owing to its biological and ecological characteristics, this crayfish is considered one of the worst invasive species worldwide, causing serious direct impacts to biodiversity (e.g. other crayfish species, fish, amphibians, macroinvertebrates and macrophytes) (Geiger et al. 2005; Twardochleb et al. 2013), ecosystem functioning (Usio et al. 2006), human infrastructure and ecosystem services (e.g., irrigation canals, water quality, rice crops, etc.) (Souty-Grosset et al. 2016). Regarding on indirect impacts (e.g., parasite-mediated) of the red swamp crayfish, crayfish plague caused by the oomycete
Aphanomyces astaci has been well studied for long due to its devastating impact over
European crayfish species, causing massive decline in populations (Holdich et al. 2009).
But little is known about other indirect impacts such as potential role of the invasive red swamp crayfish as reservoir of the amphibian chytrid fungus (indirect effect), that is crucial for future conservation strategies on amphibians. Finally, the red swamp crayfish
16 General introduction is one of the most economically valuable aquatic species (Huner 2002; Souty-Grosset et al. 2016), generating tens of billions of US dollars (USD) per year in the world. Therefore, conservation conflicts arise among stakeholders, giving rise to ups and downs in law legislation about invasive species and particularly the red swamp crayfish.
17
OBJECTIVES
In an effort to fill important gaps in biological invasion research, the overall aim of this thesis is to analyse the complex invasion process of the red swamp crayfish as well as some effects on the biodiversity and the conflicts associated to invasive species management. Hence, this thesis comprises different aspects of the invasion of the red swamp crayfish: introductions, impacts and management (Fig. 3).
Figure 3. General framework of this thesis. Each section represents one of the key points in the invasion process: 1) introductions; 2) impacts and; 3) management of invasive species, in this case, taking the red swamp crayfish, Procambarus clarkii, as study model.
First of all, we use multidisciplinary approaches such as historical reports, literature research, grey literature and genetic tools in order to elucidate the invasion history of the red swamp crayfish, Procambarus clarkii (Chapters 1-3). By using georeferenced records from global databases and doing an exhaustive literature search, the Chapter 1 explores the global invasion history of the red swamp crayfish throughout the last century. The main goal of this chapter is to understand how the global expansion of the
19 Objectives red swamp crayfish has been carried out, identifying its pathways and sources of introduction over last one century. Once global distribution was known, the Chapter 2 infers genetically the main routes of introduction in the North Hemisphere focusing on
United States, Japan, China and Europe. Furthermore, this chapter examines the genetic variability and population structure in both the native and non-native areas in order to reveal potentially unreported introductions not cited in the literature. After unravelling the main routes of introduction in the North Hemisphere according to literature and genetics, the Chapter 3, based on genetics, unveils two other unconfirmed routes of introduction into Europe. These three chapters shed light on the invasion history of the red swamp crayfish, describing its invasion routes and the main triggers of its great expansion over the world.
Secondly, the study of impacts on invaded ecosystems is crucial for conservation actions. Many direct impacts have been associated to this invasive species, however, little is known about indirect impacts such as potential transmission of emerging diseases. The Chapter 4 focuses on the role of the red swamp crayfish as potential vector of amphibian chytrid fungus (Batrachochytrium dendrobatidis) which is causing severe decline in amphibian populations. In particular, this chapter analyses the prevalence and infection intensity of Bd in the invasive red swamp crayfish across western
Andalusia (Spain); and also assesses whether the presence of crayfish affects the prevalence and infection intensity of Bd in amphibians of Doñana Natural Space, a highly protected area in Europe.
Thirdly, conservation conflicts often occur when invasive species entail both socioeconomic benefits and severe environmental impacts. As the red swamp crayfish is an economic value species, the Chapter 5 explores the particular social conflict that the commercial use of this species generate among conservationists, crayfish industry and legislation in southwestern Spain. This chapter addresses the specific case of how simplistic and rigid laws can hinder the invasive species management.
20
CHAPTER I
One century away from home: how the red swamp crayfish
took over the world
Oficialdegui FJ, Sánchez MI, Clavero M
Submitted to Reviews in Fish Biology and Fisheries
Chapter I
ABSTRACT
The red swamp crayfish (Procambarus clarkii), native to southern United States and north-eastern Mexico, is currently the most widely distributed crayfish globally as well as one of the invasive species with most devastating impacts on freshwater ecosystems.
Reconstructing the introduction routes of invasive species and identifying the motivations that have led to those movements, is necessary to accurately reduce the likelihood of further introductions. In this work, we: i) review the temporal evolution of the scientific literature on the red swamp crayfish; ii) compile georeferenced, time- explicit records of the species to provide a comprehensive understanding of its global expansion process; and iii) evaluate the role of biological supply companies in the translocations of the red swamp crayfish. The interest of the red swamp crayfish in scientific research increased steadily since the begging of the 20th century to be stabilized since the late 1960s. The number of studies related to the use of this species in aquaculture show two peaking periods, one between the mid-1960s and the early 1980s and a current one, after a continuous increase of the scientific production in this area since the mid-1980s. Research on the red swamp crayfish as an invasive species has only been numerically relevant in recent times, with the number of papers increasing since the 2000s to represent currently around 25% of the scientific production dealing with this species. The first introductions of the red swamp crayfish took place in the 1920s, but our synthesis highlights the rapid expansion of the species occurred from the 1960s, arguably promoted by the rise of crayfish industry. Commercial suppliers from native
(Louisiana) and non-native (California, North Carolina) areas in the U.S.A. have provided live red swamp crayfish specimens for scientific research around the world for decades, suggesting that the invasion process of the red swamp crayfish is more complex than generally assumed.
24 How the red swamp crayfish took over the world
1. BACKGROUND
Humans have transported plants and animals across biogeographical barriers for millennia, with cultural, leisure or commercial purposes (Forcina et al. 2015), albeit this movement of organisms has steeply accelerated since the mid-20th century (Capinha et al. 2015). When released into new areas, some of those transported species are able to survive, reproduce and stablish self-sustaining populations, becoming invasive species
(Blackburn et al. 2011). Invasive species are now a widespread conservation issue and their impacts are considered one of the biggest threats to global biodiversity (Bellard et al. 2016). Identifying the invasion routes through which species are either transported from the native areas to non-native ones or moved among non-native areas is crucial to prevent further spread and to manage future emerging invaders (Estoup & Guillemaud
2010; Bertelsmeier et al. 2018).
Freshwater ecosystems are amongst the most severely threated in the world, due to the combination of habitat degradation, hydrological alteration, global warming, overexploitation, pollution and invasive species (Reid et al. 2018). As a consequence of all these pressures, freshwater biodiversity is currently declining at a much faster rate than terrestrial or marine environments, with population declines estimated at 80% since
1970 (WWF 2016). Freshwater ecosystems are among the most invaded ecosystems in the world and are particularly susceptible to the impact of invasive species (Ricciardi &
MacIsaac 2011; Gallardo et al. 2016). Several groups of organisms are recurrent, often coexisting invaders of freshwater systems, including plants, vertebrates (especially fish) and a wide array of invertebrate taxa (see Gherardi 2007).
Sixteen freshwater crayfish species have been widely introduced into non-native areas worldwide (Logde et al. 2012), some of them being amongst the most impacting invasive species (Twardochleb et al. 2013 and references therein). The magnitude of the impact of invasive crayfish is often related to their frequent role as keystone species in freshwater ecosystems (i.e. due to their high abundances, large size, wide range of trophic interactions and their role as ecosystem engineers), affecting to both lower and upper trophic levels (Geiger et al. 2005; Reynolds et al. 2013).
25 Chapter I
Freshwater crayfish are well-known and exploited by humans in all the regions of the world where they live (Gherardi 2011). Their accessibility and nutritional value
(Tricarico et al. 2008) have contributed to make crayfish a relevant food item for many societies (Holdich 1993; Swahn 2004; Gherardi 2011; Patoka et al. 2016) and a source of economic development (Comeaux 1978; Gutiérrez-Yurrita et al. 1999). The use crayfish as food is in the roots of several cultural traditions, such as the Swedish crayfish summer festivals, in which families and friends gather to eat crayfish (Edsman 2004; Swahn
2004). Being appreciated and easily transported organisms (crayfish can survive prolonged periods out of water, Gherardi & Barbaresi 2000), crayfish species have been introduced into new areas for long (Machino & Holdich 2006; Hobbs & Lodge 2010). In
Europe, crayfish introductions have occurred at least since the Middle Ages (e.g. Gouin et al. 2003; Swahn 2004; Gherardi 2011). For example, Carl Linnaeus reported the introduction of the noble crayfish (Astacus astacus) to Sweden, promoted by King John
III in the second half of the 16th century (Hobbs et al. 1989), coinciding in time with the importation of the Italian crayfish (Austrapotamobius italicus) from Tuscany to Spain, a personal initiative of King Philip II to imitate the uses of the Tuscan court (Clavero et al.
2016).
North America possesses the largest diversity of freshwater crayfish in the world
(382 species, Crandall & Buhay 2007), but little is known on crayfish uses by aboriginal
North American inhabitants (Huner 2002). First European settlers noticed the presence of crayfish (e.g., they were already cited by Aldrovandi 1606) and crayfish could be found in some North American markets since the early 19th century (Comeaux 1978).
By the early 20th century, three main crayfish industries had been developed in North
America, targeting three different genera, namely Orconectes (in the Middle West),
Pacifastacus (Pacific Northwest) and Procambarus (Louisiana) (Comeaux 1978). These are nowadays the most widely introduced genera worldwide and the ones producing the highest biodiversity impacts (Twardochleb et al. 2013). The introduction of North
American crayfish into other continents took place at least since the late 19th century, when the spiny-cheek (Orconectes limosus) and the virile (O. virilis) crayfish were
26 How the red swamp crayfish took over the world introduced in Europe (Hobbs et al. 1989). But the most striking invasion process is that of the red swamp crayfish (Procambarus clarkii), which started in the early 20th with the introduction, of possibly hundreds of crayfish, in California and is currently the most cosmopolitan freshwater crayfish, distributed across all continents except Australia and
Antarctica (Loureiro et al. 2015).
The origins of exploitation of the red swamp crayfish is linked to the Cajuns, descendants of the French colonists in Acadia, north-eastern North America, who settled in Louisiana in the late 18th century (Gutiérrez 1998). The Cajuns’ customs, including the French taste for crayfish, gradually settled down in Louisiana and the commercial exploitation of the red swamp crayfish started growing since the late 19th century
(Gutiérrez 1998; see in Brady 2013). The first fishermen harvested crayfish from wild stocks from swamps and marshes in south Louisiana, but water bodies were soon modified or constructed to store catches and allow longer harvesting periods, developing the aquaculture-based crayfish industry (Comeaux 1978). Crayfish production steeply increased in the 1960s, due to the transformation of several lands to that aim, often in combination with rice cropping (i.e., rice-crayfish fields) (Huner 2002).
Land devoted to crayfish production increased from 400 ha in 1959 to 10,000 ha in 1970
(Clark & Avault 1975) and up to 49,000 ha in 1990 (LSU AgCenter 2016). The Louisiana crayfish industry became the most successful producer and seller of crayfish in North
America (Comeaux 1978) reaching a farm-gate value of more than $200 million
(aquaculture plus wild harvested) in 2016 (LSU AgCenter 2016).
The high profitability of the red swamp crayfish industry led several entrepreneurs to try to replicate its aquaculture-based production in other areas (Hobbs et al. 1989;
Huner 2002; Cheung 2010; Brady 2013). Transcontinental movements of the red swamp crayfish to Africa and Europe gave rise to incipient crayfish industries in countries such as Kenya or Spain (Harper et al. 2002, for Kenya; Gutiérrez-Yurrita et al. 1999, for Spain).
However, the most striking growth of crayfish production took place in China, which has recently overtaken the native production of Louisiana crayfish industry. Chinese production has increased from 6,700 tonnes in the early 1990s (Xia 2007) up to more than
27 Chapter I one million tonnes in 2017, with an actual commercial value of $42 billion nowadays
(China’s Ministry of Agriculture and Rural Affairs 2018).
Here, we review the century-lasting invasion history of the red swamp crayfish in order to describe its expansion, update the knowledge on its global distribution, report the main introduction routes and discuss the main pathways driving the translocations of this species. Based on a review of scientific and grey literature, as well as a collection of records worldwide, we (1) describe the historic variation in the research scope of the red swamp crayfish from the early 20th century to the present as well as the patterns of knowledge production in the red swamp crayfish, (2) make a thorough description of introduction and expansion events along the last one century, and (3) explore the role of commercial companies in the expansion of this species. Commercial companies that ship live specimens for different purposes (e.g., aquarium hobby, education or research) may represent a relevant, though overlooked introduction vector of the red swamp crayfish worldwide (Chucholl 2013). Information related to aquarium species and pet trade is scarce and often inaccessible (see Chucholl 2013), but researchers usually report the provenance of their model organism in scientific articles. This information could be a useful proxy for the potential role of commercial companies in the translocation of the red swamp crayfish, and other organisms, around the world.
2. METHODS
To assess the evolution of knowledge production involving the red swamp crayfish we conducted a keyword-based search using the ISI Web of Science (WOS). On April
15th 2019 we searched for the terms “Procambarus clarkii” OR “cambarus clarkii” OR
“procambarus (scapulicambarus) clarkii” in the title, abstract and keywords, and counted the number of papers resulting from the search for each year up to 2019. To refer the results of our search to the temporal patterns in the overall scientific production, we also compiled the yearly production of scientific papers in a pool of disciplines and themes that could involve crayfish-based research, by using the search (*zoolog* OR *ecolog*
OR *toxicol* OR *biolog* OR *neurolog* OR *invasi* OR "pet trade"). Combining the two
28 How the red swamp crayfish took over the world searches, we calculated for each year between 1924 (the year of the first introduction of the red swamp crayfish outside of its native range) and 2019 the number of papers dealing with the red swamp crayfish for every 10,000 scientific papers. We then assessed the temporal variation of the scope of the research involving the red swamp crayfish, with a focus on the disciplines related to the introduction and invasive character of the species, namely aquaculture/fisheries and invasion science. To do so, in addition to the synonymous scientific names (see above), we added the following terms in our search:
AND aquacul* OR astacicul* OR fisher* (for aquaculture/fisheries) and AND invasi* (for invasions).
In order to describe the progress of the global invasion of the red swamp crayfish during the last century, we collected spatially-explicit records of the species in both native and non-native areas. The search included the review of the existing scientific and grey literature as well as a compilation of geo-referenced, time-explicit records from on- line repositories of biodiversity data, namely the Global Biodiversity Information
Facility (GBIF, www.gbif.org), the U.S. Geological Survey (USGS, www.usgs.gov) and the iNaturalist platform (iNaturalist, www.inaturalist.org). Whenever a record from the literature referred to a political entity (region or county) instead to a specific locality, we assigned the record the coordinates of the centroid of the political entity. We split the records of red swamp crayfish in four time-periods: before 1950, 1951-1975, 1976-2000, and 2001-2019. As result of our search, we present a summary of the global expansion process, but full territory-specific information on this process, including all consulted bibliographic sources, is provided as Supplementary Material.
To assess the source of red swamp crayfish specimens used as study model in scientific articles, a literature search on the ISI Web of Science (WOS) using the topic
“Procambarus clarkii” was conducted as well as doing cross referencing of old literature cited in more recent studies. We focused on the commercial source to evaluate the role of biological supply companies in the shipping of live red swamp crayfish. In the screening process, we used those scientific manuscripts that detailed the source of the red swamp crayfish (provenance from wild and a commercial house or biological supply
29 Chapter I company) in materials and methods. And, we used author’s affiliations, information on place of experiment setup and acknowledgements for destination accuracy, but in case of any doubt, manuscripts were discarded from the literature search. While wild source of the red swamp crayfish (captures from the wild) was added as a record of presence in invaded areas (see paragraph above), the commercial source (crayfish obtained from commercial companies) was used as a proxy of the potential translocations of the red swamp crayfish worldwide because anyone can buy live-specimens from anywhere. As our main interest was to detect first translocations outside of its native range in the beginning of the invasion process, we exhaustively analysed all articles published annually until 1990. However, regarding on the most recent period (1991-present), due to the drastic increase in number of published manuscripts on the red swamp crayfish
(n = 3,924) as well as a more chance to capture wild crayfish nearby because it was already expanded worldwide, we selected available scientific studies published every each five years (1995, 2000, 2005, 2010 and 2015) as an indicator of all studies published over last 30 years.
3. RESULTS AND DISCUSSION
3.1 HISTORICAL VARIATION IN THE RESEARCH SCOPE ON THE RED SWAMP
CRAYFISH
Out of 19,342,413 articles published over the last 95 years (from 1924 to 2019) on zoology, ecology, toxicology, biology, neurology, invasion science and pet trade, 5,442
(<0.03%) dealt with the red swamp crayfish. While the total production of articles has constantly increased since the decade of 1950s, the interest on the red swamp crayfish intensified during the 1960s. Before the early 1960s, the ratio of publication was 1.5 articles on the red swamp crayfish for each 10,000; however, it was duplicated up to three articles in the late 1960s remaining relatively constant since then (Fig. 1a).
In the beginning of the invasion (since 1924 to 1960), there were hardy any scientific studies on the red swamp crayfish and most of them was not related to aquaculture/fisheries or invasions (Fig. 1b). First scientific studies on this species were
30 How the red swamp crayfish took over the world related with the physiology, functioning nervous and motor systems using crayfish as a model with potential applications to knowledge of human locomotion and nervous system (Stark 1968). Physiology studies are still a relevant component of the scientific research focussed on the red swamp crayfish. Articles on the use of the red swamp crayfish as aquaculture species or its potential in fisheries increased in numbers in two peaking periods: one between the mid-1960s and the early 1980s arguably in relation to the growing commercial use of the red swamp crayfish, reaching up to a 75% out of total number of articles on the red swamp crayfish in the decade of 1970s, and a current one, after a continuous increase of the scientific production in this area since the mid-1980s.
The number of studies dealing with the role of the red swamp crayfish as an invasive species has notably increased since the 2000s, reaching around 25% of total studies in the decade of 2010s.
3.2 THE INVASION HISTORY
After discarding all records with duplicate coordinates within a same year, our final dataset included 6,924 red swamp crayfish records (n = 48; n = 271; n = 923; and n = 5,682, for the four time periods considered, Fig. 2 and Fig. 3). The number of records grew progressively since the beginning of the invasion but there was a striking increase in number of records since the decade of 1990s (Fig. 2a), mainly associated to exhaustive surveys carried out in the native area and non-native areas such as America and Europe
(Fig. 2b). A clear example of this sudden increase is the number of records in Europe that passed from three records in the second period up to 2,710 records in fourth period
(Fig. 2b) due to the great expansion over the continent since its introduction in the 1970s.
However, the number of records of the red swamp crayfish remained low in Africa (<
1% of total records) and Asia (< 5% of total records) (Fig. 2), arguably due to pervasive spatial biases in the production of species occurrence data, which are common to historical and current datasets (e.g., Boakes et al. 2010). Therefore, if we take into account the available African and Asian occurrences of the red swamp crayfish, it is likely that its distribution area in Africa and Asia throughout the last century was underestimated.
31 Chapter I
Figure 1. Dynamic of articles published on the red swamp crayfish over the last ten lustrums (five- year periods) from 1925 to 2019. (a) Black line depicts the number of scientific manuscripts according to the categories (zoolog* OR *ecolog* OR *toxicol* OR *biolog* OR *neurolog* OR *invasi* OR "pet trade"). For a better interpretation, number of articles published on the red swamp crayfish were multiplied by 10,000 and grey dashed line represents the curve fit on the ratio (ratio = n * 10000 / N ) as the number of articles on the red swamp crayfish divided by the total number of scientific articles. The scientific search was based on title, abstract or keywords. (b) Percentage of published articles on the red swamp crayfish according to two main thematic categories. Total number of articles based on the red swamp crayfish for each lustrum is indicated on top of the graph. 32 How the red swamp crayfish took over the world
THE BEGINNINGS (BEFORE 1950)
The red swamp crayfish was cited outside of its native range for the first time in southern California in 1924 (Holmes 1924) and it was translocated from there to Oahu
Island, Hawaii in 1923 or 1927 (see Brock 1960; Brasher et al. 2006) and several more times to other Hawaiian Islands afterwards (Penn 1954). Few years later, one introduction event took place from Louisiana to Japan in 1927 or 1930 (see references in
Kawai 2017 and Penn 1954, respectively) and from there to China in 1930 (see Cheung
2010) (Table 1). It is noteworthy that there is a lack of exactness in the first introduction dates of both introductions (for California and Japan). As far we know, the first report which cited the presence of the red swamp crayfish in California was in 1924 (Holmes
1924), however, live-specimens of the red swamp crayfish were translocated from
California to Hawaii in 1923 (see Brasher et al. 2006) what supposes that the red swamp crayfish must have been introduced into California earlier than previous literature cited or first introduction in Hawaii was later. Although the first introduction in Japan seems to be the same invasion history and is well detailed in literature, there is an incongruence about the exact introduction date because some authors cited it in 1927 (Kawai 2017) and others in 1930 (Penn 1954).
However, there is a great consensus on the motivation to translocate live-specimens of the red swamp crayfish in California, Hawaii and Japan, which were intended to provide food for culturing the American bullfrog (Lithobates catesbeianus) (Hobbs et al.
1989). In spite of the red swamp crayfish was expanded across rice fields in California
(Riegel 1959), various Hawaiian islands (Penn 1954) and the Honshu Island in Japan
(Kawai 2017) and considered as a pest due to its burrowing activity (see Penn 1954), there was a time-lag between its introduction (1924) and the action measures to
‘eradicate’ them by mid-twentieth century (Chang & Lange 1967). On the other hand, the red swamp crayfish was introduced into China in 1930 short after its introduction to
Japan (Table 1), by Japanese civilians who presumably used the species as pets (Cheung
2010). On how the red swamp crayfish settled in China, Cheung (2010) described that the apprehension of Chinese society to everything that came from Japan in the beginning
33 Chapter I of the red swamp crayfish invasion could have stopped its expansion to other areas nearby. There was such an extent of concern, that Chinese people thought that the introduction of the red swamp crayfish was a Japanese conspiracy to harm their rice fields. In fact, Chinese population neither appreciated the crayfish nor considered it edible by mid-twentieth century (Cheung 2010), a rejection that probably also limited the expansion of the red swamp crayfish across China in the first decades after its introduction (Xinya 1988) (see in Fig. 3).
Figure 2. Red swamp crayfish records along last century. (a) Decadal evolution in the total number of records (black line) and number of records for different biogeographical areas (note logarithmic scale of Y axis). (b) Proportion of total number of records for the four time-periods used in the presentation of our results, showing total numbers for each biogeographical area: native area, non- native area in America, Asia, Africa and Europe.
34 How the red swamp crayfish took over the world
35 Chapter I
THE OUTBREAK OF RED SWAMP CRAYFISH INDUSTRY (1951-1975)
Coinciding with the rise of the Louisiana crayfish industry around 1960s (LaCaze
1970; Gary 1974), there were numerous attempts to emulate that production system through translocations of the red swamp crayfish to different areas (Fig. 2), either from
Louisiana or from another regions (see Table 1), even though wherever the red swamp crayfish was introduced, farmers usually tried to control its populations (Hobbs et al.
1989). The species was taken with that purpose to Africa (Sudan, Kenya) and Europe
(Spain) in the early 1970s, and by 1975 the exploitation of the red swamp crayfish had started to gain importance in different non-native areas, including states of U.S.A. (e.g.,
California, see in Huner 1977) and countries such as Kenya, Spain, France and Italy (see
Supplementary Material). But introductions also involved other purposes, such as mitigation of schistosomiasis (e.g., Uganda and Kenya, Hofkin et al. 1991) or supplying the pet market (e.g., Hong Kong and Taiwan or France and Finland, Hobbs et al. 1989).
The motivation for other many introductions remains unclear (e.g., different States of
U.S.A. and Mexico, South Africa or Costa Rica) (see Supplementary Material). Apart from the new introductions, the red swamp crayfish continued expanding in the territories were it had been introduced before 1950, notably in Western U.S.A. and Japan
(Fig. 3).
THE GREAT SPREADING WORLDWIDE (1976-2000)
In the late 20th century, there was an acceleration of the expansion of the red swamp crayfish in several non-native areas, including Europe (Gutiérrez-Yurrita et al.
1999; Changeux 2003), China (Xinya 1988), non-native areas in the U.S.A. (Hobbs et al.
1989) and Kenya (Harper et al. 2002). In the last quarter of the 20th century, the red swamp crayfish also arrived to different countries in South America (Colombia,
Ecuador, Venezuela), the Caribbean (Dominican Republic, Puerto Rico), and Africa
(Zambia, Egypt) (Fig. 2). In Europe, multiple secondary introductions drove the rapid expansion of the red swamp crayfish over Spain, Portugal, Italy and France (see
Oficialdegui et al. 2019a), as well as its arrival to Germany, Belgium, the Netherlands,
36 How the red swamp crayfish took over the world
Switzerland, United Kingdom and several European islands (e.g., Cyprus, Balearic and
Canary Islands in Spain, and Azores in Portugal) (see Supplementary Material). Besides, numerous importations of live-specimens took place from Spain and Kenya to French and Italian farms as well as English restaurants since late 1970s to early 1980s (Holdich
1993; Laurent 1990), which could have generated escapes or releases into the wild
(Oficialdegui et al. 2019b). By the late 1990s, the red swamp crayfish was the most important farmed freshwater crayfish species in Europe (54.6% of the total European production), being mainly farmed in Spain (Ackefors 1998) but also in Italy (D’Agaro et al. 1999). Moreover, the red swamp crayfish was highly exploited for recreational fishing
(Changeux 2003) and human consumption in France (Holdich 1993).
Interestingly, although the red swamp crayfish was present in China since 1930, only since the early 1980s Chinese scientists initiated aquaculture experiments aimed at setting up crayfish industry (Xinya 1988). The rapid advance of these initiatives, together the growth of commercial sales in pet shops, caused the spread of the red swamp crayfish across eastern China (Cheung 2010). Thus, the expansion of the red swamp crayfish in China had a delay of more than 50 years since its introductions and establishment. Time-lags are a common feature of several invasion processes (Crooks et al. 1999, Clavero & Villero 2013). In Africa, the main crayfish areas were Lake Navaisha and several watercourses in Kenya (Harper et al. 2002) and the Nile Delta in Egypt
(Hamdi 1994). Simultaneously, many other countries (e.g., Puerto Rico, Dominican
Republic, Ecuador, Zambia, among others) attempted to culture the red swamp crayfish, carrying out experiments on its adaptability and suitability indoor or directly in semi- natural areas, often involving escapes or releases into the wild (see Supplementary
Material).
37 Chapter I
Table 1. First reports of red swamp crayfish, Procambarus clarkii, over the world. Number in brackets indicates the total number of countries or states where the red swamp crayfish is established or probably established. (-) means unknown data. Italics indicate no confirmed information. * indicates eradicated into the country (Israel). References for first records are included in Appendix II.
Country Date Site of introduction Source Purpose AFRICA (8) Egypt 1980s Giza/Cairo/Nile Delta United States Aquaculture Kenya 1966 Solai/Subukia Kajansi (Uganda) Aquaculture /Disease Morocco 2008 Merja Zerga Seville (Spain) Aquaculture Rwanda 2019 Kigali - - South Africa 1962 Potchefstroom - Aquaculture Sudan 1975 Khartoum Louisiana (US) Aquaculture Uganda 1963 Kajjansi Louisiana (US) Aquaculture/Disease Zambia <1979 Livingstone Naivasha (Kenya) Aquaculture AMERICA (10) Brazil <1986 São Paulo United States Pet trade Colombia 1985 Cauca Valley - Aquaculture Costa Rica 1966 Alajuela City - - Dominican Republic 1977 Santo Domingo United States Aquaculture Ecuador 1986 Taura River - Aquaculture Guatemala 2019 Técpan - - Mexico 1955 Cananea - - Puerto Rico <1978 - - Aquaculture Venezuela 1978 - Louisiana Aquaculture US (39) Alabama 1961 Auburn - Aquaculture Alaska 2004 Kenai - - Arizona 1969 Lower Colorado Basin - - California <1924 Pasadena Louisiana - Colorado 2018 Denver - - Connecticut 2017 Near Norwich - - Delaware 2018 Brandywine Creek - - Dist. of Columbia 2016 Anacostia River - - Florida 1951 Hudson Louisiana Aquaculture Georgia 1989 Athens - - Hawaii 1923 Oahu island California Food source Idaho 1975 Nampa Nevada/California - Illinois 2001 Chicago River - - Indiana <1986 - - - Kansas 2017 Kansas City - - Kentucky <1944 - - - Maine 1980 Kennebec River - - Maryland 1963 Patuxent Area Louisiana Food source 38 How the red swamp crayfish took over the world
Table 1. (Cont.) Country Date Site of introduction Source Purpose Massachusetts 2010 Amherst - - Michigan 2013 Holland - - Minnesota 2016 Tilde Lake - - Missouri 2009 Table Rock Reservoir - - Nebraska 2014 Missouri River - - Nevada 1944 Las Vegas River California - New Jersey 2016 Saxton Lake - - New Mexico 1944 Grande River - - New York State 2002 Long Island - - North Carolina 1980s - - - Ohio 1967 Sandusky Bay - Fishing Oklahoma 1969 McCurtain Co. - - Oregon 1990s Willamette Valley - - Pennsylvania 1990 Schuylkill River - - Rhode Island 1970 Arcadia - - South Carolina <1978 - Louisiana Aquaculture Tennessee 2018 Nashville - - Utah 1978 Tooele Co. - - Virginia 1972 York-Pamunkey - - Washington State 2000 Pine Lake - - Wisconsin 2009 Kenosha Co. - - ASIA (8) China 1930 Nanjing Japan Pets Hong Kong <1960s Hong Kong - Pet trade Indonesia 2018 Java Island - Pet trade Israel* 2008 Hadera - - Japan 1927/1930 Ōfuna/Kamakura New Orleans (US) Food source South Korea <2005 Incheon - Pet trade Taiwan 1960s - Hong Kong Aquaculture/Pet trade Thailand 1987 Chiang Mai province United States Aquaculture EUROPE (12) Austria <2005 Salzburg - - Belgium 1983-85 Vielsalm - Human consumption Cyprus <1987 Athalassa dam - - England 1991 Hampstead Heath Park Kenya Human consumption France 1974 Charente-Maritime Spain/Kenya Aquaculture Germany 1975-76 Lake Hechtsee - - Italy 1977 Banna Stream Spain Aquaculture Malta 2016 Fiddien Valley China Pet trade/Aquaculture Portugal 1979 Caia River Badajoz Natural dispersion Spain 1973 Badajoz Louisiana Aquaculture Switzerland 1989 Schübelweiher - Fishing The Netherlands 1985 The Hague - Human consumption 39 Chapter I
CURRENT STATUS (2001-2019)
The red swamp crayfish has recently expanded over areas where it had been previously introduced of western and eastern U.S.A., north-eastern Mexico, European countries, China and, to a lesser extent, other territories (Fig. 3). Secondary, human- deliberated, introductions are keys in the invasion process, where established populations in invaded areas act as source of new introductions at long- and short- distance (see in Oficialdegui et al. 2019a; 2019b). It has also appeared for first time in some parts of Europe (Austria and several Mediterranean islands: Corsica, Sardinia,
Sicily and Malta), Africa (Morocco) and Asia (South Korea, Israel and Indonesia) (Fig.
3). Currently, this species is distributed in 38 countries of four continents (Table 1), but there is still potential for further expansion. New suitable areas to host red swamp crayfish population are predicted according to climate match maps in islands of
Indonesia (see in Putra et al. 2018) as well as distribution models in several yet unoccupied territories in southern South America, the Mediterranean Basin, and large parts of Africa and Australia (Larson & Olden 2012). Once the red swamp crayfish is introduced and established, populations seem to be viable in the long-term (Fig. 3 and
Supplementary Material). In fact, most of previously established populations around the world remain at present (except Alaska in US, Israel and Tenerife Island in Spain, as far as we know). This is an indication that eradication has thus far proven difficult (Gherardi et al. 2011) and calls for an effort to prevent any possible future introduction to new areas.
3.3 COMMERCIAL COMPANIES AS POTENTIAL SOURCES OF INTRODUCTIONS
We were able to identify the provenance of the red swamp crayfish used as model species in scientific articles in 729 out of 2,053 cases studied in selected years (see
Methods for details). Of them, 67% studies obtained them commercially and 33% studies captured the crayfish from the wild. The percentage of crayfish obtained from biological supply companies seems to have declined over time, being 73% of the 456 articles published before 1990 and 56% of the 273 articles published after that date (every five
40 How the red swamp crayfish took over the world years). The recent decrease in commercial-obtained red swamp crayfish in scientific research is arguably related to the increased availability of wild populations nearby due to the continuous expansion of the species since mid-20th century (Fig. 3).
Most of studies based on commercially-obtained crayfish also detailed the commercial company or area from where crayfish were bought. The main suppliers of the red swamp crayfish worldwide were a set of U.S. commercial companies based in the States of Louisiana, California, North Carolina and Wisconsin, which supplied crayfish up to 292 studies (Fig. 4a and Fig. 4b). Until 1990, these four source-states of
U.S.A. provided crayfish to eight countries, and 24 states of U.S.A., including themselves
(Fig. 4a), with an exportation rate of 100% for Wisconsin (n = 6), 92% for North Carolina
(n = 39), 48% for Louisiana (n = 64) and 46% for California (n = 72). From 1991 onwards, the state of Wisconsin has lost its role of main supplier of the red swamp crayfish. And
States of Louisiana, California and North Carolina have provided crayfish to only two countries (Canada and U.S.A.), and to 20 states of U.S.A. (Fig. 4b), with an exportation rate from 79% for North Carolina (n = 14), 74% for Louisiana (n = 62) and 40% for
California (n = 5). Importantly, Japan and China have also become important suppliers of the red swamp crayfish but their exportation rate was very low (Fig. 4a and 4b).
It is noteworthy that most of the most important suppliers of red swamp crayfish worldwide are based in non-native areas within the U.S.A. (e.g. California, North
Carolina and Wisconsin). By analysing source of the red swamp crayfish in scientific manuscripts, our results showed that there have been more introductions than generally assumed (Fig. 4). For example, even though the red swamp crayfish is native from Texas or northern Mexico, several introductions events had place from other invaded areas
(e.g., California or North Carolina). Even scientific studies carried out in Louisiana obtained crayfish from Louisianan and Californian commercial houses. Recently,
Oficialdegui et al. (2019a) showed, based on genetics, two invasion routes in U.S.A.
(westwards and eastwards from the native range) and attributed this invasion pattern to the situation of the native area in the middle of the U.S.A. as well as the role of a commercial company located in North Carolina which had been supplying most of the
41 Chapter I eastern U.S.A. invaded areas. If we focus on the movements within the United States
(Fig. 4), we observe that while commercial houses in California practically sent crayfish to everywhere, commercial houses in North Carolina mainly supplied crayfish to the east of U.S.A., which could explain the low genetic variability found in eastern U.S.A. populations (Oficialdegui et al. 2019a). It is remarkable the large number of shipments of crayfish from diverse areas that have received some states in the north-eastern U.S.A.
(e.g., New York, Massachusetts, Connecticut and Maryland) (Fig. 4). And also, Canada that has received many shipments of crayfish for long (Fig. 4) but wild-populations have been recently detected (iNaturalist 2019). Intercontinentally, we found a great amount of unperceived translocations to Europe where the invasion history of the red swamp crayfish was supposedly well-known (see Supplementary Material). Moreover, while high exportation rates of crayfish were described for commercial houses in U.S.A., most of the shipments of crayfish that took place in Asia, albeit there were a large number, were made within the countries (see Japan and China). And finally, we have found a series of shipments whose suppliers are unknown and their invasion routes cannot be reconstructed. Even though most of specimens used in scientific studies are often sacrificed, before or after the experiments, escapes from research centres have been described in literature (e.g., the exotic mummichog in Spain, Gisbert & López 2007).
Beside of research, other pathways of introductions could remain hidden in the translocation of alien species because the uptake of live-crayfish commercially might be extrapolated to schools and universities (Larson & Olden 2008), general citizens, fishermen or farmers who may obtain live-specimens (Lodge et al. 2000). Therefore, our approximation exposes the risk of shipping highly invasive species out of their native area by showing the amount of translocations that have occurred for long. In this context, scientific studies working with highly invasive species should indicate where live-specimens come from. Hence, particular attention should be paid to introduction routes, whatever the case, of highly invasive species out of their native range.
42 How the red swamp crayfish took over the world
4. MANAGEMENT IMPLICATIONS
Understanding the introduction routes followed by invasive species to arrive to new territories and untangling the motivations that have led to those movements of species is key to reduce the odds of future introductions. This review shows how a multitude of long and short distance translocations, many of them unreported, have shaped the current distribution of the red swamp crayfish, the largest for any freshwater crayfish worldwide. Linking wild occurrences of invasive species with the pathways of introduction such as escapes from aquaculture (Olenin et al. 2008) and the releases from pet trade (Chucholl 2013; Patoka et al. 2015; Faulkes 2015) is crucial to control new emerging alien species used in these human-related activities. The history of the global- scale invasion by the red swamp crayfish can be used as a world benchmark for future invasions involving commercially exploited species and can help managers and policy makers to design and implement efficient management strategies, such as the enforcement of control measures on commercial activities which involve translocations of live specimens. Efforts to control releases of dangerous freshwater species used for aquaculture or ornamental purposes should consider as a major objective to drastically reduce the availability of high-risk species in commercial houses. Furthermore, invasive species policies are generally applied at national or smaller scales, often being inconsistent across countries (Peters & Lodge 2009), when movements of alien species are a global issue (Hulme 2009). More effort should be devoted on the control of aquaculture and trade of invasive species, especially on commercial houses that could help to raise awareness of potential keepers of invasive species which may end up being released into the wild or escaped.
43 Chapter I
Figure 4. Network of the main commercial translocations of the red swamp crayfish (a) since 1961 to 1990; and (b) in 1995, 2000, 2005, 2010 and 2015 based on 334 and 158 scientific papers, respectively. The States with main commercial companies are depicted in the middle of the ellipse and recipient States (abbreviates) or countries (ISO codes) around. UNK shows unknown commercial suppliers. Black, light and dark grey arrows depict the direction and frequency of movements of crayfish: casual (<5), semi-frequent (5-9) and very frequent (>10), respectively.
44 How the red swamp crayfish took over the world
ISO country codes: MEX, Mexico; CAN, Canada; CHN, China; JAP, Japan; DEU, Germany; SWE, Sweden; CHE, Switzerland; CZE, Czech Republic; FRA, France; ESP, Spain; GBR, United Kingdom. Abbreviate United States codes: WI, Wisconsin; CA, California; LA, Louisiana; NC, North Carolina; NH, New Hampshire; MD, Maryland; CO, Colorado; MA, Massachusetts; NJ, New Jersey; NY, New York; PA, Pennsylvania; VA, Virginia; SC, South Carolina; GA, Georgia; FL, Florida; OH, Ohio; MI, Michigan; IN, Indiana; KY, Kentucky; AL, Alabama; MS, Mississippi; TX, Texas; KS, Kansas; MO, Missouri; MN, Minnesota; IL, Illinois; OR, Oregon; WA, Washington.
5. SYNTHESIS AND FUTURE PERSPECTIVES
We have described the global-scale, century-lasting invasion process of one of the most harmful invasive species worldwide. Our review combined literature search and hundreds of records from biodiversity databases to show how and why the red swamp crayfish has expanded its range over the world during the last 95 years, including an exhaustive description on the invasion process in all countries where the red swamp crayfish is, or is suspected to be, established. Finally, we also pointed out some of the potential pathways of introduction for the red swamp crayfish and discussed about the relevant role of commercial suppliers in the translocation of live-specimens worldwide.
Here, we used the red swamp crayfish but our results can be equally useful for any other alien species commercially exploited by humans.
Although we conducted an exhaustive literature search (scientific and grey literature) on the red swamp crayfish, issues associated to old literature (e.g., local language or regional reports are hard to find) could have caused information gaps in some invaded areas resulting in biased or underestimated crayfish distribution.
Concretely, we were unable to find literature or introduction reports in the first 50 years of the red swamp crayfish presence in China, albeit the species was allegedly restricted to the first introduction area (Xinya 1988). Information on red swamp crayfish distribution in Africa seemed to be spatially-biased, because many studies focused the situation in Kenya but introduction reports for other African countries were scarce and sometimes unclear (e.g. South Africa, Sudan or Zambia; see Supplementary Material).
Therefore, further studies on less represented regions (e.g., Asia or Africa) may acquire information from additional sources of species distribution data, for example, museum
45 Chapter I collections which provide an important coverage of species’ ranges mainly for the past species’ distributions (see Boakes et al. 2010) or working with local experts who can supply accurate data on past species distribution. Conversely, there is nowadays a lot of information is available in public databases but occurrences or reports are sometimes incomplete or inaccurate (e.g., imprecise geographical coordinates or lack of verification by experts). Even so, we wish also encourage administrations to develop citizen science projects that involve people in the early detection and spread of invasive species. Early detection and rapid action response is a cost-effective way of preventing establishment of alien species and avoid devastating impacts in the future.
ACKNOWLEDGEMENTS
We are grateful to Biblioteca Campus Cartuja for providing us old literature which was not available online. We also thank David Aragonés from Laboratorio de SIG y
Teledetección (LAST-EBD) at Estación Biológica de Doñana (CSIC). F.J.O. was supported by a grant associated to the project (RNM-936) funded by the Andalusian
Government.
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50 How the red swamp crayfish took over the world
SUPPLEMENTARY MATERIAL
APPENDIX I. INVASION HISTORY BY COUNTRIES AFRICA Established (Egypt, Kenya, Morocco, South Africa and Uganda), probably established (Sudan) and uncertain presence (Gabon).
Egypt The red swamp crayfish was successfully introduced from US in Giza, Cairo and Nile Delta in the early 1980s for commercial aquaculture (Hamdi 1994). Ten years after its first introduction, its distribution had extended from Nile Delta to Assiute and Qena Governorates in the Centre of Egypt (Saad & Emam 1998). A natural colonization from Sudan seems to be no probable because there were no records from the upper Nile or Lake Nasser, South Egypt (Fishar 2006). A consequence of its spread may be due to the possibility of buying live-specimens in markets of Alexandria (Zaglol & Eltadawy 2009). Currently, it is widely established in lower Nile river, mainly in the mouth.
Gabon Although the occurrence of the red swamp crayfish was not detected, there were enquiries on feasibility of culturing the red swamp crayfish in Gabon by Goldschmidt (1995). No updated information was found about the species.
Kenya An unspecified number of the red swamp crayfish from Uganda was originally introduced in 1966 into two dams located at Solai and Subukia, within the Rift Valley (Oluoch 1990). Around 1970, approximately 300 specimens of the red swamp crayfish from the Subukia dam were introduced into Lake Naivasha (Oluoch 1990), where population increased few years later as a potential aquaculture species (Parker 1975; Lowery & Mendes 1977a). In 1975, commercial exploitation began and many exportations to Europe carried out until European banned in 1983 (decree of 21 July 1983) (Gherardi et al. 2011). The red swamp crayfish was expanded within the country during 1970s, leading to the occupation of major river systems (Athi/Galana river, common in the Karen Pools, Nairobi River, Ewaso Ng’iro river, Gathanje reservoir and Nzoia River) with the exception of Tana River, Lake Rudolf and Lake Natron (Huner 1977; Lowery & Mendes 1977b; Harper et al. 2002; Foster & Harper 2006). Introductions were encouraged not only by the possible commercial activity, but also by its assumed role as a biological control on schistosome snail vectors (Hofkin et al. 1991). In 1991, it was abundant in Eldoret river system and, by 2000s, it was expanded to Lake Ol 51 Chapter I
Bolossat, Gilgil and Malewa rivers (Foster & Harper 2007). No updated information was found about the species.
Morocco According to El Qoraychy et al. (2015), the first sighting of this species was in late 2008 early 2009 in Merja Zerga, a permanent biological reserve. The occurrence of red swamp crayfish coincided with the setting, in 2007, of a Spanish company for processing and marketing wild-caught freshwater crayfish. The current distribution of the red swamp crayfish in Morocco has been mainly identified at swamps and rice fields between the provinces of Tanger-Tetouan-Al Hoceima and Rabat-Salé-Kénitra in North Morocco (El Qoraychy et al. 2015).
Rwanda The red swamp crayfish has been recently found in a pond in the surroundings of the capital, Kigali, in 2019 (iNaturalist 2019). No further information about its introduction has been found.
South Africa In 1962, two unconfirmed specimens of the red swamp crayfish were allegedly caught in Potchefstroom near Johannesburg, but no established populations were detected (van Eeden et al. 1983). Despite of their concerns, South African aquarists were rearing the red swamp crayfish illegally and selling in pet shops until 1987, when the Cape Department of Nature and Environmental Conservation confiscated all specimens from pet shops in East London, George, Cape Town and Kimberley (Anonymous 1987). In 1988, an established population was recorded in Driehoek Farm, near Dullstroom (Schoonbee 1993). By ending of 1980s, the species spread over Crocodile River basin and until 1993, when an eradication programme was put into practice (Schoonbee 1993). However, no monitoring was performed until 2016, when Nunes et al. (2017) found again low densities of the red swamp crayfish near Crocodile River.
Sudan In 1975, several hundreds of specimens of the red swamp crayfish were shipped from Louisiana to Khartoum by the Ministry of Agriculture of Sudan to examine the species’ suitability for rearing activities (Huner 1977). This introduction was accomplished by private interests with full government approval (Huner & Avault 1978). However, after the successful commercial boom of the red swamp crayfish in Europe, another event of introduction could have occurred into Sudan from Spain
52 How the red swamp crayfish took over the world
(National Research Council 1976). No updated information has been found about the species.
Uganda As an attempt to control the schistosomiasis snail vector, the red swamp crayfish was introduced in Uganda from Louisiana around 1963 (Hobbs 1976; Stoneham 1976). Until 1977, it was only well-established in isolated ponds in Kajjansi Fish Farm near Entebbe, where it had been introduced, without spreading to other major basins (Huner 1977; Huner & Avault 1978). Yet in 2006, the red swamp crayfish was present in the first place where it was introduced, near Entebbe close to Lake Victoria, but it was expanded to Lake Bunyonyi (SW Uganda) and also recorded downstream of River Kagera (Foster & Harper 2007). No updated information has been found about the species.
Zambia A legally authorized importation of 300 adults of the red swamp crayfish from Lake Naivasha, Kenya into a private experimental pool at Livingstone was made in the late 1970s (Grubb 1979). No updated information has been found about the species.
AMERICA Established (Brazil, Colombia, Costa Rica, Dominican Republic, Ecuador, Guatemala, Mexico, Puerto Rico, Venezuela and United States), probably established (Canada) and uncertain presence (Belize and Nicaragua).
Belize Although there are studies about the presence of the red swamp crayfish (Hobbs et al 1989; Huner & Barr 1991), neither full description in references nor available information was found about the species.
Brazil The first report of the red swamp crayfish dated back between ending 1970s and early 1980s (Huner 1986b). The red swamp crayfish began to be imported from US to be sold as a pet in the 1980s and their availability in shops only decreased after its ban in 2008 (Magalhães & Andrade 2014). Multiple releases led to the establishment of wild populations over several areas in Southeast Brazil near Sao Paolo (Magalhães et al. 2005). Currently, this species is established in some locations in the surroundings of São Paulo city, in São Paulo County (Loureiro et al. 2015).
53 Chapter I
Canada The presence of the red swamp crayfish was recently reported in 2017 near Vancouver (iNaturalist 2019). In 2018, news highlighted that more than 900 kg of live- crayfish coming from Maryland and Arkansas were seized in Michigan before crossing the Canada border (https://bit.ly/2FXTapW), which might present a high invasion risk for the country.
Colombia The red swamp crayfish was introduced for experimental aquaculture by a commercial enterprise in Cauca Valley in 1985 but its geographical origin remains unknown (Flórez-Brand & Espinosa-Beltrán 2011). Some specimens from the captive pool used in the experiments were accidentally released to wild in the basin of the Palmira river, Cauca Valley in 1988 (Arias-Pineda & Rodríguez 2012). Due to multiple secondary introductions after this escapement, the red swamp crayfish rapidly spread over the entire department of Cauca Valley (Flórez-Brand & Espinosa-Beltrán 2011), arriving to Cundinamarca region near Bogotá (Campos 2005). Currently, this species is established in Cundinamarca, Cauca Valley and Boyacá regions (Arias-Pineda & Pedroza-Martínez 2018; Pachón & Valderrama 2018).
Costa Rica The red swamp crayfish was introduced around 1966 in a small reservoir near Alajuela City (Centre Costa Rica) but its geographical origin remains unclear. From there, it appeared in the surroundings of San Carlos where it was successfully established (Huner 1977). Although Nannes’ letters informed that the red swamp crayfish had escaped from ponds to natural systems, no apparent problems were detected in that period (Nanne 1975). In 1994, it was already present in Cartago, Heredia, Alajuela, Guanacaste and Limón regions (Cabrera 1994). Currently, no additional information is updated (pers. comm. to F.J. Oficialdegui).
Dominican Republic The first reports of the red swamp crayfish dated back to 1977 when it was introduced by people of the US Peace Corps to be cultured under controlled conditions at the experimental stations at the Fisheries Experimental Station in Nigua (20 km south- western of Santo Domingo) and at the National Rice Experimental Station in Juma (80 km north of Santo Domingo) (Huner & Avault 1978). Currently, the red swamp crayfish is listed as one of the invasive species into the country and it is regularly captured in wetlands of Ozama River, surroundings of Santo Domingo and also northwards in Hatillo Dam, Cotuí (pers. comm. to F. J. Oficialdegui). 54 How the red swamp crayfish took over the world
Ecuador In 1986, the red swamp crayfish was introduced in the rice fields near Taura River (SW Ecuador) in the province of Guayas and subsequently, being expanded to the bordering region of Los Ríos in 1988 for aquaculture purposes in a similar way to rice fields in Louisiana (Salvador & Leyton 2000). Recently, in 2013, it was detected in Lake Yahuarcocha, province of Imbabura (North), where fishermen started to capture them (Riascos et al. 2018).
Guatemala Although its occurrence in the country is cited long ago (Hobbs 1989), no reports on its distribution have been found. Recently, the red swamp crayfish has been observed near Tecpán belonging to Department of Chimaltenango in South Central Guatemala (iNaturalist 2019).
Mexico Campos & Rodríguez-Almaráz (1992) detailed the distribution of the red swamp crayfish how native to North-Eastern Mexico, naturally inhabiting the basin of the Bravo River, but the species had also been widely introduced throughout the country (out of that basin). Although Re-Araujo (1994) cited its occurrence in the State of Baja California Norte since 1930s, albeit no confirmed. Other reports dated back from 1962, detected the red swamp crayfish in Conchos River near Camargo (Chihuahua) and near Cananea (Sonora) (Hobbs 1962). It was reported in 1968 south of Ensenada, Baja California (Clark & Ralston 1975). The range of the species expanded notably through Northern Mexico during the 1980s (Campos & Rodríguez-Almaráz 1992; Re-Araujo 1994). In early 21th century, new records were reported from Baja California Sur, Durango and Sinaloa, focusing on the expansion into new states and basins of the country (Hernández et al. 2008). The red swamp crayfish is currently widespread throughout Northern Mexico, but is also present to southwards in the State of Chiapas (Álvarez et al. 2011; Torres & Álvarez 2012; Franco-Sustaita 2014; Rodríguez-Almaraz & García-Madrigal 2014).
Nicaragua According to Huner (1977), only one specimen of the red swamp crayfish was found in Nicaragua, next to the Costa Rican border. The record seemed to be a result from a natural dispersion event from Costa Rica rather than an international importation (Huner 1977). Neither a full description nor available information has been described about the species.
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Puerto Rico Personal communications cited that the red swamp crayfish was being cultured in controlled laboratory systems in Puerto Rico (Huner & Avault 1978). The species was available from aquarium shops at least up to the 2000s, but, even though releases have occurred, established populations are not known in the wild (Williams et al. 2001). Since 2000s, the red swamp crayfish is enlisted into species prohibited from importation into Puerto Rico (Williams et al. 2001). No further information about its current distribution is known (pers. comm. to F.J. Oficialdegui).
Venezuela Approximately 1,200 specimens of the red swamp crayfish were shipped from Louisiana to Venezuela in 1978 with the aim of studying its suitability for culture in the country (Huner & Avault 1978). Years later the red swamp crayfish was captured in a pond of the Officers Club in Caracas and it is commonly available at pet shops in the city (Rodríguez & Suárez 2001).
United States States belonging to mostly the native area were not enlisted below: Arkansas (AR), Louisiana (LA) Mississippi (MS), and Texas (TX) or without accurate information about an introduction event or it is not present in the States: Iowa (IA), Montana (MT), New Hampshire (NH), North Dakota (ND), South Dakota (SD), Vermont (VT), West Virginia (WV) and Wyoming (WY).
Alabama (AL) The red swamp crayfish is native to the Southwest of State (Hobbs 1989), but in 1961, it was introduced at Auburn University Aquaculture Station and they would have been well established on the area ten years later (Huner & Avault 1978). This species is now known to be present in the Tennessee, Mobile, Black Warrior, Cahaba, Coosa, Tallapoosa and Escambia river systems (Schuster & Taylor 2004; Shelton-Nix 2017).
Alaska (AK) One specimen of the red swamp crayfish was found in the city of Kenai in May of 2004, who might come from release of a private aquarium (Tunseth 2004). However, no population is known to be established in wild yet (Nagy et al. 2019).
Arizona (AZ) First evidences of the red swamp crayfish was found into stomach of striped bass in Lower Colorado Basin, close to the Californian border in 1969 (Edwards 1974). In 56 How the red swamp crayfish took over the world
1989, it was present along Verde River within Tonto National Forest (USGS 2019) and currently, it is found in Lower Colorado Basin, Salt River, a tributary of Hassayampa River, canals of Phoenix and San Pedro River (Marsh 1999; Moody & Taylor 2012).
California (CA) The red swamp crayfish was present into a stream near Pasadena, Los Angeles County, in 1924 (Holmes 1924). However, its first introduction must have taken place earlier because there was a translocation of live-specimens of the red swamp crayfish from California to Oahu, Hawaii in 1923 (see Brasher et al. 2006). Few years later, in 1932, it was introduced as frog food in a farm of Lakeside, San Diego County (Riegel 1959). The species was also present in Santa Rosa region before 1940s, which might be the original population that colonized Las Vegas River few years later (Hobbs & Zinn 1948), and in Santa Barbara (Penn 1954). During 1950s, the red swamp crayfish was well established in southern and central California, being the only freshwater crayfish found in south of Tehachapis Mountains (Riegel 1959) and it was regularly taken in the Sacramento-San Joaquin Delta during 1960s. By 1970s, the red swamp crayfish was causing crop damage and levee destruction in rice fields of the Sacramento River delta, where the species was being exploited at small scale (Huner 1988). By then, it was well established in San Francisco Bay (Ruiz et al. 2000) and collected from Sweetwater River in San Diego National Wildlife Refuge (Cohen & Carlton 1995). However, there were areas such as Topanga Creek near Los Angeles that, though was surrounded by the red swamp crayfish decades before, remained no infested until 2001 (RCDSMM unpublished data in Garcia et al. 2015). Attempts to eradicate were carried out by active removal efforts, but low flows and below average rainfall in 2011-2014 facilitated anew its extensive establishment (Garcia et al. 2015). Currently, it is widely distributed across the entire State (Nagy et al. 2019).
Colorado (CO) In 2018, one specimen of the red swamp crayfish has been observed into a city lake in Denver (iNaturalist 2019). No further information about its introduction has been found.
Connecticut (CT) In 2017, the red swamp crayfish has been detected in Indiantown Brook, a tributary to Thames River near Norwich (iNaturalist 2019). No further information about its introduction has been found.
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Delaware (DE) In 2002, some specimens of the red swamp crayfish were being reared for culture research at Department of Agriculture and Natural Resources (Gherardi & Daniels 2004). However, there were not wild reports until some recently established populations in Brandywine Creek (North) and Broad Creek (South) (Nagy et al. 2019), probably by its proximity to Maryland and Pennsylvania States where was established before.
District of Columbia (DC) Some occurrences has been reported recently in the northern half of the state (iNaturalist 2019), however, little is known about its introduction.
Florida (FL) The red swamp crayfish is native to the North-western Florida (Hobbs 1989), however, Penn (1954) reported an introduction of 700-900 crayfish at a private crayfish farm near Hudson, Pasco County, in 1951, but this introduction seemed to have been unsuccessful (Rhoades 1976; Huner 1977). In the late 1970s, approximately 5,000-6,000 adult the red swamp crayfish brought from Louisiana were introduced into cultured ponds near West Palm Beach (Huner & Avault 1978). Currently, it is widely distributed across the State with established populations in Lake Rousseau, ponds near Tampa, near Orlando, Wakulla Springs, Lake Alice in Gainesville and Guana River near San Agustín (iNaturalist 2019).
Georgia (GA) The red swamp crayfish was introduced at some time after Hobbs’ study (1981) because it was not included there but its presence downtown Athens has long been known (around 1989), and it has recently spread to Oconee River and its tributaries (Nagy et al. 2019). One single specimen appeared in Gwinnett County in 2007. By 2008, it was collected from Etowah River (Skelton 2010). Currently, it is distributed in the northern areas, mainly in the surrounding area of Atlanta but also near Columbus and Augusta (iNaturalist 2019).
Hawaii (HI) The first introduction of the red swamp crayfish remains unclear because it was introduced from California to Oahu Island, Hawaii in 1923 or 1927 (see Brasher et al. 2006; see Brock 1960), originally to serve as food for bullfrog breeding facilities (Huner 1977). In 1934, a new batch of 400 specimens of the red swamp crayfish was carried from Santa Barbara, California, to a frog farm in Oahu Island, Hawaii (Penn 1954). Few years later, the red swamp crayfish was introduced from Oahu to the island of Hawaii and 58 How the red swamp crayfish took over the world
Maui in 1937 and 1939 (Brock 1960). By 1954, red swamp crayfish was widely distributed on the islands of Kauai and Maui and it was established on Molokai by 1977, where it was rapidly considered a pest. Currently, it is widely distributed in Oahu, Kauai (pers. comm. to F.J. Oficialdegui) and West Maui Forest Reserves (iNaturalist 2019).
Idaho (ID) The red swamp crayfish was first detected in 1975, when several specimens were collected near Nampa, Canyon County (Clark & Wroten 1978). Currently, it appears in Snake River around Morley Nelson Snake River Birds of Prey National Conservation Area (Department of Fish and Game, Idaho) and some of its tributaries such as Salmon, Clearwater and Selway Rivers (pers. comm. to F.J. Oficialdegui).
Illinois (IL) The red swamp crayfish is native from Southern Illinois (Pope, Johnson, Massac, Union, Pulaski, and Alexander counties) between Mississippi and Ohio basins (Hobbs 1989). However, its invasive range in the State has been artificially expanded. In 2001, it was collected from Chicago River (Taylor & Tucker 2005) and from Dead River near Lake Michigan in 2004, subsequently, it has spread to surroundings of Chicago in last few years (i.e.: DuPage county in 2010 and McHenry County in 2017) (Nagy et al. 2019).
Indiana (IN) Eberly (1954) did not include the red swamp crayfish on his list of the distribution of Indiana crayfish, but studies developed in the 1980s already cited this species within the state (Huner 1986; Hobbs 1989). The red swamp crayfish was considered one of the rarest species over this state, it restricted to extreme South-western Indiana, streams in Posey, Vanderburgh and Warrick counties (Page & Motessi 1995). However, in 2000, red swamp crayfish could be collected from Lake Michigan (North-Western of the state). The ongoing market and pet trade of this species was surely the reason by which this crayfish was spreading into the West Branch of the Grand Calumet River, border Illinois and Indiana (Simon 2001). No updated information has been found about the species.
Kansas (KS) In 2017, one specimen of the red swamp crayfish was detected near Kansas City (iNaturalist 2019). No further information about its introduction has been found.
Kentucky (KY) The red swamp crayfish is native to the Southwestern of the State (Hobbs 1989). This species was supposedly introduced few years before 1944 because Rhoades (1944) 59 Chapter I enlisted the species, being considered as a new entry. Currently, it has only observed in the southwest of the State near Mississippi River where it is native from (iNaturalist 2019).
Maine (ME) In Martin’s study (1997), the red swamp crayfish was not enlisted as species introduced in the State, however, it is found in Kennebec river system since 1980 (Nagy et al. 2019).
Maryland (MD) The red swamp crayfish was introduced in 1963 from Louisiana at Patuxent Wildlife Research Area (20km northeast of Washington D.C.), to serve as food for wading birds (Kilian et al. 2009). In 1981, a new batch purchased in Louisiana was carried to Pocomoke and Nanticoke rivers to try crayfish culture. From that original stock, more introductions occurred legally into other basins until 1990. Since then, other translocations might have occurred but there are no confirmed evidences, due to the establishment of aquaculture permit regulations. In 2006, it was well established in Chesapeake Bay, Delmarva Peninsula and all 14 watersheds of the Coastal Plain of Maryland (Kilian et al. 2009). These last occurrences may have resulted from introductions by anglers (Kilian et al. 2010). Currently, it is still established in the same areas (iNaturalist 2019).
Massachusetts (MA) In 2010, the red swamp crayfish was detected near University of Massachusetts in Amherst (iNaturalist 2019). In 2012, it was collected in Salisbury pond, Worcester, in order to study changes in water quality (Davis 2013).
Michigan (MI) First specimens of the red swamp crayfish were found in south-eastern shore of Lake Michigan near Holland in 2013, and four years later, a few established populations have been found eastwards, near Kalamazoo and Oakland county (Nagy et al. 2019). In addition, the red swamp crayfish is currently established in 30 small ponds near Novi and also, in Sunset Lake near Vicksburg (pers. comm. Michigan Department of Natural Resources) and also observed near Gaylord, northern of State (iNaturalist 2019).
Minnesota (MN) Only two specimens of the red swamp crayfish have been collected from Tilde Lake in 2016 (Minnesota Department of Natural Resources). 60 How the red swamp crayfish took over the world
Missouri (MO) One specimen of the red swamp crayfish was collected in a survey carried out in Table Rock Reservoir in 2009 (DiStefano et al. 2015). Recently, it has been observed in St. Louis and surroundings (iNaturalist 2019).
Nebraska (NE) In 2014, the red swamp crayfish was found in a bait dealer’s tank located on the Missouri River. Currently, it is distributed in the Missouri River downstream of Gavins Point Dam and in Lake Yankton (Schainost 2016).
Nevada (NV) Hobbs & Zinn (1948) collected several specimens of the red swamp crayfish in Las Vegas River in the fall of 1944. By the mid-1950s, the species was well established in southern Nevada (Penn 1954). Two decades later (Rhoades 1976), it was still thriving, even though there were state law prohibits selling and transporting them, probably responding to damages of irrigation systems in this very arid region (Huner 1977). The decline of the bass fishery (Micropterus salmoides) since the late-1970s led to an interest in crayfish stocking in Lake Mead as food source for fishes, which took place in 1988 with four releases, each one involving 1600 crayfish (Hager 1990). During 1986-1987, an exhaustive crayfish survey was carried out and low densities of the red swamp crayfish were found (Leavitt et al. 1989). Currently, it is also distributed in Ash Meadows across lower Colorado Basin, Southwest (Paulson & Martin 2014).
New Jersey (NJ) Some occurrences has been reported recently in the northern half of the state (iNaturalist 2019; Nagy et al. 2019), however, little is known about its introduction.
New Mexico (NM) One specimen was collected in Grande River in northern New Mexico in 1944 (Nagy et al. 2019). Recently, some populations of the red swamp crayfish have been detected across the Grande River from Albuquerque to El Paso (iNaturalist 2019)
New York State (NY) It is now established in New York State, at least from 2002, when the red swamp crayfish was found in Long Island and lower Hudson River system (Nagy et al. 2019).
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North Carolina (NC) First introduction events seem to have occurred in the beginning of 1980s but without information on localities and dates (Huner & Barr 1983; Huner 1986a). However, it can confirm because the red swamp crayfish was already sold on seafood markets in Raleigh in 1985 (Nagy et al. 2019). Cooper et al. (1998) reported wild populations of the red swamp crayfish in the Neuse, Tar-Pamplico, Yadkin-Pee Dee, and Cape Fear River basins, which might have originated from accidental releases from aquaculture facilities or aquarists. Besides of new locations in already invaded basins, Fullerton & Watson (2001) also reported the red swamp crayfish in Broad, Pasquotank, Waccamaw river basins. By 2007, it was present over almost all territory of the State (Cooper & Armstrong 2007).
Ohio (OH) First report of the red swamp crayfish dated back to 1967 in Sandusky Bay, after this first introduction, the red swamp crayfish was subsequently coming up on its tributaries in Sandusky County (Norrocky 1983). They started to be collected from several State Fish Hatcheries from various counties (e.g. Erie, Sandusky, Ottawa and Madison) in 1982 (Norrocky 1983). The heterogeneous distribution in Ohio (Northern of the State in Sandusky Bay, Northeastern near Cleveland and Centre near Colombus) suggested that it had not dispersed naturally from its native range (Norrocky 1983). Currently, it is still established in the same dispersed areas previously colonized, reaching the surroundings of Cincinnati in 2014 (iNaturalist 2019).
Oklahoma (OK) The red swamp crayfish is native to the south-eastern corner of the State (Hobbs 1989) but it is out of its native area in McCurtain County since 1969 (Reimer 1969) and after in Okfuskee County (Jones et al. 2005). Recently, it has been observed near Tulsa and in Veterans Lake near Sulphur in the South of Oklahoma (iNaturalist 2019).
Oregon (OR) By ending of 1990s, the red swamp crayfish was established in ponds and streams throughout the Willamette Valley (Pearl et al. 2005). A recently exhaustive sampling showed the widespread distribution of the red swamp crayfish in Western part of the State (Pearl et al. 2013; Nagy et al. 2019; iNaturalist 2019)
62 How the red swamp crayfish took over the world
Pennsylvania (PA) First report dated back to 1990 in the Schuylkill River (SE of Pennsylvania), after that, a few of established populations have been found near Philadelphia (Lieb et al. 2011). Currently, it is mainly established in the eastern of the State (iNaturalist 2019).
Rhode Island (RI) A female of the red swamp crayfish was found in a pond in northern Arcadia, Washington Co. in 1970 (Crocker 1979). However, no occurrence of the red swamp crayfish in inland waters of the State have been found on ongoing surveys (pers. comm. to F. J. Oficialdegui).
South Carolina (SC) During the 1970s, several aquaculture enterprises began to be developed in this state carrying on the red swamp crayfish from Louisiana (Huner & Avault 1978). Because of the existence of several production sites of crayfish over the state, the species is widely distributed into the State (Eversole & Jones 2004; iNaturalist 2019).
Tennessee (TN) This species is native from the Mississippi basin in western of this state (Hobbs 1989) but it has been detected currently in J. Percy Priest Lake near Nashville and near Manchester (iNaturalist 2019).
Utah (UT) First reports of the red swamp crayfish dated back to 1978 when some specimens were collected near St. John in Rush Valley, Tooele County. In following surveys of 1983, the species was still present in the area (Johnson 1986). Currently, it has been observed in Jordan River through Salt City (iNaturalist 2019).
Virginia (VA) One specimen of the red swamp crayfish was collected in 1972 in the York- Pamunkey drainage and another in the Potomac watershed in 1992 (US National Museum of Natural History 2011). Currently, it is widely expanded across the State, for example, it was found westards to eastwards, in ponds of Blacksburg, Sweet Briar Lake, James River in Wingina, Briery Creek Lake, Broad Branch, Maury Lake and False Cape State Park (iNaturalist 2019).
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Washington State (WA) Three live-specimens of the red swamp crayfish were captured in fall 2000 during a routine survey in Pine Lake in the Pacific Northwest State of Washington (Mueller 2001; Larson 2007). Between 2007 and 2009, the red swamp crayfish was collected from 11 lakes in the Puget Sound lowlands (Larson & Olden 2013).
Wisconsin (WI) An established the red swamp crayfish population occupied a private subdivision pond in Sam Poerio Park in Kenosha County, but it was eradicated in 2009 (Behm 2009). In the same year, other populations appeared in Washington County (Wisconsin Department of Natural Resources, https://dnr.wi.gov/topic/Invasives/fact/RedSwampCF2012.html).
ASIA Established (China, Japan, South Korea, Taiwan and Thailand), probably established (Hong Kong and Indonesia) and eradicated (Israel).
China In 1930, Japanese transported the red swamp crayfish from Japan to a garden in Nanjing, Jiangsu region, without clear reasons (Cheung 2010). Apparently, Japanese civilians brought and reared them as pet during the second Sino-Japanese war (1937- 1945) and released them in wild before going back to Japan at the end of the war (Cheung 2010). However, local Chinese people considered its introduction as a Japanese conspiracy to destroy rice fields (Xinya 1988). Because of its Japanese connotations, Jiangsu people did not like and, by consequence, did not eat the red swamp crayfish until 1980s. For this reason, its Chinese distribution was closely located around Nanjing but it was quickly spread across Eastern China in 1980s (Xinya 1988), until reaching Hong Kong in the Southeast (Hobbs et al. 1989). Nowadays there is a big business around crayfish harvesting and commercial use (Cheung 2010). Currently, the red swamp crayfish is widely distributed in more than 20 China’s provinces, being widely distributed in the middle and lower reaches of Yangtze River, where concentrate the main production areas of the red swamp crayfish (Gong et al. 2012) but also, there are established population going northwards and southwards from Yangtze River.
Indonesia Indonesia is considered as one of the most suppliers of ornamental crayfish, the red swamp crayfish among them (Patoka et al. 2015). Although the import of the red swamp crayfish is banned since 2014, its culture and transport are legal within the country. 64 How the red swamp crayfish took over the world
When the red swamp crayfish was introduced remains unknown. Nowadays, the red swamp crayfish is present in pet shops of the country and few wild populations have been found in wild in Java Island, Cisaat Subdistrict (Putra et al. 2018) and Halimun montain (pers. comm.).
Hong Kong The red swamp crayfish is cited in the city as pet and it is a likely source of the Taiwanese populations (Hobbs et al. 1989). Currently, it has been observed in Hong Kong Island (iNaturalist 2019).
Israel In 2008, the red swamp crayfish was fortuitously found in a temporary pond near Hadera, 40 km northward from Tel-Aviv (Wizen et al. 2008) and attempts, allegedly successful, to eradicate them were carried out by Israel Nature and Parks Authority (INPA). The provenance of this population is still unknown. Currently, no presence of the red swamp crayfish is detected over this country (pers. comm. to F. J. Oficialdegui).
Japan Of an initial uptake of one hundred red swamp crayfish, only twenty survived and were introduced into Japan in late 1920s (1927 or 1930, the precise date is not clear in literature) to serve as food for the American bullfrog, Rana catesbeiana (Penn 1954; Huner 1977; Kawai 1999; Cheung 2010). They were the survivors of an original shipping of 100 individuals from New Orleans (US), which were transported in beer barrels by the manager of a bullfrog farm in Kanagawa prefecture, Honshū Island. The species spread over Honshū Island due to its use as family pets (Sako 1987; Kawai 2017). The absence of episymbiont branchiobdellidan worms living on Japanese the red swamp crayfish, suggest is thought to be related to the deficient transport conditions of the first introduction event and suggests that all the red swamp crayfish currently found over Japanese Archipelago are descendants of the 20 specimens released in Kanagawa in 1927 (Kawai & Kobayashi 2005; Kawai 2017). The species spread rapidly throughout the country and, by the 1950s, it was very abundant and caused agricultural damages on rice fields (Penn 1954). By 1975, the red swamp crayfish was already present in all Japanese Prefectures, with the exception of Hokkaido (Takeda 1975) and by the 1990s it occupied the whole country (Kawai 1999). The rapid spread of the red swamp crayfish from a single introduction point observed in Hokkaido and other Japanese islands suggests that unreported or illegal introductions have occurred across the Japanese archipelago (Kawai 2017). Currently, it has been also found in Okinawa Island (iNaturalist 2019). 65 Chapter I
South Korea Although little is known about the first introduction and distribution of the red swamp crayfish, some specimens were bought for research studies in a fish market of Incheon around 2005 (Ahn et al. 2006). Recently, it has been observed in South Jeolla Province in the southwest of South Korea (iNaturalist 2019).
Taiwan First reports of the presence of the red swamp crayfish come from the 1960s (Chen et al. 2003). Aquaculture and pet trade were the main pathways of introduction in the island (Hobbs et al. 1989; Gao & Hong 2001) and subsequently, escapes and releases directly to wild caused its spread across the island. Currently, populations of the red swamp crayfish are widely established around Taipei, North Taiwan (iNaturalist 2019).
Thailand First introductions of the red swamp crayfish dated back to 1987 coming from US when this species started to be commercialized for aquaculture purposes in Chiang Mai province, northern of Thailand (Vidthayanon 2005; pers. comm. to F. J. Oficialdegui). Currently, it is located in wild (River Kwai, western Thailand) and, it is relatively easy to find it on websites of pet shops in many cities.
EUROPE Established (Austria, Belgium, Cyprus, England, France, Germany, Italy, Malta, Portugal, Spain, Switzerland and the Netherlands) and uncertain wild presence (Luxembourg, Slovakia and Sweden).
Austria In the 1990s, no the red swamp crayfish was found in the wild, but it was on sale in aquarist shops as "Red Lobster" (Pöckl 1999). Due to accidental releases from aquaria, there were at least two sightings near Salzburg where might become established (Strasser & Patzner 2005). A recent review shows its presence in this country (Holdich et al. 2009; Kouba et al. 2014) but no updated information has been found about the species.
Belgium The first specimen of the red swamp crayfish was found dead in the reservoir of Vielsalm during the first Belgian survey during the years 1983-1985 (see in Boets et al. 2009). This specimen might have originated from a nearby restaurant. This finding was allegedly casual because the red swamp crayfish was not considered present in Belgium 66 How the red swamp crayfish took over the world until more than ten years later (Arrignon et al. 1999). In 1996, a living individual in a pond nearby Cerfontaine was found during a large scale distribution survey of crayfish in Wallonia. In 2008, the red swamp crayfish was found in the nature reserve Zammelsbroek in Zammel. In 2009, populations of the red swamp crayfish were found there in three ponds situated northeast of the nearby River Grote Nete. The scattered distribution of the red swamp crayfish in Belgium suggests that the species probably escaped from nearby private ponds or was deliberately released by amateurs keeping crayfish as a hobby (see in Figure 5, Boets et al. 2009).
Cyprus It had allegedly been introduced to Cyprus in 1980s, where it flourished in the Athalassa dam and subsequently, it was introduced into two other dams (Stephanou 1987). Currently, it is still present in the Cyprus Island (Kouba et al. 2014) and newly observed in Athalassa National Forest Park (iNaturalist 2019).
England Specimens of the red swamp crayfish was imported for educational or recreational purposes as pets in domestic aquaria and culture trials several times during 1980s (Dawes 1981), subsequently released into aquatic ecosystems (Goddard & Hogger 1986; Hobbs et al. 1989). These introductions came from Kenya into wholesale fish markets (Goddard & Hogger 1986; Oficialdegui et al. 2019). Subsequently, the red swamp crayfish was present at low levels in a roadside ditch in Tilbury and River Lee in 1990 (Ellis, unpublished data). However, a high density of the red swamp crayfish was recorded in Britain in 1991 in the Men's Bathing Pond at Hampstead Heath in North London (Richter 2000) and at two separate locations in Kent during 1994 (Foster 1996). There were some occurrences in Regents Canal, London, in 2000 (Richter 2000) and later, a suspicious population in a small fishing lake near Windsor in May 2012. During surveys carried out between 2008 and 2010, the species was found in four other ponds within Hampstead Heath Park. The current distribution range of this species remains small in England (Ellis et al. 2012).
France The red swamp crayfish was introduced in 1974 in a fish farm near the Charente River in Western France (Arrignon et al. 1999) and later, unconfirmed reports indicated that attempts to import red swamp crayfish in France was carried out (Huner & Avault 1978). Changeux (2003) performed a national survey on the presence of native and invasive crayfish since 1977, aiming at documenting crayfish distribution in the country in France. Before 1990s, there were many scattered occurrences of the red swamp 67 Chapter I crayfish in France. For example, by 1978, it was found on the dam of Rouvière, in a tributary of Le Vidourle River between Hérault and Gard regions, where posteriorly its fishing was allowed (Laurent et al. 1991). in the swamps of Brière as consequence of one crayfish that escaped from a private crayfish farm in 1981 and posterior expansion by fishermen who sold it in live markets (Arrignon et al. 1999), in region of Pays de la Loire (Loire-Atlantique since 1984 and Mayenne since 1985-1987). Thus, the red swamp crayfish was already present over 10 departments from 7 regions by 1990, mainly Western France. Five years later, in 1995, its distribution range had reached 33 departments from 12 regions, particularly marshy and rice area in Brittany, Atlantic and Mediterranean watersheds/seaboards including the Camargue (Rosecchi et al. 1997; Arrignon et al. 1999; Changeux 2003). Later, there were other occurrences in different departments until reaching up to 49 departments from 16 regions in 2001 (see Figure 6 in Changeux 2003) and it was present in 61 departments by 2006 (see Figure 7 in Collas et al. 2007), reaching high population densities in Southwestern France. It was found in the Vosges department in 2008, along the upper part of the Meurthe River, northeast France (Collas et al. 2008). In the Ardennes Department, adjacent department to Belgium, it has been present in several rivers, including the Chiers River, which are tributaries of Belgian rivers (CETE de l'EST 2011). Its quick spreading was probably caused by large illegal translocations from surrounding countries (Laurent 1995b) and not only, there were importations from further countries like Kenya since 1976 (Laurent 1990). Approximately, 170 tons were imported to France from Kenya where they were sold until France imposed import ban of live crayfish (Laurent et al. 1991). As an example, the estimated exportation of crayfish companies in rice fields of Seville was around 300 tons, of which 78.1% was sent to French markets in 1990s (Gutiérrez-Yurrita et al. 1999). In fact, the red swamp crayfish was widely expanded in France in 1995 and researchers looked for a biological control to decrease the extensive populations in rivers around Paris (Laurent 1995a). At the beginning of this century, Poitou-Charentes and Aquitaine regions (Southwest France) produced annually more than 200 tons of the red swamp crayfish (Changeux 2003). Currently, the red swamp crayfish is widely distributed over West and South of France but also there are established populations in Centre and North of France (Kouba et al. 2014; GBIF 2019; iNaturalist 2019).
Germany The red swamp crayfish may have been located near Ulm since 1975-1976, this presumption is based on local fishermen who asserted catching an exotic crayfish species (Chucholl 2011). In 1993, it was discovered in 16 localities of North Rhine- Westphalia (Löbf 1995; Groß et al. 2008). The commercial success in other bordering countries could have been the trigger of introductions over this area. The discontinue 68 How the red swamp crayfish took over the world distribution pattern of this species in the area could have been consequence of translocations by men (Lake Hechtsee and Lake Riedheim) and subsequent active spread to surrounding habitats. By 2011, the estimated population was of approx. 13,400 crayfish in Lake Riedheim but any commercial activity was still developed (Chucholl 2011).
Italy Although the red swamp crayfish was being experimentally reared in a farm from 1977-1985, the first report of red swamp crayfish in wild was in Banna River, within Po Basin in Piemonte, where the red swamp crayfish appeared in 1989 after escaping from the installation (Delmastro 1992). By posterior samplings over the area during the following years, the occurrence of the red swamp crayfish was confirmed, even the species had spread to all Piedmont province (Delmastro 1999). Posteriorly, juvenile crayfish were collected during the sampling season in Lake Massaciuccoli, Tuscany, Italy, in 1994. Their provenance seemed to be cultured animals in a crayfish farm that, after its bankrupt in 1993, there were fortuitous releases to wild (Baldaccini 1995). Few years later, this species was especially abundant in this Lake and surrounding areas (Gherardi et al. 1999). In fact, this lake may have been the origin of future introductions in other regions of Italy (Aquiloni et al. 2010). Moreover, Barbaresi et al. (2007) hypothesized that foreign introduction may have happened, concretely, one population of Florence could have come from China following the immigration of a Chinese community to Florence. Others occurrences also appeared in Reno River drainage area, Emilia-Romagna, since 1995 (Mazzoni et al. 1996). Due to translocations by man, it started to appear in the many other regions of north-central Italy. This crayfish was found in Iseo Lake in 1991 (Delmastro 1992), Garda Lake (I. Confortini, pers. comm. to Aquiloni et al. 2010) which is placed between the provinces of Lombardia and Veneto (P. Turin, pers. comm. to Aquiloni et al. 2010). By 1994, it was present near Verona, province of Veneto (Morpugo et al. 2010), reaching Seriola Channel, between Padua and Venice, on September, 7th of 2002. (Mizzan & Vianello 2007). However, the distribution of the red swamp crayfish in Lombardy was mainly located to southwestern of the region between 1994 and 2006 (e.g.: provinces of Pavia, Milano and Lodi) (see Figure 1 in Fea et al. 2006). This species was also present in Region of Liguria (Gherardi et al. 1999) and by 2009, some specimens appeared in Tagliamento, Meduna, Torre river basins and mouth of Isonzo river in Friuli Venezia Giulia (De Luise, 2010). Moreover, it was found in central provinces like Umbria (Dörr et al., 2001), the Marches, Abruzzo (Gherardi et al. 1999) and Latium (Chiesa et al. 2006)) where it was well established (see Table 1 and references in Gherardi et al. 1999; Barbaresi & Gherardi 2000). In Lake Trasimeno, Umbria, the red swamp crayfish seems to have been introduced since 1985 69 Chapter I and captured by fishermen to sell in local markets (Döor et al. 2001). In addition, it was found in several lakes of this province, Lake Piediluco, and the neighbouring province of Rieti in Lake Ventina, being well established in surroundings areas of lakes and streams in the early 2000s (Döor et al. 2001). Regarding on Southern of Italic Peninsula, the occurrence of this species was not reported until recently (see Table S1 in Cilenti et al. 2017). Concretely, this species appeared in Bradano River and San Giuliano Lake in Basilicata (Caricato et al. 2013), in Campania region was detected near Napoli where control and eradication efforts were made (Stinca 2013), present in Tarsia Lake and in middle course of Crati River in Calabria by 2012 (Sperone et al. 2015). In Puglia, first report of this species dated back to a sampling on Lesina Lake in 2007, it might have colonized the lake for accidental release from aquaculture activities (Florio et al. 2008). Recently, juvenile of the red swamp crayfish were repeatedly observed in an artificial drainage ditch in Melissano (Lecce), located in the southwestern area of the Salento Peninsula in 2016 (Cilenti et al. 2017). In Sicily, the first finding of the red swamp crayfish was in June 2002, where a specimen and several crayfish exoskeleton were found in The Nature Reserve of Priola and Gorghi Tondi Lakes in western of the island (D’Angelo & Lo Valvo 2003; Dörr et al. 2006). Since there were no farming activities of this species in the nearby areas, the hypothesis was a release of some specimens by people keeping crayfish for recreational purposes. Moreover, this may be confirmed by the observation in the same period of one individual of the turtle Red-eared Slider, Trachemys scripta, another exotic species usually kept in captivity (D’Angelo & Lo Valvo 2003). Since then, the red swamp crayfish appeared in far locations of the Island; for example, it was found in Rosamarina Lake during 2012-2013 (Di Leo et al. 2014). The present situation over the Island is detailed by Faraone et al. (2017) using a citizen science approach. In this study, they found this species in different river basins or lakes, separated by hundred kilometres from previous invaded areas, suggesting multiple independent releases in wild (see Table 1 for more details). On the other hand, in Sardinia, first specimens were found in the northern part of the island (Gallura area) between 2000 and 2002, after then, more specimens were consecutively recorded to west- and southward expansion (Coghinas River and watercourses along Tyrrenian coast). Since 2010, it is also present in southern part of the island, near to Cagliari (see regional reports within Cilenti et al. 2017). In Italy, the presence of nonindigenous species in the wild has always been related to aquaculture activities, in fact, the distribution of farms in late 1990s mostly overlapped with the occurrence of new reports of invasive crayfish (see Fig. 2 and Fig. 3 in Gherardi et al. 1999).
70 How the red swamp crayfish took over the world
Luxembourg The red swamp crayfish was no present in Luxembourg before 2009 (Arrignon et al. 1999; Holdich et al. 2009; Kouba et al. 2014). However, as some populations have established themselves in neighbouring countries, it may be possible that the species was already present but not yet detected in some ponds or rivers (Delsinne et al. 2013).
Malta The first specimens of the red swamp crayfish were captured at Fiddien Valley, Western of Malta, in September, 2016 (Vella et al. 2017). Genetically-based, the origin of this invasive population might be attributed to China because of the commercial route that both countries recently agreed to establish each other (Oficialdegui et al. 2019).
Portugal Although the spreading of this species has been related to the attempt to create commercially exploitable breeding populations (Gutiérrez-Yurrita et al. 1999), the first introduction in Portugal was probably the result of the natural dispersion from a naturalized population near Badajoz in 1979 (Ramos & Pereira 1981). This species was subsequently colonizing new basins northwards and southwards of Portugal from the initial introduction point in Caia River in 1979. For example, first occurrences of the red swamp crayfish were found in Tagus (1986), Guadiana (1986), Mondego (1987), Sado (1990) and Douro basin (before 1993) (for more details, see references in Gutiérrez- Yurrita et al. 1999) being established in Maças River, North-east of Portugal, by the ending 1990s (Bernardo et al. 2011) and Minho region in the beginning of 2000s (Moreira et al. 2015). This expansion did that the annual production was around 700 tons in Portugal during 80s (Gutiérrez-Yurrita et al. 1999) but this market may have grown up to 3,000 tons in the ending 90s. Such was the case that some companies (Dutch company) was valuing to move a crayfish processing factory to the Mondego Valley and export crayfish to North European markets (see in Gutiérrez-Yurrita et al. 1999). In addition, new populations of the red swamp crayfish were not limited to Iberian Peninsula but also, the red swamp crayfish arrived to São Miguel, Azores, Portugal by human mediated, it is believed that some crayfish were released from aquaria into Peixe Lagoon (Correia & Costa 1994).
Slovakia Although there were no wild populations, Stloukal and Vitázková (2009) mentioned the fact that the red swamp crayfish occurs in garden ponds and might present a high invasion risk in Slovakia.
71 Chapter I
Spain The first importation was made on 13th and 14th June 1973 in two shipments from New Orleans, Louisiana. That introduction was sponsored by private interests with the approval of the Ministry of Agriculture (Huner 1977). A total of 250 females and 240 males were sent by plane in plastic bags, filled with 1/3 water and 2/3 oxygen. Crayfish were housed in four isolated earth ponds near rice fields and a fence was installed to avoid releases to any other water body. However, some crayfish were found on surrounding irrigation channels where they were easily caught. Although half of the first introduction crayfish died during the first year, the survival ones were success and in summer 1974, some around 500 crayfish were recaptured and they found females with up to 400 juveniles (Habsburgo-Lorena 1978). On May 10, 1974, another importation of 500Kg was made into the province of Seville, near Alfonso XIII, Isla Mayor, a village in the middle of the rice growing area. Crayfish were released into abandoned ponds, indeed, crayfish importations attempted to substitute an important eel business, which existed before. That time, there were important and considerable losses, indeed, it was estimated that only a 20% survived; however, once seen that the rest of crayfish were a great success, fishermen disseminated crayfish in nearby rivers and irrigation channels (Habsburgo-Lorena 1978). Since the moment when it became clear that this species had adapted, it was obvious that money was to be made on it and it was posteriorly expanded by fishermen throughout Spain, France and Italy during the period from the late 1970s into the late 1980s (Gutiérrez- Yurrita et al. 1999; Henttonen & Huner 1999). By 1976, there was a commercially exploitable production in the rice fields. In fact, captures increased from 480 crayfish in 1975, 1,843 kgs. in 1976, 9,800 kgs. in 1977 up to 13,119 kgs. by the end of May 1978 (Habsburgo-Lorena 1978) and . Few years later, the annual the red swamp crayfish production was around 2,900 tons in Spain, reaching the maximum Spanish production of 5,000 tons during 1987 (Gutiérrez-Yurrita et al. 1999). In 1976, the red swamp crayfish was already present in Doñana National Park, concretely in the Rocina stream (north of the Park) and Las Nuevas Canal (near mouth of Guadalquivir river) (Algarín 1980) and by 1977, the Rocina had become a fishing site. In 1979, the red swamp crayfish had arrived to rivers and streams of bordering provinces (Cádiz and Huelva). The economic benefit of this species encouraged different owners of rice fields to transfer to other regions of the country, in this way, it was introduced into Tablas de Daimiel, Ciudad Real, before 1978, the rice fields of Valencia in 1978 and Ebro Delta in 1979. It also appeared in the province of Zamora in 1979 (Alonso et al. 2000). By ending 1980s, it allegedly appeared in Minho River (Sousa et al. 2013). It was introduced into Cuenca in 1986, occupying mostly entire province by 2000 and almost region of Castilla La Mancha by 2006 (see Fig. 2 and Fig. 4 in Alonso & Martínez 2011; respectively). Due to crayfish 72 How the red swamp crayfish took over the world plague, there was a vertiginous decrease of captures of Austrapotamobius italicus in all regions between 1974 and 1980 in Ebro basin (where European crayfish is mostly distributed) (Fernández 2004). It was attributed to the introduction of the red swamp crayfish and not by occurrence of signal crayfish few years later (Alonso & Martínez 2011; pers. comm. to F.J. Oficialdegui). In addition, the spreading of the red swamp crayfish was not limited to Iberian Peninsula, indeed, red swamp crayfish arrived (by human-dispersal) to The Balearic Islands in 1993 and to San Andrés steep river bank, Tenerife Island, Canary Islands in 1997 (see reference in Gutiérrez-Yurrita et al. 1999). In fact, Alonso et al. (2000) mentioned that there already were dense population of the red swamp crayfish in the most of provinces in Spain (mainly the south half of Spain), excepting Lugo in Galicia. During 90s, there were multiples translocations of crayfish between Portugal and Spain to furnish crayfish food companies and the red swamp crayfish market mainly exported frozen animals from rice fields in Seville to Valladolid and Madrid (Gutiérrez-Yurrita et al. 1999) and therefore, it would not be rare that there have been translocations of live animals for culture in own ponds (pers. opinion). Lake Chozas, León, Northwest Spain, has been regularly monitored since 1994 without any occurrence of the red swamp crayfish, the presence of red swamp crayfish in 1997 indicated that an introduction may have occurred in 1995 or 1996 (Rodriguez et al. 2003). In 1999, the distribution of the red swamp crayfish was deeply analysed by grid in Extremadura region and sampling a total of 79,03% of this region, the occurrence of the red swamp crayfish was detected in a 69,77% (see Figure 3 in Pérez-Bote et al. 2000). In La Rioja region, the red swamp crayfish is found along the watershed of Ebro River, downstream of its tributaries (e.g.: Tirón, Oja, Najerilla, Leza-Jubera, Cidacos or Alhama rivers) and irrigation channels (Gobierno de La Rioja). Currently, the most part of Spanish territory (excepting high altitudes) is infected by the red swamp crayfish (Kouba et al 2014; GBIF 2019; iNaturalist 2019).
Switzerland About 1989, the red swamp crayfish was illegally introduced in two ponds (Schübelweiher and Rumensee) near Zurich probably to replace the Astacus astacus population that had collapsed due to crayfish plague Aphanomyces astaci. In middle of 90s, estimated population size was around 13,000 crayfish and Swiss Prime Court decided to poison the ponds to avoid an increase of those populations; however, it was immediately challenged by the local population. Then, other options were valuated as for example to try to minimise emigration or diminish the red swamp crayfish population by intensive trapping. Moreover, introductions of predatory fishes or eels
73 Chapter I were also considered as treatments to avoid the spreading of this population (Frutiger et al. 1999).
Sweden No live animals have been reported in wild, however, a 13.6% of international exportations from rice fields in Seville did go to Swedish markets, during 90s (Gutiérrez- Yurrita et al. 1999).
The Netherlands The first specimens of the red swamp crayfish known in the wild were released by a restaurant owner in 1985, in The Hague (Soes and Koese 2010). Initially, it was assumed that the species would not become establish permanently and its presence would therefore be casual. However, it was already well established in the Netherlands by 2010 (see Figure 19a in Soes and Koese 2010; Koese and Evers 2011). It was regularly reported in a number of ponds and streams, especially in the west of the country (Amsterdam, Utrecht and Den Haag), but also near the Belgian border (Breda, Tilburg), and from some localities in the east (Koese and Evers 2011). The distribution of the red swamp crayfish is closely associated with urban concentrations, reflecting the fact that the species mainly entered in the Netherlands through the consumption and aquarium trades (Soes and Koese 2010). Currently, the red swamp crayfish is widely distributed in Western of the country (Kouba et al. 2014; Nagy et al. 2019).
74 How the red swamp crayfish took over the world
APPENDIX II. LIST OF LITERATURE CITED IN INVASION HISTORY BY COUNTRIES (APPENDIX I) (190 References)
Adão, H. (1992). Procambarus clarkii (Girard, 1852) (Decapoda, Cambaridae) espécie exótica em Portugal. Consideraçôes sobre a ecobiologia e aspectos da sua biologia populacional na barragem de Monte Novo (Alentejo, Portugal). Tesina. Universidad de Evora (Portugal). Ahn, D., Kawai, T., Kim, S., Rho, H. S., Jung, J. W., Kim, W., ... & Min, G. (2006). Phylogeny of northern hemisphere freshwater crayfishes based on 16S rRNA gene analysis. Korean Journal of Genetics, 28(2), 185. Algarín, S. (1980). Problemática y perspectivas de la introducción del cangrejo. El Cangrejo Rojo de la Marisma. España: Junta de Andalucía, Consejería de Agricultura y Pesca, 25-31. Alonso, F., Temiño, C., and Diéguez-Uribeondo, J. (2000). Status of the white-clawed crayfish, Austropotamobius pallipes (Lereboullet, 1858) in Spain: distribution and legislation. Bulletin Français de la Pêche et de la Pisciculture, 356, 31-53. Alonso, F. and Martínez, R. (2011). La dispersión de los cangrejos rojo y señal en Castilla-La Mancha: ¿son válidas las medidas de gestión de especies invasoras a nivel geográfico de comunidad autónoma? Foresta, 47-48, 244-252. Álvarez, F., Villalobos, J.L., Elías-Gutiérrez, M., and Rivera, G. (2011). Crustáceos dulceacuícolas y terrestres de Chiapas. In: Álvarez F (ed), Chiapas: Estudios sobre su diversidad biológica. Instituto de Biología – UNAM, pp 209–298. Anonymous. (1987). Wildlife Newsletter. Exotic freshwater crayfish in the Cape Province. African Wildlife, 41: 89. Aquiloni, L., Tricarico, E., and Gherardi, F. (2010) Crayfish in Italy: distribution, threats and management. International Aquatic Research, 2 (1), 1-14. Arias-Pineda, J. Y., and Rodríguez, W. D. (2012). First record of the invasive species Procambarus (Scapulicambarus) clarkii (Girard 1852) (Crustacea, Decapoda, Cambaridae) from the Colombian Eastern Cordillera. Boletín de la Sociedad Entomológica Aragonesa, 51, 313-315. Arias-Pineda, J. Y., and Rodríguez, W. D. (2018). Presencia del cangrejo rojo Procambarus clarkii (Girard, 1852) en la sabana de Bogotá, Colombia. Boletín de la Sociedad Entomológica Aragonesa, 62, 283-286. Arrignon, J. C. V., Gérard, P., Krier, A., and Laurent, P. J. (1999). The situation in Belgium, France and Luxembourg. Crustacean Issues, 11, 129-140. Baldaccini, G. N. (1995). Considerazioni su alcuni macroinvertebrati dell’area umida del Massaciuccoli (Toscana). Il Bacino del Massaciuccoli, 91-103. Barbaresi, S., and Gherardi, F. (2000). The invasion of the alien crayfish Procambarus clarkii in Europe, with particular reference to Italy. Biological invasions, 2 (3), 259-264. Barbaresi, S., Gherardi, F., Mengoni, A., and Souty-Grosset, C. (2007). Genetics and invasion biology in fresh waters: A pilot study of Procambarus clarkii in Europe. In: Biological invaders in inland waters: profiles, distribution, and threats (ed. by F. Gherardi), pp. 381-400. Invading Nature: Springer Series in Invasion Ecology, Springer, Dordrecht, The Netherlands. Behm, D. (2009). Bleach set to eradicate Germantown's invasive crayfish. Milwaukee Journal
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Sentinel (November 7). Available: http://www.jsonline.com/news/ozwash/69487982.html. (November 2011). Bernardo, J. M., Costa, A. M., Bruxelas, S., and Teixeira, A. (2011). Dispersal and coexistence of two non-native crayfish species (Pacifastacus leniusculus and Procambarus clarkii) in NE Portugal over a 10-year period. Knowledge and Management of aquatic Ecosystems, 401, 28, p1-13. Boets, P., Lock, K., Cammaerts, R., Plu, D., and Goethals, P. L. M. (2009). Occurrence of the invasive crayfish Procambarus clarkii (Girard, 1852) in Belgium (Crustacea: Cambaridae). Belgian Journal of Zoology, 139, 173-175. Brasher, A. M. D., Luton, C. D., Goodbred, S. L. and Wolff, R. H. (2006). Invasion patterns along elevation and urbanization gradients in Hawaiian streams. Transactions of the American Fisheries Society, 135 (4), 1109-1129. Brock, V. E. (1960). The introduction of aquatic animals into Hawaiian waters, Internationale Revue der Gesamten. Hydrobiologie 45: 463-480. Cabrera, J. (1994). Morphometric relationships and yield in Costa Rican Procambarus clarkii (Decapoda: Cambaridae). Revista de Biología Tropical, 42, 743-744. Campos, E., and Rodríguez-Almaraz, G. A. (1992). Distribution of the red swamp crayfish Procambarus clarkii (Girard, 1852) (Decapoda: Cambaridae) in Mexico: an update. Journal of Crustacean Biology, 12 (4), 627-630. Campos, M. (2005). Procambarus (Scapulicambarus) clarkii (Girard, 1852) (Crustacea: Decapoda: Cambaridae). Una langostilla no nativa de Colombia. Revista Academia Colombiana de Ciencias, 29 (111), 295-302. Caricato, G., Canitano, M., and Montemurro, M. (2013). Alien vs native species into freshwaters of Basilicata. Quaderni ETP, 35, 11-20. CETE de l’Est (2011). Etat des lieux des espèces animales exotiques envahissantes en Champagne- Ardenne. 72 pp. Changeux, T. (2003). Evolution de la répartition des écrevisses en France métropolitaine selon les enquêtes nationales menées par le Conseil Supérieur de la Pêche de 1977 à 2001. Bulletin Français de la Pêche et de la Pisciculture, 370-371, 15-41. Chen, R. T., Ho, P. H., and Lee, H. H. (2003). Distribution of exotic freshwater fishes and shrimps in Taiwan. Endemic Species Research, 5, 33-46. Cheung, S. C. H. (2010). The Social Life of American Crayfish in Asia. In Globalization, Food and Social Identities in the Asia Pacific Region, ed. James Farrer. Tokyo: Sophia University Institute of Comparative Culture. Chiesa, S., Scalici, M., and Gibertini, G. (2006). Occurrence of allochthonous freshwater crayfishes in Latium (Central Italy). Bull Fr Pêche Piscic, 380-381, 883-902. Chucholl, C. (2011). Disjunct distribution pattern of Procambarus clarkii (Crustacea, Decapoda, Astacida, Cambaridae) in an artificial lake system in Southwestern Germany. Aquatic Invasions, 6 (1), 109-113. Cilenti, L., Alfonso, G., Gargiulo, M., Chetta, F. S., Liparoto, A., D’Adamo, R., and Mancinelli, G. (2017). First records of the crayfish Procambarus clarkii (Girard, 1852) (Decapoda, Cambaridae) in Lake Varano and in the Salento Peninsula (Puglia region, SE Italy), with review of the current status in southern Italy. BioInvasions Records, 6 (2), 153-158.
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Rhoades, R. (1944). The crayfishes of Kentucky, with notes on variation, distribution and descriptions of new species and subspecies. The American Midland Naturalist, 31 (1), 111-149. Rhoades, P. B. (1976). Personal Communication – Conversation, 15 April 1976. Address: Central and Southern Florida Flood Control District, Box V, West Palm Beach, Fla. 33402, U.S.A. Riascos, L., Geerts, A. N., Oña, T., Goethals, P., Cevallos-Cevallos, J., Berghe, W. V., ... & van der Heyden. (2018). DNA-based monitoring of the alien invasive North American crayfish Procambarus clarkii in Andean lakes (Ecuador). Limnologica, 70, 20-25. Richter, K. (2000). Ecological and Behavioural Studies on the Red Swamp Crayfish, Procambarus clarkii (Girard) as Introduced Species in Britain. University of North London. Riegel, J. A. (1959). The systematics and distribution of crayfishes in California. California Fish and Game, 45(1), 29-50. Re-Araujo, A. D. (1994). Evaluación de forrajes experimentales para Procambarus clarkii (Crustacea: Decapoda). Revista de biología tropical, 42(1-2), 347-352. Rodríguez, G., and Suárez, H. (2001). Anthropogenic dispersal of decapod crustaceans in aquatic environments. Interciencia, 26(7), 282-288. Rodriguez, C. F., Bécares, E., and Fernández-Aláez, M. (2003). Shift from clear to turbid phase on Lake Chozas (NW Spain) due to the introduction of American red swamp crayfish (Procambarus clarkii). Hydrobiologia, 506-509, 421-426. Rodríguez-Almaraz, G. A., and García-Madrigal, M. S. (2014). Crustáceos exóticos invasores. Especies acuáticas invasoras en México 1st edn. Comisión nacional para el conocimiento y Uso de la Biodiversidad, México, 347-371. Rosecchi, E., Poizat, G., and Crivelli, A. J. (1997). Introductions de poissons d'eau douce et d'écrevisses en Camargue: historique, origines et modifications des peuplements. Bulletin Français de la Pêche et de la Pisciculture, (344-345), 221-232. Ruiz, G. M., Fofonoff, P. W., Carlton, J. T., Wonham, M. J., and Hines, A. H. (2000). Invasion of coastal marine communities in North America: apparent patterns, processes, and biases. Annual Review of Ecological Systematics, 31, 481-531. Saad, A. A., and Emam, W. M. (1998) Field and Semi-field studies on the crayfish The red swamp crayfish in Egypt. Egypr, J. Aquat. Biol. & Fish, 2 (4), 331-344. Salvador, R. E., and Leyton, S. A (2000). Efectos ecológicos producidos por la introducción del cangrejo de río Procambarus clarkii, (Girard, 1852) en la zona de Taura. Guayquil Thesis. Sako, N. (1987). Procambarus clarkii and Rana catesbeiana in Japan: a history. Collect. Breed., 49, 396-397. In Japanese. Schainost, S. C. (2016). The Crayfish of Nebraska. Lincoln: The Nebraska Game and Parks Commission. http://digitalcommons.unl.edu/nebgamewhitepap/69 Schoonbee, H. J. (1993). Occurrence of the red swamp crawfish Procambarus clarkii (Crustacea: Cambaridae) in the Crocodile River at Dullstroom, Transvaal. Water SA, 19, 163-166. Schuster, G. A., and Taylor, C. A. (2004). Report on the crayfishes of Alabama: literature and museum database review, species list with abbreviated annotations and proposed conservation statuses. INHS Center for Biodiversity. Shelton-Nix, E. (Ed.). (2017). Alabama Wildlife (Vol. 5). University of Alabama Press.
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86
CHAPTER II
Unravelling the global invasion routes of a worldwide invader, the
red swamp crayfish (Procambarus clarkii)
Oficialdegui FJ, Clavero M, Sánchez MI, Green AJ, Boyero L, Michot TC, Klose K,
Kawai T, Lejeusne C
Freshwater Biology (2019): in press, doi: 10.1111/fwb.13312
Chapter II
ABSTRACT
Understanding how introduced species succeed and become widely distributed within non-native areas is critical to reduce the threats posed by them. Our goal was to reconstruct the main invasion routes and invasion dynamics of a global freshwater invader, the red swamp crayfish, Procambarus clarkii, through the analysis of its genetic variability in both native and invasive ranges.We inferred invasion routes and population structure from the analysis of a fragment (608bp) of the mitochondrial marker COI from 1,062 individuals of P. clarkii in addition to 354 GenBank sequences, for a total of 122 populations (22 natives and 100 invaded). Genetic structure was assessed using analysis of molecular variance and non-metric multidimensional scaling analyses. We analysed haplotype frequencies for the genetic variability in each locality and region. The haplotype network was depicted by using PopART software. A high haplotype diversity was found in the native range (Hd: 0.90), but also in some non- native areas, such as western United States (Hd: 0.80), areas of Mexico (Hd: 0.78) and some hotspots in Europe (e.g., southern Spain or Italy), suggesting a complex pattern of multiple introductions. We grouped all localities in five differentiated groups according to biogeographic origin: the native area, West Americas, East United States, Asia and
Europe. Additionally, the identification of 15 haplotypes shared between at least two localities, the phylogenetic network estimation and indices of genetic differentiation among localities allowed us to identify a large genetic admixture in the native range; the two independent invasion routes (i.e. westwards and eastwards) in US from the native range (Louisiana and Texas) with translocations within each area; a stepping-stone introduction from US to Japan (involving few individuals) themselves introduced to
China afterwards; the entry of P. clarkii from Louisiana (US) into southern Spain and their multiple secondary introductions over Europe as well as other possible introductions in central Europe. Our study emphasizes the need for unravelling the global invasion routes and the demographic processes underlying the introduction of exotic species (i.e., admixture, bridgehead invasion effect and propagule pressure) to control the spread of invasive species. Our findings highlight the value of genetic
90 Unravelling the global invasion routes of the red swamp crayfish analyses to identify the geographic origin of source populations as well as the variability of invaded areas in order to reconstruct invasion dynamics and facilitate management of invasive species (e.g. through environmental DNA monitoring).
INTRODUCTION
Humans have transported species across biogeographical barriers and introduced them to new territories for millennia (Forcina et al., 2015), but large-distance movements of species have increased exponentially in recent decades (Hulme, 2009; Lenda et al.,
2014) driving to an homogenization of biotas that has involved the break-down of long- established biogeographical barriers (Capinha et al., 2015). Those species that are transported by humans, released into new environments, able to survive, establish self- sustained populations, thrive, become abundant and spread geographically, are considered invasive species (Jeschke et al., 2014). Biological invasions today are perceived as major components of global change, with severe negative environmental
(Simberloff et al., 2013; Blackburn et al., 2014; Jeschke et al., 2014) and socio-economic impacts (Vilà et al., 2010). To manage invasive species, it is of vital importance to identify invasion routes (De Kort et al., 2016). However, most knowledge about the transport routes of invasive species is based on historical and observational data, which are usually scarce, confusing and sometimes inaccurate (Roman, 2006; Haydar, 2012).
Population genetic studies provide valuable tools to identify areas of geographic origin of introductions, to detect single versus multiple introductions, and to describe expansion patterns (Lejeusne et al., 2014; Cristescu, 2015; Blakeslee et al., 2017; O’Hanlon et al., 2018; Fang et al., 2018), though caution must be used when interpreting demographic history over such short timescales (Fitzpatrick et al., 2012). Such information can be useful for the management of invasive species and for the prevention of future introductions (Estoup & Guillemaud, 2010), and phylogeographic studies have been proposed as an integral tool of biodiversity conservation planning (van de
Crommenacker et al., 2015).
91 Chapter II
Many invasive species have high economic value, which often results in their deliberate introduction by humans into non-native areas where they can spread rapidly due to secondary introductions (see Audzijonyte et al., 2017; Cao et al., 2017; Huang et al., 2017). Unlike accidental introductions that are facilitated by humans, species that are deliberately introduced may have a higher chance of success because humans take action to ensure such success (Pyšek et al., 2011). For example, deliberate introductions often involve a high propagule pressure (i.e. number of introduction events and/or size of propagules; Lockwood et al., 2005; Simberloff, 2009), the genetic admixture of introduced populations (i.e. the mixing of populations from genetically distinct source populations; Dlugosch & Parker, 2008; Rius & Darling, 2014; Hufbauer, 2017), and the subsequent invasive bridgehead effect (i.e. a particularly successful invasive population serves as a source for new introductions, Lombaert et al., 2010; Estoup & Guillemaud,
2010).
Due to the high intensity of human disturbances and the high connectivity among inland water systems, freshwater environments are especially susceptible to invasions
(Strayer, 2010; Havel et al., 2015; Tricarico et al., 2016). Also, many freshwater species can be harvested from the wild and/or cultivated in farms for commercial purposes, providing a high socioeconomic value (e.g., aquaculture; FAO, 2011). Among these, several invertebrates are valued either for human consumption or as food for other cultured animals (e.g. shrimp or prawn for fishes, crayfish for bullfrogs, etc.), which provides a high economic return (Resh & Rosenberg, 2015). Freshwater crayfish are favoured for farming since they do not have a larval phase and are polytrophic, they are relatively easy to rear compared with other cultured crustaceans (Holdich, 1993), and their consumption has a long-lasting tradition in many regions worldwide (Gherardi,
2011). The red swamp crayfish, Procambarus clarkii (Girard 1852), native to southern USA and northern Mexico, has been successfully introduced into all continents except
Australia and Antarctica (Loureiro et al., 2015) mainly due to its economic value (Hobbs et al., 1989). Owing to its biological and ecological characteristics, this crayfish is considered one of the worst invasive species worldwide, causing serious damage to
92 Unravelling the global invasion routes of the red swamp crayfish biodiversity (e.g., other crayfish species, fish, amphibians, macroinvertebrates and macrophytes) and to human infrastructure and ecosystem services (e.g., irrigation canals, water quality, rice crops, etc.) (Geiger et al., 2005; Twardochleb et al., 2013).
Procambarus clarkii is one of the most economically valuable aquatic species to be farmed
(Huner, 2002; Souty-Grosset et al., 2016), generating tens of billions of US dollars (USD) per year in the world
(http://www.fao.org/fishery/culturedspecies/Procambarus_clarkii/en#tcNA0064).
The first introductions of P. clarkii out of its native area took place in the early 20th century, when it was taken to the Hawaiian Islands (1923), the Pacific drainages of USA
(1924), Japan (1927) and China (1929), with different motivations including aquaculture, fishing activities and food for cultured American bullfrogs Lithobates catesbeianus
(Holmes, 1924; Penn, 1954; Brasher et al., 2006). Procambarus clarkii was often able to spread rapidly, occupying the rivers and lakes of non-native areas (e.g. Riegel, 1959, for
California; Yue et al., 2010, for China). In the mid-1960s, a batch of crayfish was sent to
Uganda from Louisiana, then translocated to Kenya, and later to other African countries
(Huner, 1977; Lowery & Mendes, 1977). Concurrently, it artificially spread out of its native area in Mexico, and then to Costa Rica, Puerto Rico, Venezuela, and the
Dominican Republic in the 1970s (Huner, 1977), eventually reaching Brazil in the mid-
1980s (Huner, 1986). In Europe, it was deliberately and legally introduced into Spain
(Badajoz and Seville in 1973 and 1974, respectively) from Louisiana (Habsburgo-Lorena,
1978; 1986). In only 45 years, P. clarkii has colonized many countries in Europe, being widely established in Spain, Portugal, France, Italy, Belgium, Netherlands, Germany and the United Kingdom (see Kouba et al., 2014 for the entire European distribution of this species).
To date, most genetic studies of P. clarkii have focused on genetic variability at a regional scale (e.g., Barbaresi et al., 2007; Torres & Álvarez, 2012; Quan et al., 2014; Yi et al., 2018; Almerão et al., 2018). Of these, very few studies have attempted to unveil the invasion routes, and when performed, they did so only at a regional scale. Hence, almost nothing is known about the population genetics and invasion routes of P. clarkii at a
93 Chapter II global scale. The general objective of this study is to provide a comprehensive overview of the global invasion history of P. clarkii. We included not only most of the non-native range of this species in the Northern Hemisphere, but also an exhaustive sampling of its native area in order to confirm the invasion sources and routes previously in the literature and detect previously unreported ones. Hence, our specific objectives were: 1) to describe the invasion dynamics of P. clarkii at continental and global scales, identifying the main invasion routes; and 2) to examine the genetic variability and population structure of P. clarkii in the native area and across the non-native range, with special focus on Europe, to reveal potentially unreported introductions not cited in the literature.
METHODS
Sampling
We collected 1,062 specimens of P. clarkii from 72 localities: 15 native (States of
Louisiana and Texas, USA) and 57 non-native localities distributed within the Northern
Hemisphere (i.e., western US, eastern US, Europe and Japan) (Table 1 and Fig. 1).
Crayfish were individually preserved in 96% ethanol. Average sample size per locality was 14.7 ± 6.6 individuals (mean ± SE; range 2-21) (Table 1). We included in our dataset the information for 354 additional individuals from 7 native and 43 non-native localities that we obtained from data already published in previous studies (Genbank Accession numbers: AY701195; JF438001- JF438004; JN000898- JN000908; JX120103- JX120108) available from Taylor & Knouft (2006), Filipová et al. (2011), Torres & Álvarez (2012) and Li et al. (2012), respectively. Thus, a total of 1,416 individuals from 22 native and
100 introduced localities (Fig. 1a; Table 1) were used for this study. The sequences recently published by Almerão et al. (2018) were not added into our global analyses of this study because our sequences were larger than theirs. Even so, we compared their results in a subsequent analysis (see Supplementary Material, Table S1 for synonymous haplotypes).
94 Unravelling the global invasion routes of the red swamp crayfish
Table 1. Genetic diversity parameters based on the COI gene for each Procambarus clarkii locality. Note that localities are grouped in biogeographical zones (see Methods). Sequences retrieved from GenBank are shown in italics (see references in Materials and Methods section). *: 0.05 > p > 0.01; **: 0.01 > p > 0.001.
Country Locality Code Lon Lat N h Hd π R2 Tajima Fs NATIVE RANGE 179 39 0.902 0.0055 East Louisiana Poison LA2 30.22 -91.61 20 5 0.795 0.0035 0.172 0.810 0.854 Haha Bay LA4 30.15 -91.63 20 8 0.775 0.0040 0.104 -0.519 -1.573 Baton Rouge BAT 30.37 -91.19 4 3 0.833 0.0030 0.265 1.090 0.006 Morgan City MOR 29.77 -91.13 10 6 0.867 0.0043 0.157 0.215 -1.164 Pierre Part PIE 29.95 -91.28 10 8 0.956 0.0045 0.135 -0.144 -3.882** Des Allemands DES 29.80 -90.51 4 4 1.000 0.0063 0.114* 0.039 -0.884 Jean Lafitte JEA 29.73 -90.08 9 4 0.694 0.0027 0.172 -0.526 -0.061 New Orleans LAf 29.95 -90.08 1 1 West Louisiana Abbeville ABB 29.91 -92.20 4 3 0.833 0.0025 0.276 -0.754 -0.288 Alexandria ALE 31.10 -92.49 4 3 0.833 0.0088 0.327 -0.222 1.606 Woodworth WOO 31.19 -92.47 4 4 1.000 0.0036 0.223 -0.065 -1.741* Natchitoches NAT 31.74 -93.08 17 6 0.765 0.0082 0.199 1.495 2.133 Kaplan LAt 29.99 -92.26 5 2 0.600 0.0020 0.300 1.459 1.688 Calcasieu L. LO 29.87 -93.26 10 4 0.533 0.0064 0.143 -0.352 2.256 Upstream Mississippi River Monroe MON 32.50 -91.67 5 4 0.900 0.0072 0.218 0.132 0.286 Memphis MEM 35.37 -90.03 5 2 0.400 0.0007 0.400 -0.817 0.090 Horseshoe ILL 37.14 -89.34 1 1 Texas Comal COM 29.71 -98.13 18 6 0.686 0.0017 0.103 -0.917 -2.350* San Marcos SMA 29.88 -97.93 14 3 0.560 0.0010 0.157 -0.011 -0.072 Mexico Sabinas Hidalgo NL 26.48 -100.22 4 1 Río Jiménez CON 29.15 -100.76 5 2 0.400 0.0007 0.400 -0.817 0.090 Río Sabinas COC 27.97 -101.58 5 1 WEST AMERICAS Non-native US Santa Ynez SYZ 34.56 -119.88 10 3 0.378 0.0019 0.187 -1.388” 0.762 Topanga TOP 34.06 -118.59 20 5 0.679 0.0086 0.194 1.541 4.185 Ventura VEN 34.35 -119.30 4 2 0.667 0.0044 0.333 2.080 2.719 Pine PIN 47.59 -122.04 21 3 0.581 0.0018 0.186 0.883 1.537 Waverly WAV 44.64 -123.07 20 7 0.832 0.0050 0.121 -0.356 0.057 Waiau HW 19.71 -155.15 3 1 Non-native Mexico 13 4 0.782 0.0056 Teopisca CHIS 16.55 -92.48 5 1 El Arenal DU 24.04 -104.43 5 2 0.600 0.0049 0.300 1.686 3.526 Las Varas CHt 29.80 -106.69 3 1 Costa Rica Cachí Dam CR 9.83 -83.82 4 2 0.500 0.0008 0.433 -0.612 0.172
95 Chapter II
Table 1. (Cont.) Country Locality Code Lon Lat N h Hd π R2 Tajima Fs EAST USA Pee Dee FOR 36.15 -80.29 4 2 0.500 0.0008 0.433 -0.612 0.172 Pamplico LEN 35.24 -77.56 10 2 0.200 0.0003 0.300 -1.112 -0.339 Albemarle PER 36.27 -76.38 21 3 0.338 0.0010 0.111 -0.707 0.204 North Shore CHI 42.03 -87.71 20 5 0.442 0.0011 0.105 -1.888* -2.091* EUROPE Spain 355 13 0.469 0.0024 Balboa EXT 38.88 -6.87 20 2 0.395 0.0007 0.197 0.723 0.976 Manecorro MAN 37.12 -6.49 20 5 0.716 0.0044 0.143 0.214 1.571 Cantaritas AR4 37.05 -6.21 20 5 0.663 0.0045 0.140 0.243 1.596 Colomera GRA 37.38 -3.72 20 3 0.468 0.0030 0.154 0.255 2.904 Hueznar HUE 37.93 -5.70 20 2 0.100 0.0007 0.218 -1.868* 0.998 Arreo ALA 42.78 -2.99 20 2 0.479 0.0032 0.239 2.024 5.159 Elorz EL 42.80 -1.67 19 4 0.585 0.0028 0.105 -0.921 1.200 Expo EXP 41.67 -0.91 13 3 0.615 0.0041 0.200 1.009 3.086 Gijón GIJ 43.54 -5.64 15 3 0.257 0.0009 0.121 -1.317 -0.379 Jiloca JIL 40.54 -1.29 15 2 0.419 0.0021 0.210 1.078 3.248 Leza LEZ 42.44 -2.31 20 2 0.268 0.0018 0.134 -0.138 3.143 Almenara ALM 39.76 -0.18 20 1 Brugent BRU 42.01 2.61 20 2 0.337 0.0022 0.168 0.565 3.843 Ecomuseu ECO 40.72 0.72 20 3 0.195 0.0010 0.159 -2.056** 0.136 Alpedrete MAD 40.67 -4.02 20 2 0.442 0.0007 0.221 1.026 1.169 Júcar ALB 39.15 -1.81 11 3 0.618 0.0011 0.192 0.036 -0.113 Mundo MUN 38.46 -1.76 20 1 Sa Pobla SAP 39.79 3.06 5 2 0.400 0.0026 0.4 -1.094” 2.202 Soller SOL 39.79 2.79 5 3 0.700 0.0043 0.205 0.562 1.090 Carucedo CAR 42.49 -6.78 5 2 0.400 0.0013 0.4 -0.973 1.040 Chozas CHO 42.52 -5.71 4 2 0.500 0.0025 0.433 -0.754 1.716 Valparaiso VAL 42.00 -6.29 2 2 1.000 0.0115 0.5 0.000 1.946 Pisuerga VLB 41.80 -4.59 21 3 0.343 0.0017 0.105 -0.742 1.384 Portugal 114 4 0.399 0.0012 Aboboda ABO 38.74 -9.32 15 2 0.343 0.0017 0.171 0.342 2.71 Lousal LOU 38.03 -8.43 20 2 0.479 0.0008 0.239 1.262 1.311 Alpiarça POR 39.25 -8.59 20 4 0.642 0.0029 0.123 -0.727 1.429 R. de Monsaraz REG 38.48 -7.52 20 2 0.505 0.0008 0.253 1.430 1.409 Requeixo REQ 40.59 -8.53 20 2 0.100 0.0002 0.218 -1.164 -0.879” Vila-Rica VILA 41.23 -7.10 19 1 France 84 5 0.561 0.0022 Marais Bruges BOR 44.90 0.60 20 1 Briere BRI 47.34 -2.25 20 1 Tour du Valat CAM 43.51 4.67 20 2 0.526 0.0035 0.263 2.511 5.567 Lamartine TOU 43.51 1.34 21 2 0.095 0.0002 0.213 -1.164 -0.919” Rochechevreux FR1 45.47 1.22 1 1 Rochechevreux FR2 46.47 1.22 1 1 Givrezac FR3 45.40 0.22 1 1 Italy 60 6 0.731 0.0027 Bernate BER 45.49 8.80 20 2 0.479 0.0008 0.239 1.262 1.311 Monterotondo LAZ 42.05 12.55 20 2 0.505 0.0008 0.253 1.430 1.409 Fucecchio TOS 43.81 10.79 20 5 0.716 0.0045 0.150 0.257 1.608 Holland Hardinxveld-Giess HOL 51.82 4.84 20 1 96 Unravelling the global invasion routes of the red swamp crayfish
Table 1. (Cont.) Country Locality Code Lon Lat N h Hd π R2 Tajima Fs Belgium 19 2 0.409 0.0007 Bioul BIO 50.34 4.81 12 1 Ecaussinnes ECA 50.58 4.14 7 2 0.476 0.0008 0.238 0.559 0.589 United Kingdom Hampstead-Heath LON 51.56 -0.16 20 1 ASIA Japan 122 4 0.476 0.0024 Wakamatsu FUK 33.91 130.78 20 1 Hourai HOK 42.94 143.22 20 1 Waga IWA 39.44 140.78 9 2 0.556 0.0027 0.278 1.948 3.276 Ohfuna KAN 35.35 139.53 20 2 0.521 0.0026 0.261 2.266 4.362 Rakusho OKA 34.71 133.93 20 3 0.279 0.0011 0.108 -0.626 0.286 Ohtsu SHI 35.01 135.87 10 2 0.533 0.0026 0.267 1.831 3.338 Kanda TOK 35.69 139.77 13 3 0.410 0.0024 0.164 -1.335” 1.625 Saitama SA 35.85 139.65 10 2 0.533 0.0026 0.267 1.831 3.338 China 293 2 0.350 0.0017 Shanghai SH 31.03 121.23 8 2 0.571 0.0028 0.286 1.982 3.149 Jiaxing JX 30.75 120.77 10 2 0.356 0.0018 0.178 0.021 2.334 Binhu. Wuxi WXB 31.52 120.28 7 2 0.571 0.0028 0.286 1.811 2.920 Nantong NT 32.02 120.87 7 2 0.286 0.0014 0.350 -1.358” 1.514 Xiaba village XB 32.20 118.87 8 2 0.571 0.0028 0.286 1.981 3.149 Wuxi WX 31.57 120.30 8 2 0.536 0.0026 0.268 1.601 2.988 Wangjiang WJ 30.12 116.70 8 1 Maanshan MAS 31.55 118.50 10 2 0.356 0.0018 0.178 0.021 2.338 Chaohu CH 31.62 117.87 8 2 0.429 0.0021 0.214 0.458 2.469 Hefei HF 31.82 117.23 7 2 0.286 0.0014 0.350 -1.358” 1.514 Dingyuan DY 32.28 117.83 10 2 0.556 0.0027 0.278 2.057 3.451 Nanbei Port NBP 29.72 116.17 8 2 0.250 0.0012 0.331 -1.448” 1.415 Zhongxian ZX 30.28 108.03 10 2 0.533 0.0026 0.267 1.831 3.338 Jianyang JY 30.38 104.55 10 2 0.467 0.0023 0.233 1.152 2.985 Chongqing CQS 29.55 106.53 6 1 Ningbo NB 29.88 121.55 7 1 Xuyi-culture XYC 33.00 118.50 7 1 Xuyi-wild XYW 33.03 118.42 10 1 Xiaguan district XG 32.08 118.75 10 1 Baguazhou BGT 32.17 118.82 8 1 Guangfengwei CJR 30.12 116.87 8 1 Sanli township SLT 29.75 116.22 8 1 Poyang lake PYL 28.87 116.43 10 1 Youlan. Nanchang NCYL 28.52 116.12 6 1 Nanhu lake NHL 30.02 114.03 8 1 Yuni Lake YNL 30.00 112.20 7 1 Xiantao XT 30.30 113.40 8 1 Qianjiang QJ 30.40 112.60 10 1 Liangzi lake LZL 30.00 114.00 10 1 Honghu lake HLL 29.70 113.40 8 1 Changhu lake CHL 30.30 112.10 6 1 Yuanjiang YJ 28.85 112.37 10 1 Ningxiang NX 28.28 112.55 8 1 Dongting lake DTL 29.30 113.02 10 1 Dongting Lakeside DTLs 29.35 113.13 9 1 Number of sequences (N), number of haplotypes (h), haplotype diversity (Hd), nucleotide diversity (π). 97 Chapter II
Figure 1. Haplotype frequencies of Procambarus clarkii in the 122 localities distributed worldwide. The size of pie charts is proportional to the sample size. Haplotypes restricted to one sampling locality (i.e., private haplotypes) are coloured in white within pie charts, while haplotypes shared between localities are shaded using colours. Black spots show each one of the 122 localities used, and dark grey areas represent the native range of Procambarus clarkii. a) Global map; b) United States and Mexico; c) Hawaiian Islands; d) Costa Rica; e) close-up of Louisiana (US) within its native range; f) East Asia (China and Japan) and; g) Europe.
DNA extraction and sequencing
Genomic DNA was extracted from muscle tissue (gill tissue at LEN, FOR and PER localities; see Table 1 for more details) using a modified DNA salt-extraction protocol 98 Unravelling the global invasion routes of the red swamp crayfish
(composition: NaCl 25 mM, Tris 12.5 mM (pH 8.0), EDTA 12.5 mM (pH 8.0), 31.5 µL SDS
10%) and proteinase K (Aljanabi & Marinez, 1997). Logistical support was provided by the Laboratorio de Ecología Molecular, Estación Biológica de Doñana, CSIC (LEM-EBD).
A fragment of the mitochondrial gene coding for the cytochrome c oxidase subunit I
(COI) gene was amplified using the primers LCO1490 and HCO2198 (Folmer et al.,
1994). Amplifications were carried out in a 20 µl reaction volume, with 1-5 µl of genomic
DNA, 2 µl of 10x buffer, 0.8 µl of MgCl2 (50 mM), 0.16 µl dNTP (100 mM), 0.5 µl primer
LCO 1490, 0.5 µl primer HCO 2198 and 0.12 µl TAQ polymerase (Bioline). Polymerase chain reaction (PCR) consisted of an initial denaturation step at 94ºC for 5 min, followed by 30 amplification cycles (94ºC for 1 min, 47ºC for 1 min and 72ºC for 1 min) and a final elongation step at 72ºC for 5 min. Sequencing was performed by Macrogen Europe
Company.
Genetic analyses
Sequences were edited using the software SequencherTM v4.9 (Gene Codes Corp.,
© 1991–2009, Ann Arbor, MI 48108). Nucleotide sequences were aligned using the algorithm CLUSTAL W implemented in BioEdit (Hall, 1999). No insertions nor deletions
(indels) were found. A hierarchical series of tests based on the Bayesian Information
Criterion (BIC) was applied to identify the most appropriate nucleotide substitution model among 88 models tested, as implemented in jModelTest 2 (Darriba et al., 2012).
We used the nested model Tamura & Nei (1993) with 133 parameters, 1382.91 –lnL for onwards analyses. DnaSP 6.0 software was used to calculate the number of polymorphic sites (S), haplotype diversity (Hd), nucleotide diversity (π), and total number of synonymous and non-synonymous mutations, for which nucleotide sequences were translated into amino acid sequences using the Drosophila mitochondrial genetic code
(Rozas et al., 2017). The haplotype network was inferred by the TCS method (Clement et al., 2000) implemented in PopART software (Leigh & Bryant, 2015).
Because of a smaller sampling size (1 or 2 individuals), ILL, LAf, FR1, FR2, FR3 and
VAL localities were excluded from downstream analyses. Pairwise ɸST (PhiST) and
99 Chapter II hierarchical analysis of molecular variance (AMOVA) were calculated using Arlequin
3.5 (Tamura & Nei, 1993; Excoffier & Lischer, 2010). To examine the genetic differentiation between any two of the populations, ɸST calculations were calculated assuming gamma-distributed substitution rates using the Tamura and Nei model
(Tamura & Nei, 1993) to compute a distance matrix and 10,000 bootstrap pseudo- replicates were used to estimate the standard error. The p-values were corrected for multiple comparisons using the false discovery rate (FDR) control according to the
Benjamin and Hochberg (BH) correction method (Benjamin & Hochberg, 1995). To ascertain the genetic structure of populations, AMOVAs were performed based on
10,000 random permutations. Due to the large native range of P. clarkii, we classified native localities into five groups according to natural (e.g. river catchments) and administrative (e.g. country or state frontiers) boundaries: Mexico, Texas, Louisiana east
(east of the Atchafalaya River), Louisiana west (west of the Atchafalaya River), and
Mississippi River upstream (upstream starting at Monroe, Memphis and north to
Illinois). However, for all datasets, two different a priori hypotheses were tested: (a) native versus introduced localities, and (b) population grouping according to biogeographical distribution of this species into 5 zones: (1) native area, (2) West
Americas which included all samples from the USA west of Texas, including California,
Oregon and Washington State, plus all samples from Hawaii and invaded Central
America, (3) East United States (from Louisiana to the Atlantic Ocean and Chicago), (4)
Asia, and (5) Europe. Another a priori hypothesis was analysed exclusively for European populations to test whether there were one (i.e., the whole of Europe) or two genetic clusters within Europe (i.e., southern and northern areas of P. clarkii distribution). A dissimilarity matrix of Jost’s Dest distances was also calculated with 10,000 replicates using SPADE (Jost, 2008; Chao & Shen, 2012). Based on ɸST and Dest estimates, two non- metric multidimensional scaling (NMDS) analyses were used to graphically represent the differentiation among localities and their respective zones (described above) using the vegan package in R (Oksanen, 2013).
100 Unravelling the global invasion routes of the red swamp crayfish
RESULTS
Among the 1,416 specimens of P. clarkii analysed, we obtained a matrix of 608 base pairs (bp) of the cytochrome c oxidase subunit I (COI) with 54 polymorphic sites, yielding 65 haplotypes. Sequences of all haplotypes were submitted to GenBank and assigned Accession Numbers: MK026671 - MK026735. Most of the nucleotide substitutions were synonymous, but four non-synonymous changes were identified (2 substitutions at the 1st position corresponding to a change from an Isoleucine to a Valine, and from a Methionine to a Valine, respectively; and 2 substitutions at the 2nd position corresponding to a change from a Valine to an Alanine, and from a Threonine to a
Methionine, respectively). Of these four non-synonymous changes, one singleton was found in the ALB locality (Spain) and three parsimony sites were located at SMA and
COM localities in Texas (US) and in the DU locality in the invaded area of Mexico (see abbreviations in Table 1 and marked in haplotype network in Fig. 2).
The overall haplotype diversity (Hd) and nucleotide diversity (π) were 0.76 and
0.0040, respectively. The native area showed the highest haplotype and nucleotide diversity (0.90 and 0.0055, respectively), with the highest figures being found in WOO,
DES, MON and PIE localities (Table 1). For invaded areas, haplotype and nucleotide diversities varied considerably between regions: Hd: 0.80 and π: 0.0048 in non-native
US, Hd: 0.78 and π: 0.0056 in non-native Mexico, Hd: 0.46 and π: 0.0023 in Asia and Hd:
0.58 and π: 0.0022 in Europe, respectively (for localities see Table 1). Of the entire dataset, a total of 15 haplotypes were shared between at least two sampling localities (Fig. 1), of which Hap_04 was present at high frequency almost worldwide, irrespective of the native (up to 13 localities) or non-native (up to 76 localities) status of the populations.
Other haplotypes were shared between continents, including coincidences between
US and Europe (Hap_01, Hap_03, Hap_05, Hap_09 and Hap_29) and North America and Asia (Hap_02 found in invaded localities in Mexico, California, Japan and China, and Hap_40 shared between the native area and Japan). Conversely, 50 haplotypes were restricted to one locality (i.e., private haplotypes), 29 of them found in the native area
101 Chapter II and 21 being exclusive to one of the invaded localities (see Supplementary Material,
Table S2).
Figure 2. Haplotype network (statistical parsimony-based) for cytochrome c oxidase subunit I (COI) sequences of the red swamp crayfish, Procambarus clarkii. Each circle represents one haplotype and its size is proportional to the haplotype frequency. Within the network, each line between haplotypes represents a mutational change and small black dots show unsampled haplotypes inferred from the data. Localities from the same geographical region share the same colour. Haplotypes with non-synonymous changes are indicated by *.
The statistical parsimony haplotype network showed a star-like structuring centred around the Hap_04, which appeared in almost half of sampled crayfish (618 specimens) geographically widely distributed over all zones (13 native and 62 invaded localities)
(Fig. 2 and see Supplementary Material, Table S2). Moreover, there were other smaller star-like structuring around three haplotypes (Hap_01, Hap_20 and Hao_09). Hap_01 was mainly distributed in the US, both in its native (9 localities in Louisiana) and non- native range (TOP and PIN in western US; and PER, LEN, FOR and CHI in Atlantic area), but also was found in two Spanish localities (ECO and GIJ). Hap_20 was found in
Louisiana and across the western US. In addition, it closely joined (1 mutation) to 102 Unravelling the global invasion routes of the red swamp crayfish
Hap_28 which is widely found in Asia. Hap_09 was broadly distributed among the native localities in Louisiana, as well as in three invaded North American localities, and two Spanish and one French locality (AR4, CHO and FR1, respectively). In addition, this central haplotype was connected by only one mutation with Hap_40 which was present in Japan. A thorough analysis of the haplotype network in the native range (see
Supplementary Material, Fig. S1) showed no clear genetic structure except for Texas localities (Hap_15-19 and Hap_48-50), Mexico localities (Hap_61-62) and the Hap_20 which was mainly found in southeastern Louisiana. Additionally, Hap_04, Hap_01 and
Hap_09 were widely distributed over all localities in the native range as well as many localities grouped evolutionarily differentiated haplotypes (e.g., MON, NAT or LO localities), indicating a large genetic admixture in Louisiana.
Table 2. Analysis of Molecular Variance (AMOVA) within the native area of Procambarus clarkii, giving corresponding values for FCT (difference among groups), FSC (differences among localities within groups), and FST (differences among all localities). Five groups were considered: the native localities in Mexico, Texas, east Louisiana, west Louisiana and upstream Mississippi River.
Source of Variation d.f. Sum of squares Variance components Percentage of variation Mexico - Texas – E. Louisiana – W. Louisiana – Upstream Mississippi River Among groups 4 77.228 0.53574 29.04
(FCT = 0.290, p < 0.000) Among localities 15 39.135 0.17294 9.38 within groups
(FSC = 0.132, p = 0.002) Within localities 157 178.35 1.13598 61.58
(FST = 0.384, p < 0.000) Total 176 294.712 1.84466
Due to the large native area of P. clarkii, we tested whether the native area clustered into 5 groups (Mexico, Texas, Mississippi River upstream, Louisiana east, and Louisiana west). As haplotype network showed, AMOVA also revealed a slight genetic structure within the native area, where a small fraction of the total variance was due to between- group variance (29.0%); nevertheless, most of the variance was explained by variation within populations (61.6%) (Table 2).
103 Chapter II
Table 3. Analysis of Molecular Variance (AMOVA) among the native and introduced localities of Procambarus clarkii worldwide and among the five zones (native range, west Americas, east USA,
Europe and Asia), listing the corresponding values for FCT (difference among groups), FSC
(differences among localities within groups), and FST (differences among all localities).
Source of Variation d.f. Sum of squares Variance components Percentage of variation Native Area – Invaded Area Among groups 1 69.842 0.1976 14.35
(FCT = 0.143, p < 0.000) Among localities 114 932.781 0.63021 45.76 within groups
(FSC = 0.534, p < 0.000) Within localities 1293 710.207 0.54927 39.89
(FST = 0.601, p < 0.000) Total 1408 1712.829 1.37708 Native Area – West Americas – East USA – Europe – Asia Among groups 4 491.791 0.49942 36.04
(FCT = 0.360, p < 0.000) Among localities 111 510.831 0.33716 24.33 within groups
(FSC = 0.380, p < 0.000) Within localities 1293 710.207 0.54927 39.63
(FST = 0.604, p < 0.000) Total 1408 1712.829 1.38585
This may be due to the high proportion of private haplotypes (Fig. 1b and Fig. 1c).
For the whole dataset, native and invaded areas worldwide were not clustered into two different genetic groups because only 14.3 % of total variation was due to differences between groups (Table 3), indicating the high variability of P. clarkii in the invaded range. However, after classifying the whole set of localities according to their biogeographical ranges (see Methods), 39.6% of total variation was still explained by differences within localities, but 36.0% of total variance was due to differences among these 5 established zones (Table 3).
This result seems to indicate a slight genetic structure among these zones. In
Europe, a moderate genetic structure was detected between the northern and southern distribution areas of P. clarkii (Fig. 1g), where Hap_04 was predominant in South Europe
104 Unravelling the global invasion routes of the red swamp crayfish and Hap_11 in Central Europe. AMOVA analyses showed that 40.7 % of explained variance was due to differences between both genetic clusters (Table 4).
Table 4. Analysis of Molecular Variance (AMOVA) within Europe between northern and southern distribution of Procambarus clarkii, listing the corresponding values for FCT (difference among groups), FSC (differences among localities within groups), and FST (differences among all localities).
Source of Variation d.f. Sum of squares Variance components Percentage of variation North (UK, HOL, BIO, ECA, BRI) – South European distribution (rest of European localities) Among groups 1 59.646 0.40773 40.74
(FCT = 0.407, p < 0.000) Among localities 37 101.955 0.13481 13.47 within groups
(FSC = 0.227, p < 0.000) Within localities 628 287.734 0.45817 45.78
(FST = 0.542, p < 0.000) Total 666 449.334 1.00071
In a NMDS plot based on Dest distances (Fig. 3), most localities remained within the
95% confidence intervals (CI) of the native area group, except some localities from Japan,
China and western North America, in which Hap_28 was present at high frequency.
This is due to the fact that we did not find this haplotype in the native area, despite the exhaustive sampling done there. All localities from eastern US were closely grouped within the range of the native area. However, the localities from western North America were more different from each other, for instance PIN was closer to European localities, whereas VEN and CH were more similar to Asian locations. In addition, the other locality from western North America not only had a clear proximity to each other but also to localities from within the native area (COM, SMA, COC and NL), indicating a similarity among them. The result of AMOVA analysis for European populations, where two genetic clusters were found between localities from the northern and southern
European distributions, was also reinforced by NMDS plot. LON, HOL, ECA, BIO and
BRI localities (depicted by a green triangle in the NMDS plot) were situated all together outside the 95% CIs of the European group, having greater proximity to Texas and
105 Chapter II
Mexico localities from the native area and those from western America than to south
European localities.
Figure 3. NMDS analysis on Dest Jost distances. The graph depicts the pairwise dissimilarity between localities in a low-dimensional space where each point represents one population, ellipses depict established groups and dashed ellipses their 95% confidence intervals (CI). For a better interpretation, a green triangle indicates overlapping Central European localities (BIO, LON, BRI and HOL), a black “X” indicates the Mexican and Chinese overlapping localities (CHt, NB, XYc, XYw, XG, BGt, CJr, SLt, PYL, NCyL, NHL, YNL, XT, QJ, LZL, HLL, CHL, YJ, NX, DTL, DTLs ) and a black star indicates the European, Asian and Mexican overlapping localities (BOR, VILA, MUN, ALM, WJ, CQs, HOK, FUK, CHIS).
106 Unravelling the global invasion routes of the red swamp crayfish
107 Chapter II
DISCUSSION
The high haplotype diversity of P. clarkii found in some invaded localities suggests that its global invasion, driven mostly by human-mediated introductions, may have involved admixture in the native range, an invasive bridgehead effect, and high propagule pressure. However, we also detected low levels of genetic diversity in some non-native areas (e.g. Asia), attributable to potential bottlenecks or founder effects. Our results allow the identification of the likely geographic origin and main routes of invasion, helping us to understand how the invasion has happened over a long time scale (Fig. 4).
The native range of the red swamp crayfish
Admixture has been proposed to be a causal mechanism triggering the invasiveness of some introduced species (Kolbe et al., 2007; van Boheemen et al, 2017; Fischer et al,
2017; Wagner et al., 2017; but see also, Rius & Darling, 2014) by enhancing genetic variability, thus improving population growth, decreasing the risk of extinction, and favouring adaptation to novel environments. In the present study, we found the highest haplotype diversity in the native area. The vast majority of haplotypes found in invaded areas also appeared in Louisiana but not in other native populations of US or Mexico.
This pattern is arguably related to the commercial exploitation of this species in
Louisiana, where P. clarkii has been reared, harvested and sold globally by food-industry companies for a long time (Gary, 1975; Alford et al., 2017). Although some genetic clusters seem to be differentiated between east and west Louisiana (Hap_01, Hap_04 and Hap_20 predominated in Louisiana east while Hap_03 and Hap_08 predominated in Louisiana west), Texas and Mexico localities, most of the genetic variation in those areas occurred within localities. The lack of a clear genetic structure in the native area might imply a pattern of admixture owing to farming activities in Louisiana, in which crayfish are often exchanged and translocated from wild to captive populations (Huner,
2002). Similar genetic patterns of native admixed populations have been identified in other species related to aquaculture such as the topmouth gudgeon, Pseudorasbora parva,
108 Unravelling the global invasion routes of the red swamp crayfish when they were translocated together with Chinese carp species (Hardouin et al., 2018).
The exchanges found in the Louisiana populations do not seem to have occurred in
Texas, since eight private haplotypes were found in two populations and almost all haplotypes were grouped together (see Hap_15-19 and Hap_48-50). Admixed source populations, like for P. clarkii in Louisiana, can lead to high genetic variability in invasive populations, thus allowing the invasive species to face novel environments and to thereby increase invasion success in the introduced range. But this assumption should be interpreted with caution since some species are able to show high invasiveness despite low genetic variability, such as Procambarus virginalis, a potentially highly invasive parthenogenetic crayfish that is able to establish wild populations from a single released individual (Feria & Faulkes, 2011).
The invasion dynamics of the red swamp crayfish
According to the literature, the first invasion of P. clarkii took place in Hawaiian streams in 1923 (Brasher et al., 2006) and in California in 1924 (Holmes, 1924). However, the geographical origins of both invasions remain unclear. In 1934, another event of introduction occurred in the island of Oahu, Hawaii, from Santa Barbara, California, from which subsequently P. clarkii apparently spread over the rest of the Hawaii archipelago (Penn, 1954). The Hap_27 found in each of these US states (California and
Hawaii) may confirm this second introduction event of P. clarkii from California, but this result should be treated with caution since only four crayfish were sampled from
Hawaii. In the continental USA, the California introduction was followed by later introductions to Oregon in the early 1980s (Larson & Olden, 2011) and to Washington
State in the 2000s (Mueller, 2001). Theoretically, we might expect higher genetic variability in California, with a decrease of variability from the place of first introduction
(California) northwards (Washington, Oregon) due to secondary bottlenecks and/or founder effects. Our results seem to indicate a more complex pattern of invasion (Fig. 4), in which shared haplotypes between populations in the western US confirm the connectivity among localities (Hap_01, Hap_09, Hap_20 or Hap_27). The development
109 Chapter II of the crayfish industry in California, where P. clarkii has been cultured and traded for many years, seems to have contributed to the dispersal of the crayfish along the West
Coast of the US (Comeaux, 1978; Mueller, 2001). Moreover, this scenario might have been favoured by the large number of biological supply companies in this area, and also by the use of live animals for classroom observations, some of which were given to students after school-years to take home and were probably released in the wild later on
(Larson & Olden, 2008).
Regarding the origin of California populations, we were not able to identify the precise geographic origin of this invasion because though low ɸST values were found among California and native localities, and these native localities were not close to each other (see Supplementary Material, Table S3). The haplotypes found in Topanga Creek
(TOP) suggest that the origin of this invasion came from southeastern Louisiana, but the presence of the Hap_08 (mainly distributed in western Louisiana; and considerably distinct from ancestral haplotype), would contradict this idea. For the other two
California localities (VEN and SYZ), we were also unable to unveil their origins accurately because their haplotypes were not shared with the native area; however, their private haplotypes were more evolutionarily related to Hap_04 and Hap_20, which would indicate again a possible origin from southeastern Louisiana. Given that crayfish populations in Topanga Creek were recently established (around 2001, Garcia et al.,
2015), this population could come from previous established populations in California, having undergone a possible bottleneck. If so, we could be underestimating the haplotype diversity in the area and more haplotypes would be present in California.
This latter surmise is reinforced by the fact that P. clarkii has long been considered a pest in southern California (Riegel, 1959), the higher haplotype diversity in population WAV
(Oregon) and the presence of Hap_04 at a high frequency only in PIN (Washington
State). This suggests that (1) more genetic variability is to be expected in California, acting as an admixed or bridgehead zone because of its anteriority in introduction and the numerous biological supply companies which can move live crayfish, or (2) other distinct introduction events may have occurred in northern states from the native area,
110 Unravelling the global invasion routes of the red swamp crayfish which seem unlikely given the great demand for the crayfish industry in California
(Comeaux, 1978) and since the northwest US is the native range of another commercially and culturally important crayfish species (i.e., signal crayfish, Pacifastacus leniusculus)
(Holdich, 1993).
Localities from the eastern US showed a different haplotype frequency (Hap_01 was the most common haplotype) from those of the western part of the country, suggesting another independent route of invasion to North Carolina and north of Illinois with subsequent secondary events (Fig. 4). This pattern is congruent with the native area located in the middle of the US, making it easier to move crayfish in two independent directions than from coast to coast, as well as the presence of one of the biggest biological supply companies in North Carolina (US) which was supplying most of the eastern US invaded areas. Although the eastern US is a suitable area for P. clarkii (Larson & Olden,
2011), the low haplotype diversity found in eastern localities of the US (from Hd = 0.20 in LEN to Hd = 0.50 in FOR) suggests a low propagule pressure from the native area or from shipments of the biological supply company in North Carolina.
In Asia, according to the literature (Penn, 1954), 100 specimens of P. clarkii were carried from New Orleans to Japan in 1927, of which only 20 specimens arrived alive to a pond near Tokyo (Penn, 1954; Kawai & Kobayashi, 2005); two years later, P. clarkii from Japan were translocated to Nanjing, in China (Li et al., 2007). This historical report perfectly matches the genetic pattern (i.e., founder effect and strong bottleneck) found in Japanese and, overall, Chinese populations of P. clarkii (Yue et al., 2010; Li et al., 2012;
Zhu et al., 2013, this study), in which a smaller batch was introduced to Japan and subsequent invasions came from the Japan population with few founders (Fig. 4). The lack of ectoparasites of the order Branchiobdellida is often attributed to long shipments in poor conditions (Gelder & Williams, 2015; Clavero et al., 2016). Kawai & Kobayashi
(2005) found no Branchiobdellida on Japanese specimens of P. clarkii, a pattern that could support the hypothesis that all specimens of P. clarkii in Japan (and thus, in China) descend from the initial introduction at the end of 1920s. Our results show low haplotype diversity in Japanese and Chinese populations (Hd = 0.48 and 0.35,
111 Chapter II respectively) with only four haplotypes appearing in the extensive area sampled, only two of them at high frequency (Hap_04 and Hap_28), as similarly found by Li et al.
(2012). Apart from Asian populations, Hap_28 was only found in few individuals of the
CH population (Mexico) and VEN (California) but not in the native area. Surprisingly, our genetic results seem to contradict previous literature because neither Hap_28 nor
Hap_40 appeared in localities sampled around the native locations of New Orleans,
Louisiana. The finding of Hap_28 in California and the similar date of both introductions
(i.e., California in 1924 and Japan in 1927) suggests that a route of invasion from
California to Japan is more plausible (Fig. 4). Additionally, as Hap_28 was a rare haplotype in our sampling (only 5 of the 988 non-Asian individuals carrying this haplotype), another old introduction into Asia seems unlikely because a different haplotype frequency would be expected. This strong genetic bottleneck did not prevent
P. clarkii from invading successfully (Estoup et al., 2016) and becoming a pest across
Japanese and Chinese territories (Penn, 1954; Kawai, 2017). A similar pattern has been recorded for the parthenogenetic crayfish, P. virginalis, in other areas (Feria & Faulkes,
2011). Finally, the presence of the Hap_40 in TOK (Japan) and NAT (northwest
Louisiana) led to two possible hypotheses: (1) this haplotype was present in the initial translocated batch but has been lost in subsequent secondary invasions by genetic drift or bottleneck, or (2) one new introduction event has recently occurred from the native area but has not been spread yet (nor been reported). Of both hypotheses, the first one seems more plausible, but we are not able to resolve them.
Of all invaded areas, the European invasion by P. clarkii has perhaps been the best reported, with the first two events of introduction from Louisiana to Spain (Habsburgo-
Lorena, 1978) and later into other European countries (i.e., in Spain Gutiérrez-Yurrita, et al., 1999; in France Changeux, 2003 and Laurent, 1997; in Italy Gherardi et al., 1999) (Fig.
4). The invasion routes through European countries and connectivity between European populations are poorly understood, possibly because they are due to multiple and uncontrolled deliberate introductions by private citizens (Clavero, 2016; this likely also occurred with signal crayfish, P. leniusculus, see Petrusek et al., 2017). In European
112 Unravelling the global invasion routes of the red swamp crayfish populations, we found a moderately high overall haplotype diversity (Hd = 0.58; i.e. lower than for invasive US populations, but higher than in Asia). The European invasion has probably not been based on as many introduction events as invasive American populations (e.g., California) given the differences in proximity to the native area (a possible cause of the higher haplotype diversity found on the American continent); however, the large number of P. clarkii imported to Spain (100 kg, around 6,500 crayfish) probably also included high genetic variability from the native area compared to the
Asian introduction. According to our results, a clear decrease in haplotype diversity was found from the initial sites of introduction (Hd = 0.66 in rice fields near Seville, and Hd
= 0.72 in Doñana National Park) northwards, excepted for TOS in Italy (Hd = 0.72), which could be explained by intensive farming activities on Lake Massaciuccoli (Gherardi et al., 1999) or a second introduction.
The most surprising result was the finding of two independent genetic groups in
Europe. The Hap_04 was widely distributed over the Iberian Peninsula, South France and Italy, while the Hap_11 predominated in Northern France and Italy, Belgium, the
Netherlands and United Kingdom, but was not found in the Iberian Peninsula. Two possible scenarios could explain this result: (1) we did not capture all haplotypes from the first introduction in southern Europe, and northern populations have undergone a strong bottleneck; or (2) another unreported introduction from outside Europe has occurred, independently from those reported from southern Spain (Fig. 4). The first scenario is unlikely, due to the extensive sampling effort on both the Iberian Peninsula and the native area. In such a scenario, the Hap_11 should have appeared in the Iberian
Peninsula because of other high frequencies in North Europe. Moreover, Almerão et al.
(2018) found nine haplotypes in Central France, four of which seem to match with our database but not Hap_11. On the other hand, unreported introductions of P. clarkii could be a consequence of the sales in pet shops which are common and one of the primary introduction pathways in Central Europe (Chucholl, 2015; Faulkes, 2015). These results, however, also support previous historical reports (Laurent, 1990; Holdich, 2002) suggesting how live P. clarkii may have been brought from Kenya to Europe in the 1970s.
113 Chapter II
Both hypotheses could explain the presence of this haplotype across the northern
European range of P. clarkii. To clarify our results, samples from pet shops or African samples of P. clarkii should be obtained in order to resolve the likely second invasion route. The second scenario therefore seems the most plausible (as Barbaresi et al., 2007, also suggested), with a plausible introduction from Kenya to Central Europe.
National and international translocations have occurred within Europe (Fig. 4). On the one hand, we found Hap_04 and Hap_05 to be highly frequent all over the Iberian
Peninsula to South France, which perfectly matches with the literature signaling where live specimens having been translocated from South Spain (Laurent, 1997). On the other hand, the presence of the Hap_06 at higher frequencies in southern Portugal and dated reports of introduction events across Portugal seem to confirm the spread of P. clarkii from south (near the first introduction site in 1973; Cruz & Rebelo, 2007) to north
Portugal (Gutiérrez-Yurrita et al., 1999). In addition, Hap_06 was also found in MAD
(Spain) and LAZ (Italy), suggesting a connection among these invaded areas as well as
POR in Portugal and LEZ in the Ebro Basin, Northern Spain (Fig. 4). Another possible connection was between TOS in Italy and southern Spain, with most haplotypes shared, suggesting another possible invasion route. Continuous exchanges and secondary translocations of P. clarkii through invaded areas have produced a very complex invasion process, which could accelerate the invasiveness of this kind of species
(Wagner et al., 2017).
Our results provide a clear example of how different features of introduction events and invasion processes (e.g. genetic admixture, propagule pressure or secondary introductions) can generate contrasting genetic diversity patterns across non-native populations of a global invader (Roman & Darling, 2007). For example, Asian populations of P. clarkii underwent a strong bottleneck as a consequence of the introduction of few individuals in a single introduction event, which arguably originated from an already introduced population (probably in western US) that might have already gone through previous bottlenecks. Genetic diversity was notably higher in the P. clarkii populations in western US, probably due to the existence of numerous
114 Unravelling the global invasion routes of the red swamp crayfish introduction events (e.g. facilitated by vicinity to the native range and the development of biological supply companies), involving large batches of individuals with high genetic admixture. The European case is apparently intermediate, with numerous individuals imported from an admixed native range to SW Spain, from which the species expanded across the continent through multiple secondary introductions, involving a clear loss in haplotype diversity. However, higher genetic variabilities were found in European (Petrusek et al., 2017) and Japanese populations of P. leniusculus
(Usio et al., 2016) in comparison to that reported by us for P. clarkii, arguably due to the combination of several introduction events involving large batches of individuals and coming from a variety of origins in the US, including native and non-native populations.
A striking pattern deriving from our results is that the invasiveness of P. clarkii does not seem to depend, at least in the short- and mid-term, on the genetic diversity of introduced populations. Although genetic diversity can fuel invasiveness by allowing the efficient adaptation of introduced populations to spatial and temporal variability in the recipient ecosystems, the relationship between those two features is obscure (Estoup et al., 2016). There is growing evidence that the loss of genetic diversity in introduced populations can be compensated through epigenetic processes (Estoup et al., 2016). The most extreme example of high invasiveness with low genetic variability is the clonal species P. virginalis (Feria & Faulkes, 2011), which is able to thrive in a wide variability of environmental conditions (Andriantsoa et al., 2019).
Apart from informing about invasion routes, our results might also be relevant for new approaches for the detection and surveillance of invasive species. Environmental
DNA (eDNA) is a rapidly emerging monitoring tool for freshwater invasive species based on the persistence of DNA fragments in the environment (Ficetola et al., 2008;
Mauvisseau et al., 2018). Large-scale phylogeographic studies provide accurate datasets for improving invasive species detection protocols based on eDNA (Ficetola et al., 2008;
Larson et al., 2017). Admixture in both native and invasive ranges, as well as the bridgehead invasive effect, has led to large intraspecific genetic variability within and among invaded areas, which may reduce the efficacy of eDNA protocols (Wilcox et al.,
115 Chapter II
2015). In fact, the spatial gradients in genetic variability and the presence of different genetic clusters in Europe reported here, probably led to the failure of eDNA probes in detecting French populations of P. clarkii (Tréguier et al., 2014; Mauvisseau et al., 2018), which had worked well with the less variable Chinese populations (Cai et al., 2017). Our study may thus be useful for the development of better site-specific eDNA-based protocols to detect P. clarkii (Manfrin et al., 2019).
CONCLUSIONS
Our results illustrate extensive admixture of P. clarkii in its native area, report two independent invasion routes in the US (i.e., westwards and eastwards), and support the historical reports of a single introduction event into Asia involving few individuals.
They also suggest that Europe may have received P. clarkii through more introduction routes than the frequently reported imports into Spain. To find other likely introduction routes, more effort should be put on sampling in previously unstudied sites (e.g., Texas, pet shops, biological supply trade and/or Southern Hemisphere countries where other introduced populations might act as sources of invasion, for example, African or South
American populations).
We have traced the complex scheme of invasion of P. clarkii (Fig. 4), with a key role for human-mediated dispersal. The economic value of P. clarkii and the ease with which it is transported have favoured the spread of the species worldwide (largely for aquaculture, the aquarium trade and other forms of human exploitation as food) as the consequence of multiple subsequent introduction events. Genetic admixture, invasive bridgehead effects, extensive genetic variation in the native area and high propagule pressure are apparent drivers of genetic variability across its broad geographic distribution. Such extensive genetic variability in invaded areas should be taken into account to improve management measures based on mtDNA for environmental detection of this invasive species. Overall, invasive species, and invasive crayfish in particular, continue to be artificially introduced into more countries through the aquarium trade (e.g. fish species, Strecker et al., 2011; crayfish species, Patoka et al., 2014
116 Unravelling the global invasion routes of the red swamp crayfish and particularly P. virginalis, Faulkes 2015). The example of the successful worldwide invasion of P. clarkii highlights the high spread potential of intentionally introduced freshwater species, especially those species also involved in aquaculture (Naylor et al.,
2001). Once a species has been introduced in a new territory, management strategies aimed at reducing the spread and impacts of invasive species should focus on avoiding secondary introductions and would benefit from the early detection of potential invasion hubs.
ACKNOWLEDGEMENTS
We thank R. López, M. A. Bravo, and Á. Vidal for field and lab assistance, as well as the many colleagues who kindly provided samples from Spain [J. Molero, J. L. Grau,
E. Carnerero, A. Pradilla (CCEDCV - EL PALMAR), J. Rucio & M. Alcántara (Aragón
Government), L.M. Álvarez (Asturias Government), Parque Martioda & J. Carreras
(Alava Provincial Council), J. L. Moreno, N. Franch (Parc Natural del Delta de l'Ebre),
Centro de Acuicultura Vegas del Guadiana (Extremadura Government), L. Valls, F. J.
Galindo (Andalusia Government), F. Unzu & L. Lopo (La Rioja Government), J. Bosch,
Presi, Kike, F. J. Oliva, I. Vedia, A. Mestre and F. Mesquita], Europe: [F. Ribeiro, A. F.
Filipe, I. Roessink, D. Ghia, E. Tricarico, R. Cammaerts, P. Fossat, D. Nicolas, E. Petit, O.
Adrien and J. Cucherousset] and the United States [R. Fagundo & B. Williams (North
Carolina Museum of Natural Sciences), T. García, J. Olden & L. Kuehne, E. R. Larson, C.
Gabor and R. Keller]. We also thank both Laboratorio de Ecología Molecular (LEM-EBD) and Laboratorio de SIG y Teledetección (LAST-EBD) at Estación Biológica de Doñana
(CSIC) for providing logistical support. In addition, we thank Eric R. Larson and another anonymous reviewer who improved the final version of this manuscript. F.J.O. was supported by an Andalusia Government grant. This study was funded by the
Andalusian Government (RNM-936) and the bilateral CNRS-CSIC PICS programs
(PICS07360 and PIC2015FR4).
117 Chapter II
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SUPPLEMENTARY MATERIAL
List of legends of supplementary material
Table S1. List of haplotypes used for this study and their synonyms for the 608 bp COI gene fragment used. GenBank accession numbers in bold were obtained from Almerão et al. (2018), and those in italics were obtained from (Li et al., 2012). A same GenBank accession number can appear in more than one of our haplotypes due to different number of base pairs in the alignment. * indicates that the other study found the haplotype in this zone but was not detected in our study.
Haplotype GenBank References Native Area Asia West America East US Europe Hap_01 X X X X Hap_02 X JN000899 Torres & Álvarez 2012 JN000905 Hap_03 X X X X JX120108 Li et al. 2012 KY112716 Almerão et al. 2018 AY701195 Taylor & Knouft 2006 JF438002 Filipová et al. 2011 Hap_04 JN000901 Torres & Álvarez 2012 X X X X X JX120103 Li et al. 2012 KY112709 Almerão et al. 2018 Hap_05 KY112710 Almerão et al. 2018 X X Hap_06 X Hap_07 X Hap_08 JX120107 Li et al. 2012 X X JF438004 Filipová et al. 2011 JN000904 Torres & Álvarez 2012 Hap_09 X X X X JX120105; Li et al. 2012 KY112711 Almerão et al. 2018 Hap_10 X Hap_11 JF438003 Filipová et al. 2011 X Hap_12 X Hap_13 KY112711 Almerão et al. 2018 * X Hap_14 X Hap_15 X Hap_16 X Hap_17 X Hap_18 X Hap_19 X JF438001 Filipová et al. 2011 Hap_20 X X * KY112712 Almerão et al. 2018
126 Unravelling the global invasion routes of the red swamp crayfish
Table S1. (Cont.) Haplotype GenBank References Native Area Asia West America East US Europe Hap_21 X Hap_22 X Hap_23 X Hap_24 X Hap_25 X Hap_26 X Hap_27 X JN000903 Torres & Álvarez 2012 Hap_28 X X JX120104 Li et al. 2012 Hap_29 X X Hap_30 X Hap_31 X Hap_32 X Hap_33 X Hap_34 KY112709 Almerão et al. 2018 X * Hap_35 X Hap_36 X Hap_37 X Hap_38 X Hap_39 X Hap_40 X X Hap_41 X Hap_42 X Hap_43 X Hap_44 KY112712 Almerão et al. 2018 X * Hap_45 X Hap_46 X Hap_47 X Hap_48 X Hap_49 X Hap_50 X Hap_51 X Hap_52 X Hap_53 X Hap_54 X Hap_55 X Hap_56 X Hap_57 X Hap_58 X Hap_59 X
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Table S1. (Cont.) Haplotype GenBank References Native Area Asia West America East US Europe Hap_60 JN000898 Torres & Álvarez 2012 X JN000900 Hap_61 Torres & Álvarez 2012 X JN000906 Hap_62 JN000902 Torres & Álvarez 2012 X Hap_63 JN000907 Torres & Álvarez 2012 X Hap_64 JN000908 Torres & Álvarez 2012 X Hap_65 JX120106 Li et al. 2012 X KY112713 Almerão et al. 2018 X KY112714 Almerão et al. 2018 X KY112715 Almerão et al. 2018 X KY112717 Almerão et al. 2018 X
128 Unravelling the global invasion routes of the red swamp crayfish
Table S2. Haplotype frequencies grouped by the 5 biogeographical zones. Bold type indicates haplotypes shared between two localities at least.
Haplotype Native Area Asia West Americas East US Europe Hap_01 18 6 44 2 Hap_02 1 Hap_03 7 2 2 12 Hap_04 35 144 17 3 419 Hap_05 2 69 Hap_06 54 Hap_07 1 Hap_08 7 6 Hap_09 35 4 1 3 Hap_10 9 Hap_11 91 Hap_12 2 Hap_13 1 Hap_14 1 Hap_15 1 Hap_16 2 Hap_17 2 Hap_18 2 Hap_19 1 Hap_20 11 19 Hap_21 1 Hap_22 1 Hap_23 1 Hap_24 5 Hap_25 1 Hap_26 1 Hap_27 2 Hap_28 269 5 Hap_29 1 3 Hap_30 1 Hap_31 1 Hap_32 1
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Table S2. (Cont.) Haplotype Native Area Asia West Americas East US Europe Hap_33 1 Hap_34 1 Hap_35 1 Hap_36 1 Hap_37 1 Hap_38 1 Hap_39 1 Hap_40 2 1 Hap_41 1 Hap_42 1 Hap_43 1 Hap_44 1 Hap_45 1 Hap_46 1 Hap_47 2 Hap_48 3 Hap_49 2 Hap_50 9 Hap_51 1 Hap_52 1 Hap_53 1 Hap_54 1 Hap_55 1 Hap_56 2 Hap_57 1 Hap_58 1 Hap_59 1 Hap_60 3 Hap_61 9 Hap_62 1 Hap_63 3 Hap_64 1 Hap_65 1
130 Unravelling the global invasion routes of the red swamp crayfish
Table S3 Pairwise ɸST values (below diagonal) and pairwise Jost’s Dest values with confidence intervals (above diagonal) for the mtDNA COI of Procambarus clarkii. Both significant p- values of
ɸST following a FDR correction and confidence intervals of Dest not enclosing 0, are represented in bold. Locality names are abbreviated as in Table 1.
Please download this table in the following link: https://drive.google.com/open?id=144DnzQ_oaBnU4Xhds16x1CrZi90UUMPI
Figure S1. TCS haplotype network for cytochrome c oxidase subunit I (COI) sequences of the red swamp crayfish, Procambarus clarkii, in the native range. Each circle represents one haplotype and its size is proportional to the haplotype frequency. Within the network, each line between haplotypes represents a mutational change and small black dots show unsampled haplotypes inferred from the data. Localities from the same location share the same colour.
131
CHAPTER III
Brought more than twice: the complex introduction history of the
red swamp crayfish into Europe
Oficialdegui FJ, Sánchez MI, Lejeusne C, Pacini N, Clavero M
Submitted to Biological Invasions
Chapter III
ABSTRACT
One of the biggest challenges in understanding and managing biological invasions is the identification of the routes of introduction. This information is often incomplete because of unnoticed, unreported and, sometimes, illegal translocations. Reports on the introduction of the red swamp crayfish (Procambarus clarkii) into Europe specify it was introduced for the first time to southern Spain (1973 and 1974) from Louisiana; from there, it rapidly spread throughout several European countries. While other importation events have been suggested in the literature, there is no evidence that these led to wild populations in Europe. Our present study unambiguously confirms two introduction routes into Europe from non-European areas previously invaded. By using mtDNA, we found evidence of shared haplotypes between the Lake Naivasha in Kenya and central
Europe, as well as between the western United States or Asia, and Malta. These findings strongly support historical reports found in the literature and contribute to explain the high genetic variability of the red swamp crayfish populations in Europe, shedding new insights into this complex global invasion.
134 Introduction events of the red swamp crayfish into Europe
INTRODUCTION
Identifying introduction routes is important to understand, predict and manage biological invasions (Estoup and Guillemaud 2010; Cristescu 2015). This identification may be hampered because introduction events are often unreported, incompletely reported (e.g. information on the origin of introduced populations is lacking) or reported in inaccessible sources (e.g. administrative documents). Molecular tools offer opportunities to reconstruct unreported invasion routes by comparing the genetic profiles of native and introduced populations (Estoup and Guillemaud 2010; Fitzpatrick et al. 2012).
The red swamp crayfish Procambarus clarkii (Girard 1852), native to southern US and northern Mexico, has been purposely introduced to all continents except Australia and Antarctica to be bred and consumed as food (Hobbs et al. 1989) or to be sold as an aquarium species (Faulkes 2015). It is currently considered the most cosmopolitan freshwater crayfish in the world, as well as one of the most ecologically destructive invaders (Gherardi and Acquistapace 2007). It was first brought to Europe from
Louisiana (US) in 1973 when it was introduced to the rice fields of the middle Guadiana
Basin, Badajoz province in Spain, and again in 1974 introduced to the rice fields of the lower Guadalquivir, Seville province (Habsburgo-Lorena 1978). It has generally been assumed that the European range of the red swamp crayfish, currently distributed throughout 12 countries, results from the expansion of these two founder populations, driven by multiple, uncontrolled translocations (Kouba et al. 2014 and references therein). This assumption seems to be supported by the high haplotype diversity that the red swamp crayfish populations show in SW Spain, and by the observation that it progressively decreases as one moves towards the North, as recently reported by
Oficialdegui et al. (2019). However, one haplotype was found exclusively across western
Europe (NW France, northern Italy, Belgium, the Netherlands and United Kingdom), which was not detected in the Iberian Peninsula, and neither in southern France or in southern Italy (Oficialdegui et al. 2019). This finding was interpreted as the probable result of an unreported introduction into Europe, in addition to those that occurred in
135 Chapter III
Spain in the 1970s. The possibility of other plausible introduction routes into Europe has been suggested in the literature, but never explicitly supported. For example, Barbaresi et al. (2007) proposed that red swamp crayfish could have been brought to Italy from
China, while Laurent et al. (1991) and Goddard & Hogger (1986) suggested that it could have been introduced to France and to the United Kingdom from Kenya for human consumption.
Here, we investigate whether the invasion of Europe by the red swamp crayfish may have involved routes different from the initial inoculums from Louisiana to SW
Spain in the 1970s. We focussed specifically on possible routes originating from Africa and/or Asia. To that aim, we: 1) analysed the mitochondrial DNA (mtDNA) of red swamp crayfish from Lake Naivasha in Kenya; and 2) compiled and re-analysed recently published sequences from red swamp crayfish populations, both in Europe and elsewhere.
METHODS
We collected a total of nine red swamp crayfish from Lake Naivasha (Kamere
Landing Beach, 0°48.88’ S; 36°19.46’ E), Kenya. Muscle tissue was extracted on site and individually preserved in 96% ethanol until subsequent genetic analyses. Extraction of mtDNA, use of cytochrome oxidase 1 (COI) primers and sequencing protocols were performed as in Oficialdegui et al. (2019). We also re-assessed 10 sequences (559 bp;
Genbank Accession number: MF170527-36) recently published by Vella et al. (2017), referring to populations established in Malta, as well as the 65 haplotypes (608bp) reported by Oficialdegui et al. (2019). A total number of 84 sequences was thus analysed.
After alignment and trimming of all sequences (Kenyan and Maltese), a matrix of 559 bp was obtained.
RESULTS
The nine red swamp crayfish specimens from Kenya shared a common haplotype, while a different haplotype was common to the 10 specimens from Malta reported by
136 Introduction events of the red swamp crayfish into Europe
Vella et al (2017). The Kenyan haplotype was identical to the one found in western
Europe but was not present anywhere else (Hap_11, GenBank: MK026681 in
Oficialdegui et al. 2019). Because of the trimming of sequences, the Maltese haplotype corresponds to two haplotypes (Hap_28, MK026698 and Hap_41, MK026711; in
Oficialdegui et al. 2019) occurring mainly in Asia, with Hap_28 being much more widespread than Hap_41. Interestingly, Hap_28 is also present in western North
America, where the red swamp crayfish is also an invasive species (Oficialdegui et al.
2019) (Fig. 1).
137 Chapter III
Figure 1. Haplotype distribution and plausible invasion routes of P. clarkii from non-European areas into Europe (modified from Fig. 1 in Oficialdegui et al. 2019). (A) Introduction routes are depicted by continuous (plausible routes) and discontinuous red arrows (hypothetical route), the dark grey arrow indicates the known introductions from Louisiana in 1970s, and the yellow and blue shading depict the area where Hap_28 and Hap_11 are distributed; (B) Hap_11, in Kenyan population; (C) Hap_28, main haplotype in Asian distribution and Hap_41 is the private haplotype in OKA population; (D) European haplotype distribution with presence of Hap_11 and Hap_28 in Malta (more likely, see discussion). * indicates shared haplotypes between non- European and European areas showed in this study.
DISCUSSION
The findings of shared haplotypes between Kenya and north-western Europe, as well as between Asia (or western North America) and Malta, show that the well- documented introductions into Spain of the 1970s were not actually the only ones which developed wild populations in Europe. We thus demonstrate that the invasion process of the red swamp crayfish in Europe, and plausibly elsewhere, is more complex than generally assumed.
The red swamp crayfish was introduced from the US to several African countries, including Uganda, Kenya, Zambia, Sudan, South Africa (Huner 1977; Mikkola 1996),
Egypt (Hamdi 1994), and recently Morocco from Spain (El Qoraychy et al. 2015). It was particularly successful in Lake Naivasha (Kenya), where a large population was established from only 300 individuals introduced in 1966 from a population that had been previously introduced from Uganda. This successful population in Kenya rapidly reached extremely high densities (Harper et al. 2002). Commercial exploitation was initiated in the early 1970s, with exports starting from 1975 until sudden stopping in
1983 (Harper et al. 2003), when the importation of live crayfish was banned by most
European countries (Laurent et al. 1991). According to Laurent et al. (1991), tonnes of live crayfish were imported to France from Kenya, as well as from Spain, between 1975 and 1983. Goddard and Hogger (1986) reported that red swamp crayfish from Kenya were also imported to England for the wholesale fish market, although they thought that there were no populations established in the wild by the mid-1980s. In any case, our
138 Introduction events of the red swamp crayfish into Europe results strongly suggest that imports originating from Kenya have given rise to self- sustaining populations in north-western Europe.
The presence in Malta of one haplotype shared by several red swamp crayfish populations in Asia and in the western US suggests another unreported introduction route into Europe. Barbaresi et al. (2007) suggested that red swamp crayfish could have been imported to Italy (specifically to Tuscany) from Asia and released into the wild. If such introduction had taken place, Italy would be a good candidate as the origin of the red swamp crayfish introduced to Malta, acting as a stepping-stone between China and the final destination. However, Oficialdegui et al. (2019) found neither the Maltese haplotype nor any other one related to Asia in Italy, despite describing high haplotype diversity in the country, especially in Tuscany. This pattern does not support the stepping-stone role of Italy, but it neither rules it out. In this sense, further analysis for the genetic variability of southern Italian populations would be needed, where the species is known to be present since at least 2002 (Deidum et al. 2018).
On the basis of the genetic variability of the red swamp crayfish recently reported
(Oficialdegui et al. 2019), we propose that the Maltese population could have originated from releases of live specimens sold by biological supply companies or by direct imports of live crayfish from Asia, possibly from China. Crayfish movements mediated by biological supply companies are known to have generated self-sustaining wild populations (Faulkes 2015). Several biological supply companies are located in the western US (California), where the Maltese haplotype is known to be present, and live red swamp crayfish could be readily obtained. However, the red swamp crayfish could have also been imported directly from China to Malta. The two countries have long- lasting diplomatic relationships (dating back to 1972) and commercial, cultural and scientific exchanges have recently been facilitated by a Memorandum of Understanding, signed in 2014 (Camilleri, 2019). One of the main cooperation lines signalled in that agreement was aquaculture, a relevance that resulted in the promotion of a Malta-China
Joint Research Laboratory in Aquaculture (Parliament of Malta, 2016). The very recent origin of the Maltese population, first detected in 2016 (Vella et al. 2017), suggests that
139 Chapter III these initiatives may have facilitated the arrival of the invasive crayfish. But, since the red swamp crayfish has been detected in the country together with other invasive decapods widely used by aquarists (e.g., Procambarus virginalis or Atyopsis moluccensis)
(Deidum et al. 2018), the aquarium trade could also have been the actual means of introduction of the red swamp crayfish into Malta; the genetic identity of this population coupled with Maltese-Chinese relationships described above would suggest an Asian origin, this being a relevant information for the prevention of new introductions.
This study provides evidence that the account of red swamp crayfish expanding across Europe from introductions in SW Spain from Louisiana is a simplification of a considerably more complex invasion process. Our findings have implications for the understanding of the invasion by the red swamp crayfish, one of the most harmful invasive species in Europe (Nentwig et al. 2018), and poses questions on how genetic admixture resulting from introductions from multiple sources could influence the genetic variability of European populations and their ability to become active invaders
(Wagner et al. 2017; but see Rius and Darling 2014, Estoup et al. 2016).
ACKNOWLEDGEMENTS
We thank the Laboratorio de SIG y Teledetección (LAST-EBD) at Estación Biológica de Doñana (CSIC) for providing logistical support. F.J.O. was supported by an
Andalusia Government grant. This study was funded by the Andalusia Government
(RNM-936) and PICS (PIC2015FR4 and PICS07360).
REFERENCES Barbaresi S, Gherardi F, Mengoni A, Souty-Grosset C (2007) Genetics and invasion biology in fresh waters: a pilot study of Procambarus clarkii in Europe. In Biological invaders in inland waters: Profiles, distribution, and threats (pp. 381-400). Springer, Dordrecht. Camilleri J (2019) Malta's commerce agreement with China. The official business portal of the Malta Chamber of Commerce, Enterprise and Industry. Available at: https://www.maltachamber.org.mt/en/blogs/155 (accessed February 15th 2019). Cristescu ME (2015) Genetic reconstructions of invasion history. Molecular ecology 24(9):2212- 2225.
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Deidun A, Sciberras A, Formosa J, Zava B, Insacco G, Corsini-Foka M, Crandall KA (2018) Invasion by non-indigenous freshwater decapods of Malta and Sicily, central Mediterranean Sea. Journal of Crustacean Biology 38(6):748-753. El Qoraychy I, Fekhaoui M, El Abidi A, Yahyaoui A (2015) Heavy metals in Procambarus clarkii of Rharb Region in Morocco and Its Safety for Human consumption. Journal of Environmental Science, Toxicology and Food Technology 9 (10):38-43. Estoup A, Guillemaud T (2010) Reconstructing routes of invasion using genetic data: why, how and so what?. Molecular ecology 19(19): 4113-4130. Estoup A, Ravigné V, Hufbauer R, Vitalis R, Gautier M, Facon B (2016) Is there a genetic paradox of biological invasion?. Annual Review of Ecology, Evolution, and Systematics 47:51-72. Faulkes, Z. (2015). The global trade in crayfish as pets. Crustacean Research 44:75-92. Fitzpatrick BM, Fordyce JA, Niemiller ML, Reynolds RG (2012) What can DNA tell us about biological invasions?. Biological Invasions 14(2):245-253. Gherardi F, Acquistapace P (2007) Invasive crayfish in Europe: the impact of Procambarus clarkii on the littoral community of a Mediterranean lake. Freshwater Biology 52(7): 1249-1259. Goddard JS, Hogger JB (1986) The current status and distribution of freshwater crayfish in Britain. Field studies 6(3):383-396. Habsburgo-Lorena AS (1978) Present situation of exotic species of crayfish introduced into Spanish continental waters. Freshwater Crayfish 4:175-184. Hamdi SAH (1994). Studies on the red swamp crawfish Procambarus clarkii (Girard, 1852) (Decapoda: Camaridae) in the River Nile, Egypt. M. Sc..Thesis, Fac. Sci., Cairo Univ., 136 pp. Harper DM, Smart AC, Coley S, et al. (2002) Distribution and abundance of the Louisiana red swamp crayfish Procambarus clarkii Girard at Lake Naivasha, Kenya between 1987 and 1999. In Lake Naivasha, Kenya (pp. 143-151). Springer, Dordrecht. Hobbs H, Jass J, Huner JA (1989) A review of global crayfish introductions with particular emphasis on two North American species. Crustaceana 56:299-316. Huner JV (1977) Introductions of the Louisiana Red Swamp Crayfish, Procambarus clarkii (Girard); an update. Freshwater crayfish 3:193-202. Laurent PJ, Leloirn H, Neveu A (1991) Remarques sur l'acclimatation en France de Procambarus clarkii (Decapoda Cambaridae). Publications de la Société Linnéenne de Lyon 60(5):166-173. Lowery RS, Mendes AJ (1977) Procambarus clarkii in Lake Naivasha, Kenya, and its effects on established and potential fisheries. Aquaculture 11(2):111-121. Mikkola H (1996) Alien freshwater crustacean and indigenous mollusc species with aquaculture potential in Eastern and Southern Africa. Southern African Journal of Aquatic Sciences 22:90- 99. Nentwig W, Bacher S, Kumschick S, Pyšek P, Vilà M (2018) More than “100 worst” alien species in Europe. Biological invasions 20(6):1611-1621. Oficialdegui FJ, Clavero M, Sanchez MI et al. (2019) Unravelling the global invasion routes of a worldwide invader, the red swamp crayfish (Procambarus clarkii). Freshwater Biology, in press. DOI: 10.111/fwb.13312 Parliament of Malta (2016). Annual Report and Financial Statements 2015. Available at: https://www.parlament.mt/media/88319/08227.pdf (accessed February 15th 2019)
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Rius M, Darling JA (2014) How important is intraspecific genetic admixture to the success of colonising populations?. Trends in ecology & evolution 29(4):233-242. Vella N, Vella A, Mifsud CM (2017) First Scientific Records of the Invasive Red Swamp Crayfish, Procambarus clarkii (Girard, 1852)(Crustacea: Cambaridae) in Malta, a Threat to Fragile Freshwater Habitats. Natural and Engineering Sciences 2(2):58-66. Wagner NK, Ochocki BM, Crawford KM, Compagnoni A, Miller TE (2017) Genetic mixture of multiple source populations accelerates invasive range expansion. Journal of Animal Ecology 86(1):21-34.
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The invasive red swamp crayfish (Procambarus clarkii) increases infection of 26
the amphibian chytrid fungus (Batrachochytrium dendrobatidis) 27
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ABSTRACT 32 Emerging infectious diseases are increasingly recognized as a severe threat to 33 wildlife. Chytridiomycosis, caused by Batrachochytrium dendrobatidis (Bd), is considered 34 one of the most important causes for the decline of amphibian populations worldwide. 35
Identifying potential biological reservoirs and characterizing the role they can play in 36 pathogen maintenance is not only important from a scientific point of view, but also 37 relevant from an applied perspective (e.g. disease control strategies), especially when 38 worldwide distributed invasive species are involved. We aimed (1) to analyse the 39 prevalence and infection intensity of Bd in the invasive red swamp crayfish (Procambarus 40 clarkii) across western Andalusian region, Spain; and (2) to assess whether the presence 41 of crayfish affects the prevalence and infection intensity of Bd in amphibians of Doñana 42
Natural Space (DNS), a highly protected area within Andalusian region. First, we found 43 that infection prevalence in crayfish guts was 1.5% regionally (four out of 267 crayfish 44 were qPCR positive to Bd, all of them belonging to the same Andalusian population); 45 qPCR positives were histologically confirmed by finding zoosporangia of Bd in 46 gastrointestinal walls of the red swamp crayfish. Second, we found a higher prevalence 47 of Bd infection in the localized protected region, DNS (19% for crayfish and 28% for 48 amphibians on average), a place with great diversity and abundance of amphibians. Our 49 analyses showed that prevalence of Bd in amphibians was related to the presence of the 50 red swamp crayfish, indicating that this crayfish can be a suitable predictor of Bd 51 infection in co-occurring amphibians. These results suggest that the red swamp crayfish 52 might be a possible reservoir for Bd, representing an additional indirect impact on 53 amphibians, a role that had not been previously recognised in its invasive range. 54
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INTRODUCTION 61
Globalized infectious diseases, particularly fungal pathogens, are one of the 62 greatest threats to wildlife (Fisher et al. 2012). Amphibians are the most threatened 63 vertebrates globally, being highly susceptible to climatic change, habitat fragmentation 64 and overexploitation (Stuart et al. 2004). Moreover, they are affected by the most 65 devastating panzootic so far, amphibian chytridiomycosis, which is closely related to 66 amphibian declines (Voyles et al. 2009; GISD 2018; Scheele et al. 2019). The amphibian- 67 killing pathogen, Batrachochytrium dendrobatidis (hereafter “Bd”), infects keratinizing 68 epithelial cells in amphibian skin (Berger et al. 1998; van Rooij et al. 2015), causing 69 osmotic damage and often death (Daszak et al. 2003; Voyles et al. 2009). Bd is considered 70 as one of the 100 worst invasive species worldwide (GISD 2018), whose origin has been 71 recently attributed to Asia, and the pathways of its spread at the global scale are mainly 72 linked to amphibian pet trade (O’Hanlon et al. 2018). Global trade is also resulting in the 73 evolution of further hypervirulent fungal lineages across a diverse range of host species 74 and biomes (Farrer et al. 2011). 75
Because pathogenicity of Bd is temperature-dependent, prevalence of Bd is usually 76 higher in early spring and lower in late summer or autumn (Kriger and Hero 2007), 77 which affects the temporal patterns of exposure (Walker et al. 2010; Xie et al. 2016). The 78 highest rates of Bd-infection often coincide with the breeding season of many amphibian 79 species during winter, making them more vulnerable (Gervasi et al. 2017). However, 80 viability of the Bd pathogen decreases through the seasons in areas where water 81 temperature rises over 30ºC and waterbodies dry up during the dry season (Piotrowski 82 et al. 2004, Doddington et al. 2013). Thus, amphibians with a highly variable response to 83 different levels of infection can be reservoirs of Bd, maintaining this pathogen in the 84 environment (Woolhouse et al. 2001; Gervasi et al. 2017; Brannelly et al. 2018). Others 85 have hypothesized on the saprophytic feeding of Bd, which is able to live without an 86 amphibian host (Speare et al. 2001). In addition, in spite of being an amphibian fungal 87 disease, there is increasing interest in evaluating the potential role of non-amphibian 88 species, which co-occur in aquatic ecosystems, as reservoirs or carriers of 89
147 Chapter IV chytridiomycosis (van Rooij et al. 2015). Identification of potential reservoirs and 90 characterization of the role they play in Bd maintenance and disease dynamics is not 91 only important from an evolutionary perspective, but can also provide important 92 insights on disease control strategies. 93
The red swamp crayfish (Procambarus clarkii), native to north-eastern Mexico and 94 south-central US, has been intentionally introduced worldwide, becoming the most 95 widespread crayfish in the world (Oficialdegui et al. 2019). Over the last 45 years, 96 southern European freshwaters have been widely invaded (Kouba et al. 2014), causing, 97 among others, dramatic declines of many amphibian populations and negatively 98 impacting their community structure (Ficetola et al. 2012), with an important decrease 99 in amphibian richness (Cruz et al. 2006). The recent finding that the red swamp crayfish 100 and other crayfish species can become infected by Bd in their native range (Louisiana, 101
US), and potentially transmit infection to amphibians (McMahon et al. 2013, Brannelly 102 et al. 2015a, but see Brannelly et al. 2015b), suggests the need for evaluating the role of 103 invasive crayfish as carriers of the amphibian fungus in its invaded range. This is 104 especially important in places where crayfish species are particularly prolific and 105 colonize a wide range of aquatic environments where amphibians occur (GISD 2018). 106
The red swamp crayfish is especially abundant in the southern part of the Iberian 107
Peninsula, including the Doñana Natural Space (DNS), a protected area located in 108 western Andalusia where the conservation value of ponds for amphibians and other 109 aquatic fauna and flora has been highlighted by many studies (Gómez-Rodríguez 2007; 110
Díaz-Paniagua et al. 2010). Due to its commercial value, humans have intentionally 111 translocated the red swamp crayfish among water bodies, promoting its rapid 112 colonization across Europe (Oficialdegui et al. 2019). Thus, studies evaluating the role 113 of the red swamp crayfish as potential reservoir in this area are of particular interest, as 114 crayfish may rapidly disperse Bd over long distances and contribute to the global Bd 115 pandemic in Europe. We conducted a study to explore these questions that comprised 116 two sections: (1) a regional survey across western Andalusia, with the goal of assessing 117 the prevalence of infection and infection intensity of Bd in the red swamp crayfish in 118
148 The invasive red swamp crayfish increases infection of the amphibian chytrid fungus
Andalusian streams and ponds (crayfish infection survey); and (2) a specific sampling 119 in a localized protected area within Andalusia region (DNS), a particularly rich area in 120 amphibian populations where Bd had been circulating through the environment 121
(Hidalgo et al. 2012), with the aim of determining whether prevalence of infection and 122 infection intensity of Bd in amphibians was positively related to the presence of the red 123 swamp crayfish in ponds in DNS (crayfish/amphibian interaction). 124
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Study area 127
In the crayfish infection survey, our sampling locations were located in 11 streams 128 of the provinces of Seville, Cadiz and Huelva, in western Andalusia (South-western 129
Spain) and three ponds within Doñana Natural Space (DNS) (Fig. 1A, see Table 1). The 130 region is dominated by the Guadalquivir basin and characterized by a Mediterranean 131 climate, with a long dry summer season, mean annual temperatures between <10°C and 132
18°C, and mean altitude of 200 m asl. Most streams in this Mediterranean region are 133 intermittent, where streams have little flow during wet season water remains in puddles 134 along the stream during dry season. Subsequently, for the study of crayfish/amphibian 135 interaction, we sampled six different ponds in the Doñana Natural Space (DNS), a 136 protected area within western Andalusia which is particularly important for amphibian 137 breeding and conservation, in order to analyse the prevalence and infection intensity of 138
Bd in relation to the presence of the red swamp crayfish (Fig. 1B). The DNS, situated in 139 the mouth of the Guadalquivir River, comprises the Doñana Natural Park and the 140
Doñana National Park, and is considered one of the largest and most important 141 wetlands of Europe. It has been declared a UNESCO Biosphere Reserve, a Ramsar site, 142 a Natural World Heritage Site, and part of the Natura 2000 network (see e.g. García- 143
Novo and Marín-Cabrera 2005). As other very scarce Mediterranean temporary ponds, 144 this system of more than 3,000 water bodies that integrate DNS is considered a priority 145 habitat under the European Union Habitats Directive (Code 3170: European 146
Commission DG Environment, 2007). 147
149 Chapter IV
148 Figure 1. (A) Map of the study area showing crayfish sampling points in streams of western 149 Andalusia (“HUE” Huelva, “SEV” Seville and “CAD” Cadiz). Black spots indicate locations with 150 Bd- and the black triangle shows the only Bd+ location. Area depicted by a square indicates the 151 protected area of Doñana Natural Space (DNS) and greys represent the Doñana Natural Park 152 (light grey) and Doñana National Park (dark grey). (B) Protected areas of Doñana Natural Space 153 (Doñana Natural Park and Doñana National Park). Sampling sites with crayfish (black squares) 154 and ponds without crayfish (black triangles) are numerated as follows: (1) ANS, (2) OLL, (3) PAJ, 155 (4) TAR, (5) ZAH, and (6) PDR. Dark grey patches depict temporary ponds and grey circles depict 156 permanent water bodies. 157 150 The invasive red swamp crayfish increases infection of the amphibian chytrid fungus
Table 1. Procambarus clarkii sampling in 14 streams of western Andalusia, Spain, in 2015. For each 158 location we show the coordinates, date of sampling, altitude (m), number of crayfish captured (n), 159 type of habitat, prevalence of Bd (%) with 95% confidence intervals in parentheses, and infection 160 intensity (Bd load in Zoospore Equivalents, mean ± SE). 161 162 163 Code Location Lat Long Altitude n Type of habitat Prevalence (CI) Bd load GDP Guadalporcún 36.948 -5.358 282 20 Stream 0 - GUA Guadiamar 37.657 -6.229 203 20 Stream 0 - HUE Hueznar 37.933 -5.697 422 20 Stream 0 - MAN Manecorro 37.124 -6.489 6 25 Lagoon 0 - MAR Martinazo 37.028 -6.438 8 20 Pond 0 - STG Guaperal 37.098 -6.463 7 20 Pond 0 - HOR Hornueco 37.585 -6.548 134 20 Stream 0 - JAR Jarrama 37.707 -6.480 288 20 Stream 0 - OLI Olivargas 37.786 -6.816 236 20 Stream 20.0 (5.7 – 43.7) 4.1 ± 1.2 VIL Villar 37.688 -6.725 229 19 Stream 0 - VLV Valverde 37.568 -6.738 233 16 Stream 0 - PAL Palmones 36.286 -5.595 122 20 Stream 0 - VAL Valle 36.084 -5.693 58 20 Stream 0 - YES Yeso 36.371 -5.842 22 7 Stream 0 - 164 165 Sample collection 166
To address the first objective, adults of the red swamp crayfish were collected 167 during late spring and early summer of 2015 from 14 sites distributed across the western 168
Andalusia region, Spain (Fig. 1A, Table 1). For most sampling sites, 20 crayfish were 169 captured using fyke nets (Table 1), subsequently euthanized by decapitation, and their 170 gastrointestinal (GI) tracts were removed. Samples were fixed in 70% ethanol and 171 maintained in the refrigerator until laboratory analysis. For the second objective, 172 sampling was performed in six ponds of DNS between early March and early April of 173
2018 (Fig. 1B, Table 2 and 3). While crayfish were processed as for the first objective, 174 amphibians were captured by hand nets and skin swabs were collected by gently 175 rubbing a sterile cotton-tipped swab along their entire body. Subsequently, swabs were 176 kept in Eppendorf vials with a drop of pure ethanol and maintained in the refrigerator 177 until qPCR analysis. 178
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Real-time PCR TaqMan assay for Bd-quantification 180
In the laboratory, to ensure no contamination between samples, we used disposable 181 blades to dissect the GI tract from each crayfish. In order to avoid contamination by 182 amphibian tissue inside the GI tissue of crayfish (i.e. in case crayfish had fed on infected 183 amphibians), the sample was washed twice with deionized pressure water. By using 184
PrepMan Ultra Sample Preparation Reagent (Applied Biosystems), we extracted DNA 185 from a small part of the GI tract of crayfish, and from swabs for amphibians. We 186 quantified Bd DNA from GI tracts of crayfish and swabs using standard real-time 187
Polymerase Chain Reaction (qPCR) procedures (Boyle et al. 2004). Amplifications were 188 carried out in a 15μl volume reaction, which included a Bd-specific TaqMan probe 189
(Chytr MGB) for the quantification of zoospore equivalents (ZE). We included 190 amplification standards of 0.1, 1, 10 and 100 zoospore equivalents (ZE, where one ZE is 191 equivalent to a single zoospore) prepared from an isolate of known cell density (IA042, 192
Spain) and a negative control in each plate. All samples, diluted (1/10), standards and 193 the negative control were analysed in duplicate. We considered Bd positive samples if 194 both replicates resulted positive and the amplification curves had the expected 195 sigmoidal shape (otherwise, a 1/100 dilution was made to prevent inhibition problems). 196
In case of contradictory results, the sample was repeated a third time. Infection intensity 197 was reported in zoospore equivalents (ZE) and log-transformed values to show results 198 in graphs. We considered ZE values of 0.1 (Cycle Threshold, CT < 37.0) or higher as 199 positive for infection. Infection intensities are reported as the mean and standard error, 200 unless otherwise noted. 201
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Histological analyses 203
Only crayfish samples that were determined Bd positive by qPCR in the first 204 objective (crayfish infection survey) were confirmed histologically. Samples were fixed 205 in neutral buffered 10% formalin, embedded in paraffin and sectioned, before being 206 stained with haematoxylin and eosin using routine methods (Drury & Wallington 1980). 207
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Statistical analyses 209
While the prevalence of infection and infection intensity was only calculated in crayfish 210 for the crayfish infection survey, we calculated the prevalence of Bd+ and infection 211 intensity for both crayfish and amphibians for the crayfish/amphibian interaction. For 212 both objectives, we calculated prevalence as the proportion of Bd+ with respect to the 213 total number of sampled individuals (crayfish and amphibians), and infection intensity 214 was calculated as Bd load [log10 (Bd load+1)] in each sample. For the second objective, 215 we pooled all amphibian species (see Table 3 for species sampled) and both sexes to 216 increase sample size. As we captured adults and juveniles (non-mature phase), we used 217 life stage as independent variable to control for different time of permanence in the 218 water. We analysed the effects of presence/absence of crayfish and amphibians' life stage 219
(juveniles and adults) on the prevalence of infection and infection intensity of 220 amphibians. We used a Generalized Linear Model (GLM) with a binomial distribution 221 to analyse the prevalence of infection and, in a second analysis only with Bd+- 222 amphibians, we used a linear model to analyse the infection intensity in amphibians 223 from ponds in DNS. The factor “pond” was nested within the factor “presence of 224 crayfish in each pond”; both factors were included in the model because one level of the 225 factor “pond” could not combine presence and absence of crayfish at the same time but 226 ponds could differ in infection intensity. Normality of the residuals of both models was 227 visually inspected and they did not differ from normality. All analyses were performed 228 with JMP 12 software (SAS Inc.). 229
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Crayfish infection survey 232
Only four crayfish out of 267 were positive to Bd, which means a total prevalence 233 of 1.5% (95% CI, 0.4 – 3.8). All Bd+-individuals were found in one population, Olivargas 234 stream (OLI, Fig. 1A), where the prevalence reached 20% and average infection intensity 235 was 4.1 ± 1.2 ZE (Table 1). After detecting qPCR Bd+, histological analyses confirmed the 236 presence of zoosporangia in all four infected GI walls of the red swamp crayfish (Fig. 2). 237
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238 Figure 2. Section of GI tissue from one adult Bd+-red swamp crayfish, Procambarus clarkii, at 239 different scales. Black arrows point out two different zoosporangia containing zoospores of Bd. 240 A) The two zoosporangia at 50 µm; B) the two zoosporangia at 20 µm; C) one of the two 241 zoosporangia at 10 µm; and D) the other zoosporangium at 10 µm. 242 243
Crayfish/amphibian interaction 244
We sampled six different ponds (Fig. 1B) and captured a total of 37 crayfish and 165 245 amphibians of seven different species. A total of seven crayfish were qPCR-positive for 246
Bd infection in their GI walls, which means a prevalence of 18.9% (95% CI, 8.0 – 35.2), 247 and average infection intensity was 1.1 ± 0.5 ranging from 1.7 to 2.0 ZE. On the other 248 hand, a total of 46 amphibians of three species (Triturus pygmaeus, Pleurodeles waltl and 249
Pelophylax perezi) were qPCR positive for Bd infection (Table 3), reaching an average 250 prevalence of 27.6% (95% CI, 21.2 – 35.4) and average infection intensity was 6.5 ± 1.9 251 ranging from 0.1 to 68.0 ZE (Table 2). Except for P. perezi, the other two species were 252 widely represented in our samples. We rarely found positives in anurans (one 253 individual of H. meridionalis) with prevalence of infection of 2.4% (95% CI, 0.1 – 12.6). 254
However, the prevalence of infection in urodelans was of 36.6 % (95% CI, 28.1 – 45.7), 255 being the prevalence within species of 35.1% (95% CI, 24.5 – 46.8) for T. pygmaeus and 256 154 The invasive red swamp crayfish increases infection of the amphibian chytrid fungus
39.1% (95% CI, 28.1 – 45.7) for P. waltl. Prevalence in anurans was significantly lower 257 than in urodelans (p < 0.0001, Fisher’s exact test), but there was no difference between 258 urodelan species (p = 0.8579, Fisher’s exact test). 259
260 Table 2. Procambarus clarkii and amphibian sampling in six ponds of Doñana Natural Space 261 (southwestern Spain), in 2018. For each location we show the coordinates, altitude (m), number of 262 crayfish and amphibians captured (n), prevalence of Bd (%) with 95% Confidence Intervals in 263 parentheses and infection intensity (Bd load in Zoospore Equivalents, mean ± SE). 264 265 Crayfish Amphibians P. clarkii Code Location Lat Lon Altitude n Prevalence (CI) Bd load n Prevalence (CI) Bd load presence ANS Ánsares 37.122 -6.605 23 yes 9 0 (0 – 33.6) - 5 20.0 (0.5 – 71.6) 0.1 OLL Santa Olalla 36.981 -6.473 9 yes 19 26.3 (9.1 – 51.2) 2.0 ± 1.0 40 57.5 (40.9 – 73.0) 9.3 ± 2.3 PAJ Las Pajas 36.98 -6.470 9 yes 9 22.2 (2.8 – 60.0) 1.7 ± 0.0 49 32.7 (19.9 – 47.5) 5.2 ± 4.2 TAR Taraje 36.989 -6.493 10 no - - - 30 10.0 (2.1 – 26.5) 0.1 ± 0.0 ZAH Zahillo 36.987 -6.507 11 no - - - 20 10.0 (1.2 – 31.7) 0.2 ± 0.1 PDR P. del Raposo 36.996 -6.493 8 no - - - 21 4.8 (0.1 – 23.8) 0.1 37 18.9 (8.0 – 35.2) 1.1 ± 0.5 165 27.9 (21.2 – 35.4) 6.5 ± 1.9 266 267 268 Table 3. The 7 amphibian species sampled in Doñana Natural Space in 2018. Life stage is 269 represented by juveniles (J) and adults (A). Prevalence of Bd (%) with 95% Confidence Intervals 270 in parentheses and Bd loads on amphibians (Zoospore Equivalents, mean ± SE) are shown. 271 272 Species n Life Stage (J/A) Prevalence (CI) Bd load Pleurodeles waltl 46 29/17 39.1 (25.1 – 54.6) 5.7 ± 3.8 Discoglossus galganoi 2 0/2 0.0 (0 – 84.2) - Epidalea calamita 4 1/3 0.0 (0 – 60.2) - Hyla meridionalis 29 0/29 0.0 (0 – 11.9) - Pelobates cultripes 5 3/2 0.0 (0 – 52.2) - Pelophylax perezi 2 2/0 50.0 (1.3 – 98.7) 0.1 Triturus pygmaeus 77 0/77 35.1 (24.5 – 46.8) 7.2 ± 2.1 273 274
Our generalized linear model (binomial distribution) explained a 20.6% of total 275 variance (generalized adjusted R2) for the prevalence of infection. The prevalence was 276 affected by the presence of crayfish (χ2 = 28.5, p < 0.0001), the pond (χ2 = 8.8, p = 0.0121) 277 and life stage of amphibians (χ2 = 4.8, p = 0.0280) (Fig. 3A). By using only qPCR Bd- 278 positives, our linear model explained a 22.1% of total variance (generalized adjusted R2) 279
155 Chapter IV for infection intensity. The Bd infection intensity on amphibians was significantly 280 affected by pond (F = 2.90, d.f. = 4.45, p = 0.0342), but not by the presence of crayfish in 281 the pond (F = 1.83, d.f. = 1.45, p = 0.1838) or life stage (F = 1.23, d.f. = 1.45, p = 0.2737) (Fig. 282
3B). 283
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DISCUSSION 285
Identification of non-amphibian hosts is crucial to understand the virulence, 286 distribution, spread and persistence of Bd in aquatic ecosystems worldwide. McMahon 287 et al. (2013) observed encystement of Bd within GI tracts of the red swamp crayfish from 288
Louisiana, the native area of this species. Importantly, our study confirmed the 289 histological evidence of Bd embedded within GI tracts of the red swamp crayfish in its 290 invaded range, indicating that the red swamp crayfish may be a potential carrier of this 291 disease wherever it invades. The presence of the red swamp crayfish could thus imply, 292 besides a predatory effect on amphibians, an indirect effect through the transmission of 293
Bd, thus promoting the decline of amphibians in Europe or elsewhere. Our results 294 indicated that the prevalence of Bd-infection in amphibians was significantly high when 295 the red swamp crayfish co-occurred in the same pond and, to a lesser extent, showed 296 effects of pond (site) and amphibian life stage, being the juvenile stage more susceptible 297 to be Bd positive. For infection intensity we only found a slight effect of pond, probably 298 attributable to the high prevalence and average infection intensity of Bd in amphibians 299 from the Santa Olalla pond. Therefore, our results suggest that non-amphibian species, 300 and the red swamp crayfish in particular, could play a crucial role in maintaining and 301 spreading this emerging infectious disease (chytridiomycosis) in amphibians. 302
303
156 The invasive red swamp crayfish increases infection of the amphibian chytrid fungus
304 Figure 3. (A) Prevalence and (B) infection intensity of Bd in amphibian species from six 305 sampled ponds in Doñana Natural Space. The grey area includes ponds with presence 306 of the red swamp crayfish. In figure 3A, while the number of Bd-negative are depicted 307 in white bars, Bd-positives amphibians are depicted in black bars. 308 309 Due to its great impact on amphibians, chytridiomycosis is considered the worst 310 infectious disease in vertebrates (Gascon et al. 2007; GISD 2018). Sudden high mortality 311 rates in several amphibian species have been related to this emerging panzootic 312
157 Chapter IV infectious disease in the last decades (Berger et al. 1998; Bosch et al. 2001; Daszak et al. 313
2003; Bosch and Martínez-Solano 2006; Skerrat et al. 2007; Scheele et al. 2019). In 314
Andalusian streams, we found a prevalence of Bd of 1.5% in wild red swamp crayfish, 315 which was similar to that found in a previous study in its native area (3.3% in the wild 316 during spring) (Brannelly et al. 2015a). However, all Bd+-crayfish were found in the 317
Olivargas stream, reaching a prevalence of 20% in this location. On the other hand, the 318 red swamp crayfish in DNS showed a total prevalence of 18.9% (seven out of 37 sampled 319 crayfish). While only two out of three ponds were Bd+ for crayfish, we detected Bd+ in 320 amphibians in the six sampled ponds, with an average prevalence of 27.9% reaching the 321 maximum prevalence in the Santa Olalla pond (57.5%). Furthermore, our results showed 322 that three out of seven sampled amphibian species were positives to Bd; these were 323
Triturus pygmaeus, Pleurodeles waltl and Pelophylax perezi, all of which breed in DNS 324
(Gómez-Rodríguez et al. 2009), and the two first are listed as near threatened in the 325
UICN Red List. Even so, it is worth noting that, although the prevalence is still high in 326 amphibian species, and Bd seems to be broadly detected over DNS, no massive 327 mortalities of amphibians have been detected so far. 328
It is well known that the infection of Bd is context-dependent on the amphibian 329 host (Scheele et al. 2017), the fungal virulence (Fisher et al. 2009), and environmental 330 determinants such as temperature, altitude, seasonality and/or rainfall (Berger et al. 331
2004; Kriger and Hero 2007; Walker et al. 2010; Doddington et al. 2013; Ruggeri et al. 332
2018); however, how the fungus can be maintained in the environment is not well 333 known. Unfavourable environmental conditions for the Bd fungus, such as summer 334 desiccation of many temporary ponds, could prevent outbreaks of this disease 335
(Piotrowski et al. 2004), but water availability in permanent ponds could maintain Bd 336 zoospores (Ruggeri et al. 2018). Moreover, some amphibian species, or some individuals 337 within a species, may function as reservoirs of Bd since they are not highly susceptible 338 to infection in spite of harbouring the pathogen and transmitting it to others (van Rooij 339 et al. 2015; Scheele et al. 2017; Brannelly et al. 2018). Finally, other studies in addition to 340 ours have demonstrated infections of Bd in non-amphibian taxa such as reptiles (Kilburn 341
158 The invasive red swamp crayfish increases infection of the amphibian chytrid fungus et al. 2011), waterfowl (Garmyn et al. 2012; Johnson and Speare 2005), fish (Liew et al. 342
2017) and crayfish (McMahon et al. 2013; Brannelly et al. 2015a). The generalist strategy 343 of the Bd fungus, which is able to infect a wide range of hosts, may have profound 344 evolutionary consequences for the pathogen, including the evolution of highly 345 pathogenic strains. 346
The particularity of DNS, which encompasses a system of > 3,000 waterbodies with 347 a singular hydroperiod where temporary ponds fill up with rainfall during autumn and 348 winter and most become completely dry during summer (Díaz-Paniagua et al. 2010), 349 could play a crucial role in the dynamics of Bd. Given that our study shows the presence 350 of Bd in several amphibian species across DNS and high prevalence of Bd in nearby 351 sampling sites that were studied ten years ago (Hidalgo-Vila et al. 2012), the presence of 352 the red swamp crayfish as well as the singular hydroperiod in DNS could explain the 353 existence of Bd through the environment without a sudden decline in amphibian 354 populations. Although our results showed loads of Bd in some amphibian species and 355 individuals in particular, further studies are necessary to assess the role of these 356 amphibian species and particular individuals as reservoirs in DNS. We show that the 357 presence of the highly invasive red swamp crayfish is a relevant factor in the prevalence 358 of Bd. If the red swamp crayfish acts as vector competent of Bd, it would be a good step 359 in understanding the dynamic system of this disease in places where both taxa co-occur, 360 and its participation in transmitting infection to amphibians (see references in van Rooij 361 et al. 2015). Although the introduction pathway of Bd in Doñana ponds is unknown, the 362 positive relationship between the presence of the red swamp crayfish and prevalence of 363
Bd in amphibians suggests an unrecognised indirect impact of crayfish on amphibian 364 populations of Doñana besides of those already known direct effects (Arribas et al. 2014). 365
To conclude, our study suggests a role of the red swamp crayfish as potential 366 reservoir of the Bd pathogen and suitable predictor of its prevalence in amphibians. The 367 suitability of the red swamp crayfish as a potential Bd reservoir (McMahon et al. 2013), 368 together with its ubiquitous presence in temperate habitats worldwide and its positive 369 relationship with high prevalence of Bd in amphibians, may represent new pathways of 370
159 Chapter IV introduction for this disease globally. Furthermore, the invasion routes of the red 371 swamp crayfish have been genetically described for invasive populations of the 372
Northern Hemisphere (Oficialdegui et al. 2019). If the red swamp crayfish is able to 373 transmit this disease to amphibians wherever it invades, then the invasion routes of the 374 red swamp crayfish may help elaborate more adequate conservation plans in order to 375 control the spread of this infectious disease. This study highlights the need to include 376 non-amphibian hosts, especially invasive species such as the red swamp crayfish, as part 377 of the dynamics of Bd infection in order to predict possible outbreaks of 378 chytridiomycosis, as well as its contemplation in future conservation strategies on 379 amphibians. Further research should consider experiments on the role of the red swamp 380 crayfish in the transmission of Bd to amphibians, longitudinal studies to assess the effect 381 of hydroperiod together with the presence of the red swamp crayfish simultaneously, 382 as well as the inclusion of non-amphibian hosts in epidemiological models to estimate 383 clinical outcomes of this panzootic disease. 384
385
ACKNOWLEDGEMENTS 386
We are strongly grateful to C. Serrano, M.A. Bravo, V. Castaño and R. Arribas for 387 their help in the fieldwork, M. Wass for his help in the laboratory, and C. Díaz-Paniagua 388 and The Monitoring Team on Natural Resources and Processes of the Doñana Biological 389
Station for their valuable and helpful comments. We also thank the Laboratorio de 390
Histología at Museo Nacional de Ciencias Naturales (CSIC) and the Laboratorio de SIG 391 y Teledetección (LAST-EBD) at Estación Biológica de Doñana (CSIC) for providing 392 logistical support. This study was funded by the Andalusian Government (RNM-936). 393
F.J.O. was supported by an Andalusian Government grant. 394
395
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161 Chapter IV
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163
CHAPTER V
Rigid laws can set back invasive species management
Oficialdegui F. J, Delibes-Mateos M., Green A. J., Sánchez M. I.,
Boyero L., Clavero M.
Under review in Conservation Biology
Chapter V
INTRODUCTION
Social conflicts are widespread and increasing phenomena that arise from the diversity of values, beliefs and interests that exist within societies (Redpath et al. 2013).
Conservation conflicts are frequent in the management of invasive species, particularly when such species simultaneously provide socioeconomic benefits and cause environmental impacts (van Wilgen & Richardson 2014). Conflicts arise between groups that promote use of invasive species, and other groups that oppose use for the benefit of biodiversity (i.e. conservationists) (Crowley et al. 2017). When restrictions to exploitation are applied in response to conservationists’ demands, the users of invasive species perceive that conservationists’ values and interests are being imposed, conforming a typical win-lose scenario of conflict (Redpath et al., 2013) (Fig. 1). We argue here that this scenario might be justified whenever restrictions result in a real benefit for biodiversity, understood as a common societal good (Treves et al. 2017), but can be untenable if such conservation benefits are not obtained. The absence of tangible benefits of restrictions to the use of invasive species can generate a perception of pointlessness towards the management of this problem and a loss of social support for conservation
(Fig. 1), with the potential of producing a deregulation rebound effect. Here, we illustrate this with the ups and downs in application of regulations for the use of the red swamp crayfish (Procambarus clarkii) in Spain, considered one of the 100 worst invasive species in Europe.
CRAYFISH USES, PERCEPTIONS AND REGULATION
Native to southern United States and northeast Mexico, the red swamp crayfish has been introduced in many regions worldwide (Oficialdegui et al. 2019), causing severe negative effects over many aquatic organisms and the functioning of the invaded ecosystems (Geiger et al. 2005). It was first introduced to Spain for commercial purposes in 1973, rapidly spreading across the Iberian Peninsula, where it is now almost omnipresent (Kouba et al. 2014). This species is a key economic resource in the
Guadalquivir marshes in SW Spain, where it was introduced in 1974 and is
166 Rigid laws can set back invasive species management commercially exploited since the late 1970s (Gutiérrez-Yurrita et al. 1999). This exploitation has generated an exportation-based industry (Martín-López et al. 2011), which currently processes some 4,000 tons of wild-caught red swamp crayfish yearly, involving between 150,000 and 200,000 work-days and generating a gross income of around €20M (Conde & Domínguez 2015). This is a critical economic sector in an area that offers few other job opportunities.
Figure 1. Left: general framework for conservation conflicts, modified from Redpath et al. (2013). In a typical conflict situation, non-conservationist (a) and conservationist (b) stakeholders strive to win, independently of the interests of the others. Conflict resolution aims at redirecting this win-lose situation towards a more neutral scenario that requires trade-offs or, ideally, to a win- win scenario. Right: dynamic scenarios resulting from the regulation of the invasive red swamp crayfish (Procambarus clarkii) in Spain. (1) Legal restrictions supported by environmental NGOs ban crayfish trade (conservationists are winners; stakeholders related to crayfish exploitation are losers). (2) Since control of crayfish populations is not feasible (more so, eradication), the ban does not provide environmental benefits (conservationists obtain a neutral outcome). (3) Economic and job opportunity losses not linked to any environmental benefit drive the social discrediting of conservationist stakeholders and lead to the downgrading of societal targets for biodiversity conservation, especially of those dealing with invasive species (everyone loses). (4) Legal modifications shield the interests of the crayfish industry (and those of other users of invasive species) by an unspecific allowance of the use of invasive species for reasons of “social and economic nature”.
167 Chapter V
In 2007, the Spanish Natural Heritage and Biodiversity law (Ley 42/2007) ordered the creation of the Spanish Catalogue of Exotic Invasive Species ('Spanish Catalogue'), which should list non-native species that constitute a “severe threat” for native biodiversity or socio-economy. According to the 42/2007 Law, listing a species in the
Catalogue implied the general prohibition of its possession, transport and commercialization. The Spanish Catalogue, issued in 2013 (Real Decreto 630/2013), listed the red swamp crayfish, but included an additional provision excluding it from the restrictions mandated by the 42/2007 Law. On March 16th 2016, the Spanish Supreme
Court declared null and void this exception, in response to the lawsuit filed by several environmental NGOs. This decision implied the prohibition of red swamp crayfish commercial exploitation and was received with widespread alarm and social protests in the Guadalquivir marshes area (Fig. 2). The Andalusian Regional Government rapidly passed a crayfish “control plan” (BOJA 2016, August 3rd) to unblock crayfish exploitation. This plan established that crayfish fishermen were responsible for
“population control” operations, while crayfish should be delivered to the crayfish industry processing facilities, to be traded. Finally, in 2018 the Spanish Government modified the 42/2007 Law, in order to "balance the essential fight against exotic invasive species and their exploitation for hunting or angling” (Ley 7/2018). In particular, it included the possibility of delisting a species from the Spanish Catalogue, or eluding the restrictions that the listing involves, due to “overriding reasons relating to a public interest, including those of a social and economic nature”. The preamble of this legal modification specifically mentions the “great concern” generated by the prohibition of the commercial exploitation of the red swamp crayfish.
POINTLESS CONFLICTS AND WAYS OUT
Social conflict around the management of invasive species is often unavoidable
(Estévez et al. 2015), but when restrictions on the use of invasive species result in improvements for the environment they should be implemented even in the face of opposition, considering the environment protection as a common good. But the ban of
168 Rigid laws can set back invasive species management the red swamp crayfish trade was a highly controversial measure without tangible environmental benefit, because this species cannot realistically be controlled in the
Guadalquivir marshes (see Gherardi 2006). In fact, the crayfish industry already extracts around 250 million crayfish individuals yearly, an activity that has not resulted in any significant population control and that, most probably, would not be maintained by publicly-funded conservation organizations in a commercial-ban scenario. In summary, what initially seemed like a victory for conservationists over the crayfish industry turned into a loss of conservationists’ reputation and ultimately led to a downgrading of invasive species management (Fig. 1).
Figure 2. (A) Crayfish delivered to be processed in one of the crayfish industries in Isla Mayor, and (B) demonstration on May 9th 2016 in Isla Mayor, Seville, against the Spanish ban on crayfish exploitation.
We argue that the undesirable process related to invasive species management featured in Figure 1 has its roots in the rigidness of the 42/2007 Law, in which drug-like restrictions (i.e. no possession, no transport, no commerce) were invariably applied to all listed species. The law thus treated a complex problem as a simple one, a mistake that has been shown to reduce the likelihood of effective management of invasive species
(Woodford et al. 2016). Rigid legal restrictions overlook the fact that the impacts of invasive species are context-dependent, that they often change over time and across space, and that best management practices must adapt to this variability (Gosnell et al.
2017; Beever et al. 2019). Moreover, simplistic legal restrictions can be reversed in an equally simplistic manner in response to public opinion (e.g. Miller et al. 2018). This 169 Chapter V leads to a win-lose scenario in which the interests of the users of invasive species are imposed over conservation goals (Fig. 1). In this sense, the 2018 modification of the
42/2007 Law was precipitated by the red swamp crayfish, but also applies to a set of socio-economically relevant invasive species involving a highly diverse set of environmental problems and management options (e.g. the rainbow trout, Oncorhynchus mykiss; the largemouth bass, Micropterus salmoides; the American mink, Neovison vison; or the Barbary sheep, Ammotragus lervia).
Legal instruments regulating invasive species should overcome simplistic approaches usually demanded by either conservationists (i.e. rigid ban) or the users of such species (i.e. full exploitation). As an alternative, invasive species policy should rely on a more holistic, flexible and adaptive management (Novoa et al. 2015; Shackleton et al. 2019). Regulatory measures complementing total bans could focus on introduction vectors (e.g. ballast water management, Woodford et al. 2016) or specific features that determine success at different stages of the invasion process. For example, the ban of wild-caught bird trade issued by the European Union in 2005 as a response to avian influenza resulted in a drastic decline in the number of bird species introduced to
Europe (Cardador et al. 2019). Interestingly, this unprecedented conservation success did not rely on species blacklisting and did not involve serious conflicts with the pet bird market, which rapidly adapted to a captive-bred offer with smaller invasive potential (Cardador et al. 2019). In our specific case study, red swamp crayfish policy could focus on banning the possession and transport of live specimens to avoid further spread of the species, while permitting the current trade involving processed crayfish.
The implementation of policy regulating invasive species usually lacks social consensus, as in most cases laws are passed because of lobbying by one of the conflicting parties. Alternative, collaborative approaches are needed to design and implement flexible legal frameworks on invasive species management (Estévez et al. 2015; Novoa et al. 2018; Shackleton et al. 2019). This would help not only to avoid sterile social conflicts among involved stakeholders, but also to gain the support of society at large, thus helping to maintain the many recent advances in the management of biological
170 Rigid laws can set back invasive species management invasions and the resulting environmental improvements (Estévez et al. 2015; Crowley et al. 2017). These kind of approaches would also avoid uncertainties in the regulation framework arising from drastic legal changes, such as those related to crayfish exploitation in Spain, which have been shown to reduce the willingness of stakeholders involved in conservation conflicts to collaborate (Pollard et al., 2019). Societal stakeholders must feel involved in the planning of regulations affecting their interests and that these regulations support a common good, in order to avoid misconception, opposition and setbacks.
ACKNOWLEDGEMENTS
This work was funded by the Andalusian government RNM-936. M. Delibes, A.
Rodríguez, C. Gutiérrez-Expósito, E. Revilla, P. Pons, J. Román, C. Rodríguez, P.
González, F. Ribeiro and V. Hermoso discussed previous versions of the manuscript and helped us to improve it. F.J.O. was supported by an Andalusian Government grant.
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Gutiérrez-Yurrita PJ, Martinez JM, Ilhéu M, Bravo-Utrera MA, Bernardo JM, Montes C. 1999. The status of crayfish populations in Spain and Portugal. Crustacean Issues 11:161-192. Kouba A, Petrusek A, Kozák P. 2014. Continental-wide distribution of crayfish species in Europe: update and maps. Knowledge and Management of Aquatic Ecosystems (413):05. Martín-López B, García-Llorente M, Palomo I, Montes C. 2011. The conservation against development paradigm in protected areas: Valuation of ecosystem services in the Doñana social–ecological system (southwestern Spain). Ecological Economics 70:1481-1491. Miller RS, Opp SM, Webb CT. 2018. Determinants of invasive species policy: Print media and agriculture determine US invasive wild pig policy. Ecosphere 9(8):e02379. Novoa A, Kaplan H, Kumschick S, Wilson JRU, Richardson DM. 2015. Soft touch or heavy hand? Legislative approaches for preventing invasions: Insights from Cacti in South Africa. Invasive Plant Science and Management 8:307-316. Novoa A, et al. 2018. A framework for engaging stakeholders on the management of alien species. Journal of environmental management 205:286-297. Oficialdegui FJ, Clavero M, Sanchez MI, Green AJ, Boyero L, Michot TC, Klose K Kawai T, Lejeusne C. 2019. Unravelling the global invasion routes of a worldwide invader, the red swamp crayfish (Procambarus clarkii). Freshwater Biology: in press. DOI: 10.111/fwb.13312 Pollard CR, Redpath S, Bussière LF, Keane A, Thompson DB, Young JC, Bunnefeld N. 2019. The impact of uncertainty on cooperation intent in a conservation conflict. Journal of Applied Ecology (in press) Redpath SM, et al. 2013. Understanding and managing conservation conflicts. Trends in ecology & evolution 28:100-109. Shackleton RT, et al. 2019. Stakeholder engagement in the study and management of invasive alien species. Journal of environmental management 229:88-101. Treves A, Chapron G, López-Bao JV, Shoemaker C, Goeckner AR, Bruskotter JT. 2017. Predators and the public trust. Biological Reviews 92(1):248-270. van Wilgen BW, Richardson DW. 2014. Challenges and trade-offs in the management of invasive alien trees. Biological Invasions 16:721-734. Woodford DJ, Richardson DM, MacIsaac HJ, Mandrak NE, van Wilgen BW, Wilson JRU, Weyl OLF. 2016. Confronting the wicked problem of managing biological invasions. NeoBiota 31:63–86.
172
GENERAL DISCUSSION
We live in a globalized world, where travelling great distances in a short period of time is feasible for hundreds of millions of persons and for an incalculable amount of goods. Transport of species across geographic barriers, occurring at a clearly accelerated rate during the last decades (Seebens et al. 2017) have led to important losses of biodiversity (Blackburn et al. 2019) and ecosystem services (Vilà et al. 2010), as well as to the blurring of natural biogeographic boundaries, leading to the homogenization of biotas worldwide (Capinha et al. 2015). Those species transported and introduced in an area outside of their native range are called alien species (Gilroy et al. 2017), and those alien species that are able to establish, thrive, become abundant, and spread over non- native areas are named as invasive species (Blackburn et al. 2011). Studying well-known and globally distributed invasive species is of special interest to predict and halt invasions of others less known alien species (Larson & Olden 2012). Understanding why invasive species are introduced, why they succeed in establishing viable populations in new areas and how they are able to expand their non-native rages to become widely distributed are key issues to understand the invasion process (Seebens et al. 2019) as well as to prevent and reduce the impacts of biological invasions (Estoup & Guillemaud
2010). Addressing questions at broad scales is of great importance for exploring complex patterns in invasion processes (e.g., admixture, invasive bridgehead effect or propagule pressure; Fig. 1 and 2). While the first three chapters of this thesis focus on understanding of invasion routes as a measure of prevention for emerging invasive species, taking as reference a cosmopolitan, human-dispersal and well-known invasive species; the other chapters address questions about its impacts on invaded ecosystems and launching particular concerns about the management of invasive species once established (Fig. 3). By using a multidisciplinary approach, this thesis describes the invasion history of one of the most harmful invasive species in freshwater ecosystems worldwide, the red swamp crayfish, Procambarus clarkii; analyses the role of commercial companies in translocation of live-specimens; infers its main invasion routes; explores
175 General discussion the indirect effects on native amphibian communities; and assesses the societal conflict derived from its use as commercial invasive species and its management.
RECONSTRUCTING A GLOBAL INVASION
Often, the distribution ranges and introductions outside of the native area of many invasive species remain poorly described in the literature, mainly those from the past.
By contrast, other alien species gather many occurrences and their translocations are well-known so that studying their invasion processes over time can be useful for management of emergent invasive species (see Larson & Olden 2012). By compiling reports described in global databases, scientific and grey literature, I was able to illustrate the changes in the geographical distribution of the red swamp crayfish throughout the last century (Chapter 1). The species started to be introduced out of its native range in the 1920s (to California, Hawaii and Japan, and from there to China), but its spread was limited. Following the increasing commercial interest of the red swamp crayfish in Louisiana by the 1960s, the species started being introduced into new areas with the aim of simulating Louisiana crayfish exploitation, and to spread within them, as well as within the first non-native territories (i.e. California and Japan). By then, numerous translocations took place worldwide being introduced into several countries in Central America, Africa and Europe. More recently, the striking rise of crayfish industry in China as well as the commercialization of this species in pet shops has occasioned an increase in wild populations of this species over several Asian countries.
Hence, the red swamp crayfish has become one of the most cosmopolitan species. In last years, the number of red swamp crayfish records has grown exponentially, mainly in
Europe and U.S.A., thanks to extended monitoring as well as to be easily traceable by non-expert citizens due to its size, particular morphological features (see in Loureiro et al. 2015) and, burrowing activity in the riverbeds (see in Barbaresi et al. 2004), or be observable in the diet of several predators such as birds or otters. It is hence that citizen science can be an extremely powerful and solid tool for mapping actual species distributions (see Sumner et al. 2019) and early detection of highly invasive species.
176 General discussion
The red swamp crayfish, as many other invasive species (e.g. pampas grass, common carp, signal crayfish, etc.), has been translocated by humans for long because of multiple reasons. For example, its introduction has been related with food sources
(food for American bullfrogs in California, Hawaii and Japan), fisheries (sport fishing or live bait in western U.S.A.), live seafood for restaurants (e.g., France or the Netherlands), aquaculture (astaciculture or wild harvest in Louisiana, China, Spain or Kenya), pet shops and aquarium suppliers (in Brazil in central Europe or Indonesia) (Lodge et al.
2000), but also other additional ways of introductions have to be considered such as school science programs in western US (Larson & Olden 2008). In the Chapter 1, I explored how the red swamp crayfish has been globally translocated over time.
Although this species has been mainly used as study model for physiology studies, the tendency of published papers changed to invasion issues once the species has been introduced and naturalized worldwide. Thus, I showed that the great expansion of this species because of the growth of crayfish industry decades before favoured an increase in studies focused on invasion science since the decade of 1990s. Other crayfish species have also rapidly grown in popularity as ornamental, pet and seafood species such as the common yabby Cherax destructor, the redclaw Cherax quadricarinatus, the signal crayfish Pacifastacus leniusculus or the marbled crayfish Procambarus virginalis that have been globally translocated due to its economic benefit in aquaculture or pet trade
(Holdich 1993; Chucholl 2013; Patoka et al. 2015; Souty‐Grosset et al. 2016). Therefore, more conservation measures should be paid on translocated species with a greatly invasive potential associated to aquaculture or wild-harvesting and pet trade, as for example, cooperative engagement between hunters/anglers or pet keepers and scientifics in order to prevent future escapes and releases of potentially risk invasive species.
Additionally, the Chapter 1 focused on the potential role of biological supply companies in the translocation of invasive species, which sell live-specimens of laboratory species. Our findings showed that multiple, often unrecorded translocations had occurred from native to non-native areas and vice versa, as well as among non-
177 General discussion native areas, entailing a conservation concern because invasion history can be much more complex than assumed. I pointed out that many translocations of live-crayfish were carried out by scientists, though research-derived escapes or releases are rarely described in the literature. Although I focused here on translocations for scientific research, the resulting figures can be seen as surrogates for translocations with other motivations, could have been promoted by common citizens, anglers, farmers or school programmes (Faulkes 2015). These results imply that the amount of uncontrolled translocations of the red swamp crayfish, and arguably of other invasive species, may have been underestimated. Maceda-Veiga et al. (2016) described some positive actions through cooperative efforts between scientifics and aquarists in order to prevent the releases of invasive species into the wild. This engagement could help to avoid releases from aquarium keepers but also, to halt many additional translocations of highly invasive species, especially. Therefore, multidisciplinary approaches are necessary to unveil routes of introduction which have not been portrayed in literature.
Despite the fact that describing the invasion history based on georeferenced records and invasion reports are useful for mapping species-distributions, many established populations are hardly traceable because of the lack of available information, especially when introductions that took place long ago. Identifying the routes of introduction of alien species is crucial for facilitating their management and regulation in the early stages of invasion process (Cristescu 2015). In this context, molecular tools can help reconstruct the invasion routes, to disentangle the status of cryptogenic species or to describe the genetic structure of invasive populations (Estoup & Guillemaud 2010;
Lombaert et al. 2010; Clavero et al. 2016). In the Chapter 2, I showed the main invasion routes of the red swamp crayfish in the Northern Hemisphere using genetic tools that allowed unveiling previously unreported and unconfirmed introduction routes into
Europe (Chapter 3). To reconstruct the globally invasion routes genetically, one critical piece of the puzzle is to describe exhaustively the genetic structure in the native range
(Fitzpatrick et al. 2012). In order to infer the invasion routes and population structure, I analysed a fragment (608bp) of the mitochondrial marker COI from 1,062 individuals of
178 General discussion the red swamp crayfish plus 354 GenBank sequences in a total of 122 populations, of which 22 were populations from the native area (Chapter 2). Interestingly, high haplotype diversity was found in the native range with no clear population structure
(arguably due to genetic admixture) but also in some non-native areas (e.g., the western
United States or southern Europe), suggesting a complex pattern of multiple introductions. A similar pattern of multiple introduction events has been recently found in other invasive aquatic species such as the oriental shrimp (Lejeusne et al. 2014) or signal crayfish (Petrusek et al. 2017). By contrast, the opposite pattern (low haplotype diversity) was described in Asian populations, consistent with the founder effect of few crayfish introduced in Japan giving rise to the current established populations in Asia
(Kawai 2017). These results provide a clear example of how different features of invasion processes (e.g. genetic admixture, propagule pressure or secondary introductions, see
Fig. 1 and 2) can generate contrasting genetic diversity patterns across non-native populations of a global invader. In turn, these different invasion scenarios propose further questions. For example, if there is low haplotype diversity (e.g. in Asia), how does a population become invasive (i.e. the genetic paradox)? (Estoup et al. 2016). These scenarios, together with other cases such as the parthenogenetic marbled crayfish, which is able to establish wild populations from a single released individual (Feria & Faulkes
2011), suggests new evolutionary paradigms in invasion biology (Andriantsoa et al.
2019). Moreover, a successfully established invasive population can serve as source of new invasions, consequently, multiple secondary introductions can have place. If there is an evidence for adaptive evolution in invasive populations that act as source of new introductions, it is a key evolutionary question in invasion science (i.e., invasive bridgehead effect; Lombaert et al. 2010; Bertelsmeier & Keller 2018). So, by having already faced novel environmental conditions, are bridgehead populations key in evolving higher invasiveness of subsequent introductions? (Bertelsmeier & Keller 2018).
In fact, this invasive bridgehead effect might have taken place in Kenya or China, where previously successfully established populations seems to have been the sources of invasive populations in the northern Europe and Malta (Chapter 3), but further studies
179 General discussion focusing on invasiveness and evidence for adaptive evolution of previously established populations are needed to elucidate whether the invasion bridgehead effect has taken place in a specific invasive population. This complex scenario of invasion with high genetic admixture and multiple secondary introductions (Chapters 1-3) has led to large intraspecific genetic variability within and among invaded areas, entailing potential applied issues in conservation biology (Chapter 2). Specifically, large haplotype variability may reduce the efficacy of environmental DNA (eDNA) protocols to detect and survey invasive species (Wilcox et al. 2015). Our genetic dataset and the large-scale phylogeographic study can provide an accurate baseline for improving detection of global protocols in this invasive species based on eDNA (Ficetola et al. 2008; Larson et al. 2017).
TRANSMISSION OF EMERGING DISEASES
Due to their effects and consequences on ecosystems, biodiversity and human well- being, impacts of biological invasions have been long discussed (Simberloff et al. 2013).
Particularly, the red swamp crayfish is a keystone species in freshwater ecosystems with severe effects on the whole food web, consequently, its impacts have been well studied across different invaded areas (Geiger et al. 2005; Anastácio et al. 2000; Souty-Grosset et al. 2016). In particular, the presence of the red swamp crayfish has been subject to reduction of amphibian richness (Cruz et al. 2006), nonlethal inquiries in tadpoles
(Nunes et al. 2010) and amphibian survival (Arribas et al. 2014). While many studies have shown direct impacts on amphibians (e.g., predation; see references above), very few studies have focused on indirect impacts. Identifying the role of the red swamp crayfish in pathogen maintenance seems vital for disease control strategies, mainly since this invasive species is distributed worldwide. In the Chapter 4, I showed the role of the red swamp crayfish in the transmission of chytridiomycosis, caused by the fungus
Batrachochytrium dendrobatidis (Bd) and considered one of the most important causes for the decline of amphibian populations worldwide. While previous studies had found Bd encystement in the wall of the crayfish gut (McMahon et al. 2013), our histological
180 General discussion analyses seemed to show the presence of Bd zoosporangia in the wall of the crayfish gut, which might represent the development of Bd cycle in a non-amphibian host. Although the prevalence of Bd in the red swamp crayfish in Louisiana, native area, was 3.31%
(Brannelly et al. 2015a); contrasting patterns were found in crayfish in invaded streams of south-western Spain and within Doñana Natural Space where the prevalence of Bd was from 1.5% to 18.9%, respectively (Chapter 4). This increase in prevalence might be due to the high abundance and diversity of amphibians in this highly protected area of the south-western Spain, where Bd is long maintained in the environment. Thus, I demonstrated that the presence of the red swamp crayfish significantly altered to infection of Bd in amphibians, indicating that this crayfish can be a suitable predictor of
Bd infection in co-occurring amphibians. The potential of the red swamp crayfish as reservoir of Bd has received little attention to date (McMahon et al. 2013, Brannelly et al.
2015a, Brannelly et al. 2015b) but the results of this thesis highlight the need to consider non-amphibian hosts, especially invasive species like the red swamp crayfish, as part of the dynamics of Bd infection in order to predict possible outbreaks of chytridiomycosis, as well as its consideration in future conservation strategies of amphibians.
INVASIVE SPECIES MANAGEMENT
Inflexible and simplistic legal tools (most commonly, full bans based on black- listing) can be counterproductive for the management of invasive species due to the generation of sterile and unsolvable social conflict. In the Chapter 5, I dealt with the societal conflict underlying on the commercial use of invasive species and its management. Specifically, I focused on the veto that Spanish laws imposed to all red swamp crayfish uses in 2016. If laws forbid the use of socio-economically relevant invasive species, there is no resulting benefit for biodiversity afterwards (eradication of established populations with large densities of the red swamp crayfish is virtually unfeasible), this may generate a perception of pointlessness towards the management of invasive species and a loss of social support for it and for conservation in general. In this chapter, I advocated for minimizing (as avoiding them is virtually impossible) social
181 General discussion conflicts around the management of invasive species. Such conflicts should be faced only when a benefit for the common good (i.e. biodiversity conservation) is expected. I discussed the specific case of the management of the red swamp crayfish in Isla Mayor
(SW Spain), offering relevant lessons to be learnt for the management of biological invasions worldwide. The implementation of policies regulating invasive species usually lacks social consensus because in most cases laws are passed because of lobbying by one of the conflicting parties. Collaborative approaches are needed to design and implement flexible legal frameworks on invasive species management (Estévez et al.
2015; Novoa et al. 2018; Shackleton et al. 2019). Social sciences and humanities have taken a growing interest in biological invasions in the last decades (Vaz et al. 2017). New frameworks for resolving conflicts of interest are urgently needed and have already started to be developed (Gaertner et al. 2016; Novoa et al. 2018).
Overall, this thesis investigates the invasion process of the red swamp crayfish from a multidisciplinary approach (literature review, field samplings, and molecular analyses) and multiple perspectives, from the comprehension of the global invasion history, the impacts on native biodiversity to the conflicts and management of biological invasions. Results derived from this thesis shed light on the understanding of invasion routes of the red swamp crayfish at the large- and short-scale, its role in maintaining and spreading an emerging infectious disease of amphibians (chytridiomycosis) and expose social conflicts among different stakeholders. The outcomes of this thesis serve and can be applied to the study of other aquatic invasive species whose targets are aquaculture or pet markets (e.g., fishes) and raise key questions to be addressed in future research.
FUTURE PERSPECTIVES
This thesis sheds light about various stages of the invasion process of the red swamp crayfish (Procambarus clarkii) and leads to many other evolutionary, epidemiological, management and social questions. Firstly, the Chapter 1 opens the way for assessing the role of aquaculture species and biological supply companies or pet shops as potential sources of invasive populations, which have been related with
182 General discussion escapes or releases into the wild (Nunes et al. 2015). To what extent are the biological supply companies able to have influence on the spread of invasive species? For example, by having genetic samples or sales information from the main biological suppliers of the red swamp crayfish and using the results obtained from the Chapters 2 and 3, invasion routes could be inferred and sources of wild invasive populations identified. Also increasing conservation awareness among those managers of aquarium shops that act as source and domestic invasive species keepers, something that it is of vital importance to prevent further releases of invasive species into the wild (Maceda-Veiga et al. 2014).
Moreover, scientific studies should include where highly invasive species are captured or obtained from, thereby having valuable information about their provenances and translocations.
Another evolutionary question that arises from this thesis is whether invasive bridgehead populations are potentially more successful invaders or not. According to
Bertelsmeier & Keller (2018), the bridgehead populations evolve higher invasiveness; hence, invasive populations from Kenya or China to Europe (Chapter 3) could be a starting point for experimental approaches on adaptive evolution of introduced populations. Moreover, low haplotype diversity scenarios of invasion process (e.g. Asia in Chapter 2) can provide interesting scenarios to test relevant evolutionary hypothesis such as the genetic paradox: how individuals may compensate the loss in genetic diversity by adaptive phenotypic plasticity or epigenetic changes (Estoup et al. 2016)?.
Among the multiple factors that must be considered into epidemiological models to study the dynamic of the fungus disease chytridiomycosis. The Chapter 4 underlines the role of non-amphibian hosts as potential carriers of the disease, especially when globally introduced invaders are involved (the red swamp crayfish, Chapter 1).
However, while we have demonstrated that the red swamp crayfish is able to act as reservoir, the ability to transmit the diseases to co-occurring amphibian species is still unknown. Therefore a promising area of research is to develop experimental approaches to test the role of the red swamp crayfish as vector of Bd.
183
CONCLUSIONS
1. The growth of the crayfish industry in Louisiana (US) based on the red swamp crayfish (Procambarus clarkii), in the 1960s and 1970s, led to the worldwide expansion of this species.
2. Biological supply companies may have played an important, though largely overlooked, role in the translocation of the red swamp crayfish worldwide.
3. The history of invasion by the red swamp crayfish is more complex than generally assumed, involving an extensive genetic admixture in native populations, different scenarios of propagule pressure and multiple secondary introduction events.
4. Two previously unconfirmed routes of introduction of the red swamp crayfish into Europe, from Kenya and China, have been identified by analysing patterns of genetic variability.
5. The presence of the red swamp crayfish increases the infection of the chytrid fungus Batrachochytrium dendrobatidis in co-occurring amphibians.
6. The presence of zoosporangia of Batrachochytrium dendrobatidis in the gastrointestinal tissue of the red swamp crayfish suggests a role for invasive crayfish as reservoir of this emerging disease.
7. Prohibiting the use of exploited invasive species when control or eradication are not feasible management options may trigger sterile conflicts and drive a loss of social support for the management of invasive species.
8. Legislation regulating invasive species should introduce flexibility in the application of management options, in order to adapt to the several context dependencies and ensure the efficacy and sustainability of the strategies to deal with this first-order environmental issue.
185
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ACKNOWLEDGMENTS / AGRADECIMIENTOS
Aunque esta tesis haya sido corta, no lo ha sido mi estancia en la EBD. He tenido la oportunidad de conocer grandísimas personas, a las que de una forma u otra me gustaría agradecer el haber creado el ambiente idóneo para disfrutar enormemente durante todo este tiempo.
En primer lugar me gustaría agradecer la oportunidad de hacer esta tesis a mis directores. Marta, gracias por confiar en mí, por apoyarme en todo lo que se ha propuesto y por haberme animado a dar cada paso durante estos poco más de tres años. A Miguel, me alegro de haberte tenido al lado, ha sido fabuloso conocerte como persona y, por supuesto, como investigador. Por esas charlitas justo antes de irte a casa y por incluirme en todo el mundo “cangrejil”. A Luz, aunque haya sido un poco más a distancia, gracias por siempre estar ahí dispuesta a echar una mano para cualquier cosa que he necesitado. En definitiva, os estoy muy agradecido a los tres por el apoyo que he tenido desde que planteamos la tesis al inicio, durante estos tres años, hasta esta última y estresante etapa con mails, discusiones y revisiones casi a diario. Gracias por estar conmigo! Indudablemente, también me gustaría resaltar a varias personas que de una o de otra manera me han acompañado a lo largo de esta tesis. Raquel, gracias por estar ahí desde el inicio, en las aventuras del campo y de laboratorio. Me alegro de haber coincidido contigo, siempre resolutiva en campo, en laboratorio, despacho y dónde sea. Eres una crack! Todo lo que pueda decir de ti, es poco. Miguel Ángel, gracias por estar para cualquier cosa, conocer mucha parte de las aguas andaluzas, incluida Doñana, ha sido todo un privilegio hacerlo a tu lado. Andy, gracias por contar conmigo para poder colaborar con otros centros de investigación, primero en Holanda y luego en Francia. Han sido dos experiencias muy gratificantes, a nivel personal y científico, y no habrían sido posibles sin ti. Gracias José Guerra, desde el primer día al pie del cañón para echarme una mano en todo lo que he necesitado, es un grandísimo placer haberte conocido. A quarter part of my thesis I was abroad. It allows me to know a lot of amazing people. Thanks all of them for supporting me and making me feel at home. Let’s start from the beginning!! By order, David, merci beaucoup!! Thanks for showing me the amazing world of the proteomics. It is a pity that our work was not reflected on this thesis but papers together come out soon, sure!. I would like to say a big “Thanks” to Iris. It was a pleasure to meet you. Thanks for teaching me to get me on better at lab (a lot of French unknown rules haha) and inviting me to join to any meetup in my shorts stays in Clermont-
195 Acknowledgements/Agradecimientos
Ferrand. And, of course, I would like to thank all people at the department, phD students, postdocs, technicians and Spanish-speaking group for all these beers, indoor football, hiking nearby, more beers, cups of coffee and so on (ahh!! and some scientific discussion as well, of course :P). Muchas gracias Jaime por adentrarme en el mundo de los anfibios y sus hongos. Me ha molado mucho meterme en este tema y más con un investigador como tú. Ha sido una gozada trabajar contigo. Bedankt Edwin and Ivo!! Both of you were two admirable hosts. Thanks for making me feel one more from the beginning of my stay, for the discussions, for making everything easier, for letting me carry out all my experiments and helping me to get it. It was a great pleasure to meet you and are still in touch with both of you, thanks a lot!. I am also so lucky to have met many people there: the Spanish team (Pablo, Ariadna, Paula), Maira, Bernardo, Jelle, Arie, Usman, Seb, Arianna, Bregje, Marlies, John and a lot of good people!! Christophe!! Merci beaucoup!! Qué bueno fue ir a Roscoff a disfrutar del apacible, seco y caluroso invierno jaja. Fue un placer ir a trabajar allí, esas pausas café y las discusiones genéticas. Gatito!! Jajaja Gracias por todo =) I am also delighted to have learnt from you, Claire, thanks for teaching me the microsatellites world and helping me at lab. I would like to say thanks to other PhD-Students, post-docs and researchers who I knew along my short-stay at Roscoff Biologic Station. Muito Obrigado Filipe Ribeiro for giving me positive comments to my thesis. It was a pleasure to have met you! Next time I promise you to have learnt a bit of Portuguese :P Y gracias Pablo Burraco por ese buen feedback. Ya buscaremos el nexo entre los dos para hacer algo juntos, aunque sea tomar unas cervezas jaja.
Al departamento de ecología de humedales, ha sido un placer compartir estos años con vosotros, ese dinamismo, Jornadas de Departamento, los “You Speak I Eat” aunque muchas veces era una excusa para ponernos hasta las cejas a las 12h. Muchas gracias a todos: Iván, Poli, Elena, Luis, Cristina Ramo, Juan, Cristina Pérez, Miguel Ángel, Pablo, Laura, Javier, Christoph, MariJose, y un grandísimo etc.!! Gracias a todo el personal de la Estación Biológica de Doñana por todo el cariño, apoyo y ayuda prestada durante la tesis (y antes). Desde aquellos que nos saludan tempano por la mañana y nos despiden por la tarde: Pedro, Jesús, Manuela, María, Franciscos,… Incluso a veces dando las buenas noches y buenos días a la misma persona. A todo el personal de dirección y coordinación por facilitarnos permisos, estancias, ferias de divulgación y un gran etcétera: Bego, Irene, Guyonne, Sofia,… Aunque hay muchas personas detrás de cada justificación y proyecto, me gustaría dar un especial
196 Acknowledgements/Agradecimientos agradecimiento a Antonio López y Carmen Velasco por soportar todos los jaleos del PICS, que si este no es, que es el otro proyecto, ahora hay que pedir a un sitio luego a otro, vaya quebraderos de cabeza nos han dado :D A Antonio Páez y Sonia porque siempre os tendré en mente con esa canción de Rafael sonando a todo volumen jaja. Me habéis tratado genial y eso es de agradecer, y mucho. Gracias. Se te echará de menos Antonio, disfruta que lo tienes bien merecido!! En general, a todo el resto de personal con el que quizá he tenido menos trato pero son igual de importantes, gracias a todos. A Fran, Elvira y todo el resto de personal de limpieza. A toda la gente de los laboratorios, LEM: Ana, Mónica, José María y Antonio por echarme un cable siempre que lo he necesitado en mi época de laboratorio. Gracias Juanele por los consejos al inicio :D; LAST: A David por estar siempre disponible para hacer un mapa, que si ahora Doñana, que si haplotipos, que si luego introducciones... Gracias por soportarme! :P; LEF: muchas gracias Nene por echarme una mano en cualquier duda que he tenido, entrar en este laboratorio es sinónimo de buen ambiente y trabajar con una sonrisa en la cara, eres fantástico! Ánimo y no cambies por más que se tuerzan las cosas =). Y también, a la gente de seguimiento por su apoyo para el muestreo de Doñana. Y los informáticos en especial a Jesús! ¡¡¡Gracias a todos!!!
Y como una tesis no solo es leer artículos, campo, laboratorio, análisis estadísticos, escritura, congresos, presentaciones, y zascas de las revistas. Hay mucha gente a la que hay que agradecer que estos años hayan estado cargados de risas, aventuras y buen ambiente. Empezaré por la “familia laboral”, por mis primic@s: Vane, compañera de despacho, de debates transcendentales, de final de tesis, de –ic@s, ha sido un verdadero placer haber compartido estos años contigo y, aunque Murcia es la única comunidad autónoma que no haya visitado (por algo será), siempre queda pendiente ese viaje a tu tierra. El inglé, lo colorè y la florè ya no serán lo mismo para mí :D Víctor, por las risas junto a “la Vane” en el despacho salvando el mundo, y por cómo has sabido huir en esta época tan bonita que hemos tenido en el despacho, ehh?? Nos abandonaste vilmente. Gracias por darle esa vidilla siempre necesaria al despacho =) Irene, prima lejana jaja gracias por dejarme grabado ese momento happy cantando y bailando por el césped de Coimbra, es algo que será difícil de olvidar :D. Si una cosa queda clara es que, el trabajo en grupo y el razonamiento científico tiene sus frutos para la vida cotidiana, y si no, que se lo digan al pobre hombre que asaltamos como hienas hambrientas buscando desayuno jaja. Obragoda adapetados!. Y por supuesto a todo el resto del grupo: MJ, José Antonio, Alberto, Andrew, Casper, Adam, Charles, Bia,…
197 Acknowledgements/Agradecimientos
Miguel, eres un crack! Grandes risas nos hemos echado juntos, desde la pinza, hasta el Fran-CIS-co!, porque haya muchas más cervezas (no al mismo ritmo, eso es imposible). Lucía, fuiste la primera estudiante que codirigí y puedo decir que aprendí un montón, fue un auténtico placer trabajar a tu lado. Sé que todo lo que te propongas lo vas a conseguir, a por ello!. Muchas gracias a los Figuerolid@s, por haberme dejado ser ese +1 :P Por todas las cenas en casa (de Martina), por acompañarnos en la boda (de Martina), y muchos etcéteras (de Martina) jaja. Sois la leche :P Gracias Jordi por recordarme lo dura que es la vida del biólogo (me he dado cuenta al escribir estas poco más de 200 páginas :D), menos mal que se compensa con barbacoas y calçots. Ramón, gracias por estar ahí siempre tan atento de todos nosotros. Josué, eres muy grande (no tanto como Segovia, su grandeza es incalculable, todos lo sabemos), siempre dispuesto a echar una mano, lástima que seas merengón jaja. Rafael, ya son muchos años que nos conocemos, muchas risas y momentos inolvidables con y sin urbazón; por más que lo he controlado espero no haya ninguna referencia extraña por alguno de los capítulos. Y esto sería un poco todo lo relacionado con los cangrejos, si lo has entendido tú ya me quedo a gusto, porque ha quedado todo clarinete. Alazne, gracias por aportar ese toque norteño al despacho, por estar al otro lado y no saber si estamos hasta que no levantamos la mano :D. A la mini-hi.., a Jessica, no te dejes hacer bullying y dales caña a estos carrozas, a ver cuando estalla esa traca como cual Valenciana que eres, tengo ganas ya de oírla. Alberto, por deleitarnos con licor café del bueno-bueno, del que no duermes esa noche =). Y a los apegados como yo :D Merche, por esos cafés en la Alameda como tradición dominical y por enseñarnos el buen jamón en la feria de tu pueblo. Isa, por acercarme las verbenas norteñas al sur, en tu pueblo, las necesitaba. Óscar, dueño y gran señor de Santipower, gracias por esas conversaciones y por cuidar de nuestra alimentación durante días al invitarnos a una lasagna a tu casa. Sois geniales! Que haya muchas más quedadas. A todos los que acabaron esta etapa doctoral ya, Burraco, se te echa de menos cabesita, un poco de locura en este mundo siempre es agradecida, tengo la impresión que volveremos a coincidir y tener más diversión con banderas. Jesús, gracias por robarme la despedida de soltero!! Tendré que perdonarte por ser FPU aunque cobres menos $. A Noiña por enseñarme su tierra, su “sol” y perdernos (literalmente) por los bonitos alrededores de su pueblo y por sus 5 minutitos más. A Lucas, Luismi, Edu, Rosa, Goyo, Joan, Isa, Paloma, y muchos que han pasado incluso antes de que empezara con esta aventura cangrejil. Gracias! =) No me olvido del Dream-Team, la edad pasaba factura, las nuevas generaciones nos bailaban y aunque hubiera muchos disgustos (no ganábamos ni una), han caído
198 Acknowledgements/Agradecimientos muchas risas pasando más tiempo tomando cervezas que jugando, incluso quedando a jugar y tomando cervezas directamente, discutiendo estrategias que dejaban de funcionar en el minuto 1, qué desastre :D Gracias a todos por esos buenos momentos: David, Néstor, Albertos, Airam, Grego, Juan, Duarte, Carlos Camacho, Marcelo,… Y cómo no de BAMI, grandes pachangas: Xim sigo esperando tu política de fichajes y cláusula de entrada para doctorandos :D, Riki, Rubén, y muchos otros que se acercaban para darle cuatro patadas al balón :D Gracias a los que cada semana o mes o más bien de ciento a viento jugamos al padel. Entre lesionados, estancias y dejadez, lo mejor será ir directamente al post-partido :D Gracias a Jorge, Rubén, Ale, Víctor, Berni, Javis, Rosa y Grego por esos tantos interminables, partidos reñidos y pelotazos que liberan tensión después de escribir en la fase final de la tesis. A todos los doctorandos y los que no, la verdad es que nos hemos juntado un buen grupo y que siempre quedará un bonito recuerdo de esta etapa: Amparo, Álvaro, Antonio, Carlines, Eneko (a ver si echamos un mus), Dani, Fer, Hyeun-Ji, Jota, Juan, Lola, Lina, Mar, Mari, Marina y Kike, María Lucena, Marta y David, Sara, Sarai, Sol-Basti, Vary, Vero, y un larguísimo etc. Gracias Paula por esta portada, por plasmar fantásticamente las ideas que tenía en mi cabeza. =) Piliiii ya tenemos la tesis cangrejil!! Ahora a hacer cangrejos y anillamientos flamenquiles en modo póster. Gracias por ser como eres!
Estar en el Sur supone estar lejos de casa la mayor parte del año. Es siempre una gozada que, aunque vuelva de ciento a viento nunca parezca que esté lejos. Gracias a todos los amig@s de La Resaka, School-team, Sodomitas, y a mis 3 hombres + 1 y medio. A mi familia romana: grazie di cuore per farmi sentire come in casa ogni volta che vengo da voi. Y por supuesto a la familia, prim@s y mi hermana porque siempre hay ganas de juntarse y echar un vermout largo más binguico. Pero especialmente, gracias a mis padres porque ellos apostaron por mí desde siempre desde cuando dije con 14 años que quería ser Biólogo, por seguir mi sueño a mi lado sin pedir explicaciones, por buscar soluciones a cualquiera de los problemas que surgían, por tratar de ayudar y echar una mano cuando la he necesitado. Mamá, eres única, fuerte y luchadora! Gracias por transmitirme esos valores que me han hecho llegar donde estoy. Papá, ver estas líneas era una de tus siguientes metas, esta portada es aquello que comentamos justo antes y que dijiste: “ya veremos”. Echo de menos que me preguntes: “y esto de los cangrejos, ¿para qué?”, sólo para picarme. Muchas gracias porque gran parte de la persona que soy, es gracias a ti.
199 Acknowledgements/Agradecimientos
Martina, Martina, Martina. Siempre recordaré el día que me pasaste este proyecto y me animaste a ir a por él, a por el sí!! Fue probablemente una de las mejores decisiones y fue gracias a ti. Gracias por ser ese impulso que siempre necesito, por estar a mi lado en las buenas y en las no tan buenas, por aguantarme, por tener (poca) paciencia conmigo :P, por esos bailes espontáneos por casa, por decir “pues vale” después de hincar rodilla, por tantísimas cosas y por muchas otras que están por llegar… Te lo mereces todo y más! Acuérdate que siempre estaré “lo más lejos a tu lado” y aunque haya veces que haya algún kilómetro entre los dos “ya no queda nada entre tú y yo, ya no queda nada entre los dos”. Y para acabar, me gustaría hacerlo como el primer mail que me mandaste y el mensaje que dejaste en el ordenador mientras estaba fuera.
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