Effects on soil and alternatives for biological control of the invasive plant Carpobrotus edulis

   7(6('('28725$0(172 '(('%651051+.#0&#.6'40#6+8'5(14 $+1.1)+%#.%10641.1(6*'+08#5+8' 2.#06 %#421$41675'&7.+5   &ULVWLQD9LHLWHV%ODQFR    '5%1.#&'&17614#/'061+06'40#%+10#. 241)4#/#&'&17614#/'061'0/'&+1#/$+'06''4'%745150#674#+5  6$17,$*2'(&203267(/$  Instituto de In v estigaciones Agrobiológicas de Galicia Avda. de Vigo, s/n 15705 Santiago de Compostela Tfno.: 981 590 958 Fax: 981 592 504 [email protected] PROBLE MINISTERIO DE CIENCIA, INNOVACIÓN Y UNIVERSIDADES AUTORIZACIÓN DO DIRECTOR DA TESE Effects on soil and alternatives for biological control of the invasive plant Carpobrotus edulis Dr. Serafín González Prieto, Investigador Científico do CSIC adscrito ao Instituto de Investigacións Agrobiolóxicas de Galicia (IIAG-CSIC), INFORMA: Que a presente tese correspóndese co traballo realizado por Dna. Cristina Vieites Blanco, parcialmente desenvolvido nos laboratorios do Departamento de Bioquímica de Solos do IIAG-CSIC baixo a miña dirección, e autorizo a súa presentación considerando que reúne os requisitos esixidos no Regulamento de Estudo de Doutoramento da USC, e que como director desta non incorre nas causas de abstención establecidas na Lei 40/2015. En Santiago de Compostela, a 17 de xullo de 2018 Asinado: Dr. Serafín González Prieto 4 6.5. Synthesis .................................................. 215 6.6. Supplementary material ..................................... 216 6.7. References ................................................. 228 7. Conclusions ................................................... 233 8. Resumen general en castellano .................................. 239 8.1. La planta invasora Carpobrotus edulis ........................... 241 8.2. Estudio de los efectos de C. edulis en el suelo.................... 243 8.3. Potencial de Sclerotinia sclerotiorum y Pulvinariella mesembryanthemi para control biológico de C. edulis .............. 245 8.4. Distribución potencial y dinámica de población de P. mesembryanthemi en áreas invadidas por C. edulis ................ 247 8.5. Variabilidad genética de poblaciones de P. mesembryanthemi a escala mundial............................. 248 8.6. Bibliografía ................................................ 249 9. Agradecimientos ............................................... 255 1. Introduction 1. Introduction 7 1. Introduction 1.1. BIOLOGICAL INVASIONS The distribution of a species is limited by their capability to overcome physical barriers to dispersal, by their environmental requirements (e.g. climate, soil, water and food availability) and biotic constraints (e.g. predation, competition, diseases) (Boulangeat et al., 2012, Guisan and Thuiller, 2005, Mott, 2010). Species distribution is dynamic, changing with environmental conditions (which can lead to regional extinctions or expansions), evolution and modifications of physical barriers (leading to vicariance, when populations are separated, or to dispersal, when they disappear) (Holt, 2003, McNeely, 2001, Sanmartín, 2009). The limitations on species distribution has led to the high biodiversity of the planet (McNeely, 2001). Human activity significantly impacts species distribution, decreasing their range (and even causing their extinction) or increasing their distribution, either directly through transportation or indirectly due to changes in the environment and climate (McNeely, 2001, IUCN, 2017a). These species that are deliberately or accidentally introduced outside their native range are called alien species (DAISIE, 2018). From the arrival of an alien species to the new ecosystem until its consideration as invasive, three different stages can be differentiated (see Fig. 1.1 and Table 1.1 for further details): a. introduced or casual species: when they have been transported by humans outside their natural range (Richardson et al., 2000), b. naturalized or established species: they survive and regularly reproduce in the new environment, being able to sustain populations on its own (Richardson et al., 2000), c. invasive species: those whose introduction or spread threatens or negatively affects the ecosystems as indicated by the European_Parliament_and_Council (2014). It is usually assumed that around 10 % of the imported species become introduced, 10 % of the introduced species become naturalized and 10 % of these naturalized species become invasive (Williamson and Fitter, 1996) (Fig. 1.1). Cristina Vieites Blanco 8 Stage Dispersal barrier1 ALIEN Survival barrier (abiotic factors of local environment)1,5 INTRODUCED CASUAL Reproduction & biotic barriers1,5 NATURALIZED ESTABLISHED Dispersal & environmental barriers (abiotic and biotic)1 INVASIVE Process transported outside native range2 released or escaped2 established viable populations2 spread outside released area2 Advanta g eous traits or situations utility or association to humans2,5, hitchhikers5, seed longevity5, wide native range5 original from a similar environment2, generalist2, high number of individuals released2,3,5, fast growth5 asexual reproduction or hybridization3, rapid adaptation by epigenetic changes3, behavioural changes2, high genetic variation3, fast growth5, efficient resource use5 better competitor3, absence of enemies3, phenotypic plasticity5, rapid evolutionary changes3, good dispersal ability5, high fecundity5, short generation time5, habitat connectivity5, anthropo g enic disturbance5 Limitations no capacity to survive longdistance transport5, no association with humans5 no local adaptations3, absence of a suitable niche2 population bottleneck3, migratory behavior2, small populations are vulnerable to stochastic events2 no outcompetition of natives3, slow spread5, isolation5 Management prevention4 eradication4 eradication, containment4 eradication, mitigation4 Fig. 1.1. Stages of an alien species, from its introduction into a new area to its invasion: barriers to overcome, processes, positive and negative traits and situations and management options for each step. References: 1Richardson et al. (2000); 2Duncan et al. (2003); 3Allendorf and Lundquist (2003); 4Blackburn et al. (2011); 5Theoharides and Dukes (2007). 10% 10% 10% 1. Introduction 9 Table 1.1 Approximate number of naturalized species in the world and percentage of naturalized species within each group, with respect to the total number of known species. 1van Kleunen et al. (2015); 2Lever (2002); 3Lever (2003), 4Lever (2005); 5Lever (1985); 6Ahyong et al. (2018); 7The_Plant_List (2013); 8Eschmeyer and Fong (2018); 9Uetz (2018); 10IUCN (2017b); 11IUCN (2018a); 12IUCN (2018b); 13Bouchet (2006). The introduction of alien species began long ago. Humans spread outside Africa to other continents, since 100,000 years ago, has been linked to the transportation of domesticated species used for feeding or protection (e.g. dogs, pigs, wheat) and other stowaway species (e.g. rats, diseases, parasites) across their dispersal barriers (McNeely, 2001, Mack et al., 2000). With the development of trading and long-distance traveling humans’ capacity to move species outside their native range increased (McNeely, 2001), especially after late 15th century with the sailing voyages between Europe and America. The beginning of the 16th century entailed the commencement of a new period in biological invasions due to its facilitation by changes in demography, agriculture, commerce and industry and the beginning of colonialism and exploration voyages (Hulme, 2009, Preston et al., 2004). Species were transported for explorers and colonists for feeding, medicines or even nostalgic and aesthetic reasons (Mack, 2001, Lever, 2005). The introduction of alien species increased in the 19th century, with the Industrial Revolution, due to the more intensive international trade and migrations (Hulme, 2009). Military activities have also led to the introduction of species to new regions since the 19th century due to the transportation of cattle (to feed armies), plants (which were used for producing natural remedies or materials) or by inadvertently transporting other species (McNeely, 2001, Wearn, 2016). Also, to control accidentally-introduced pests, other species have been introduced to control the former, resulting in some cases in new invasions (McNeely, 2001). Cristina Vieites Blanco 10 Currently, terrestrial transport networks, migration, importations and tourism are ligated to the introduction of alien plants (Vilà and Pujadas, 2001). The globalization of the economy since the 20th century has enhanced the demand of new species (McNeely, 2001). Foreseeable, invasions will be facilitated by new intercontinental transportation networks (with higher volumes of merchandise that will travel further distances), global change, increase of human activities in the Artic due to loss of ice (such as shipping, mineral exploitations, fisheries or tourism), refugee movement and military transport (Ricciardi et al., 2017). It is expected that the use of new crops (and their pathogens), new farmed species (such as insects) and soil bacteria and fungi (to increase crop production) will entail new invasions (Ricciardi et al., 2017). Invasive species are a worldwide problem and pose one of the major threats to biodiversity (European_Commission, 2011, McGeoch et al., 2010, UNEP, 2002) as well as to the structure and functioning of ecosystems (Ehrenfeld, 2010, Hernández et al., 2014, Lake and Leishman, 2004, Le Maitre et al., 2011). Invasive plants can affect natives by modifying water and fire regimes, soil nutrients availability and ecosystem energy budgets (Mack et al., 2000). Invasive animals can affect natives by predation, competition or alteration of the invaded habitat (Mack et al., 2000). The introduced species can be sometimes very competitive due to the absence of coevolution with the native species (e.g. the brown tree snake in Hawaii) (McNeely, 2001). Since the 16th century, large extinctions of species are related to alien predators and diseases (Bellard et al., 2016, Loehle and Eschenbach, 2011), being alien species considered by some authors as the second most common cause of species extinctions (Bellard et al., 2016). However, extinctions are usually the result of a set of causes (which add to the effect of the invasion), and in some cases the contribution of alien species to extinctions might be overrated as the invasion by aliens and the extinction of natives can be triggered by the same anthropogenic disturbances (Pyšek et al., 2017). Around 27% and 62% of plant and vertebrate extinctions, respectively, are believed to be related to alien species (Bellard et al., 2016). Aliens affect more importantly amphibians, reptiles and mammals, for which the introduction of new species is the major cause of extinction, and to island endemisms (Bellard et al., 2016). Almost one fourth of the threatened or nearly threatened species are affected by invasive species (Maxwell et al., 2016). Alien invaders can also have economic, social and human health impacts (Pimentel et al., 2005, Vitousek et al., 1996). For instance, microbial pathogens’ 1. Introduction 11 invasions, favoured by tourism, global commerce and changes in climate or environmental conditions can affect biodiversity and human’s economy and health (Ricciardi et al., 2017). Pathogens can rapidly undergo genetic changes and our capacity to detect and identify them is limited (Ricciardi et al., 2017). Invasive alien species' damage has been estimated as € 12.5 billion per year in the EU (European_Commission, 2011), more than € 100 billion per year in the USA, more than € 10 billion in Australia, around € 6 billion in South Africa (Pimentel et al., 2001) and globally could reach € 1200 billion per year (Saad et al., 2009). However, damage estimations are highly variable depending on the methods used and the number of species included (Marbuah et al., 2014). In order to fight the increasingly concerning problem of invasive species, improving the knowledge on them is crucial and, consequently, several programs such as the GISP (Global Invasive Species Programme) and DAISIE (Delivering Alien Invasive Species Inventories for Europe) have been created. In addition, their control and eradication is the 5th target of the 2020 Biodiversity strategy (European_Commission, 2011). 1.2. INVASIVE PLANTS Around 4 % of the vascular plants in the Earth have become established outside their native range (van Kleunen et al., 2015). In Spain, 12% of the flora is invasive (Sanz-Elorza et al., 2004) and over € 50 million have been expended, mainly in removal of invaders, from 1997 to 2007 (Andreu and Vilà, 2007). Galicia (NW Spain) exceeds the national average, with 14% of its flora being invasive (Romero-Buján, 2007). 1.2.1. Species invasiveness and habitat invasibility That an alien plant becomes invasive or not in a given habitat mainly depends on the interaction between the characteristics of the habitat and the species (Alpert et al., 2000, Rodríguez-Echeverría, 2010). Despite the importance of the early identification of potential invaders (Gibson et al., 2011), to date much remains to learn regarding this matter (Fuentes-Ramírez et al., 2011). Plants invasiveness is commonly related with high dispersion and competition capacities, generalist characteristics and the ability to exclude natives (Lorenzo et al., 2010, Fuentes-Ramírez et al., 2011, Gibson et al., 2011, Hui et al., 2011, Daehler, 2003, Parker, 2001). Another factor of invasive Cristina Vieites Blanco 12 species success is a lower regulation of invasive populations by predators, parasites and other enemies (“enemy release hypothesis”), which derives from the lack of specialist consumers, low impact from specialist predators of native species and less pressure from generalist predators in comparison to native species (Keane and Crawley, 2002, Maron et al., 2014). Needing to expend less resources in defence, the invasive plant can allocate more resources to growth and reproduction, obtaining a competitive advantage (Callaway and Ridenour, 2004). Several experiences seem to corroborate this hypothesis, although exceptions have also been found (Keane and Crawley, 2002, Daehler, 2003). Another explanation for invaders success is the “novel weapon hypothesis”: the absence of a coevolution between invasive and native plants can be a competitive advantage for the invader, as natives may be more susceptible to allelopathic or antimicrobial substances exuded by the invader’s roots (Callaway and Ridenour, 2004). Habitats most prone to invasions, for instance islands, riverbanks and coastal areas, have usually low levels of competition, altered disturbance regimes, low environmental stress or a combination of these factors (Alpert et al., 2000, Carboni et al., 2010). Frequently, invasive plants take advantage of disturbances, particularly those that diminish competition and increase the amount of unused resources (Alpert et al., 2000, Lake and Leishman, 2004, Daehler, 2003, Werner et al., 2008). However, disturbances by themselves may not be sufficient to enhance invasion if they are not accompanied by an increase in nutrient availability (Lake and Leishman, 2004). Among the main adaptations of invasive species to increased fertility are rapid growth, higher competitive ability than natives, high leaf area and less leaf construction cost (Alpert et al., 2000, D'Antonio and Meyerson, 2002, Daehler, 2003, Lake and Leishman, 2004). These adaptations to high resources availability suggest they will be benefited by an increase in atmospheric CO2 concentration (Manea and Leishman, 2011). It has been suggested that invasibility and the deviation of the disturbance regime from the natural one may be directly related and, thus, even the suppression of a disturbance could lead to invasion in some cases (see Alpert et al. (2000) and references therein). Under low-nutrient conditions, invasive species rarely overcome natives (Daehler, 2003), although exceptions have been found (Funk, 2013). A high capacity of competition for limiting resources in resource-poor environments is associated with an easier acquisition of resources, a higher efficiency in their use and a more efficient conservation (Kurokawa et al., 2010, Werner et al., 2008, Funk and Vitousek, 2007, Morris et al., 2011, Funk, 2013, Sardans et 1. Introduction 13 al., 2016), as well as an increase in fertility through N2 fixation, rapid decomposition of litter or modification of fire regime (Funk and Vitousek, 2007, Saad et al., 2009). 1.2.2. Effects of the invasion 1.2.2.1. Effects on economy, society and human health Although in some cases the introduction of alien plants can produce economic benefits and important human services (e.g. crops), invasive species can negatively affect the economy in several ways, such as loss of ecosystem services, impacts on human health or expenses derived from the control of the invasion and the restauration of the invaded habitat (Sharma et al., 2005). Usually, the introduction of new species is not accompanied by a cost-benefit analysis (McNeely, 2001) and the economic impacts of aliens are usually not assessed due to the absence of quantitative data of their impacts or to the difficulty of including nonmarket costs (Duncan et al., 2004, Evans, 2003). Alien plants can affect human health by producing toxins [e.g. giant hogweed (Heracleum mantegazzianum Somm. & Lev.) (Nielsen et al., 2005)] or allergens [e.g. common ragweed (Ambrosia artemisiifolia L.) (Bohren, 2006)]. For instance, allergy costs due to common ragweed are predicted to be around 300-350 million € per year in the next years (Richter et al., 2013). The effects of invasive plants on ecosystem services are rarely quantified, despite of their importance for human society (Charles and Dukes, 2007). They include provisioning (food, fiber, etc.), cultural (such as recreation or cultural heritage), supporting (e.g. maintenance of nutrient and water cycles) and regulatory (e.g. regulation of air quality) services (Charles and Dukes, 2007, Vilà and Hulme, 2017). For example, negative effects on crop production have been reported by Cirsium arvense (L.) Scop. in North America, producing yield losses of around 45 % (Vilà and Hulme, 2017). 1.2.2.2. Effects on the environment Invasive plant can affect both abiotic - such as water and fire regimes and soil nutrient status, and biotic characteristics - like native species biodiversity (Brooks et al., 2004, Le Maitre et al., 2002, Enright, 2000, Ehrenfeld, 2003, Ehrenfeld, 2010, Vilà et al., 2006); in some cases enhancing their own invasion (Ehrenfeld, 2003). The effects of an invasive species can vary depending on the characteristics of the invaded ecosystem (Ehrenfeld, 2003, Dassonville et al., 2008). Cristina Vieites Blanco 20 1Mattrick (2006); 2Kettenring and Adams (2011); 3D'Antonio and Meyerson (2002); 4Yelenik et al. (2004); 5Wilson et al. (2011); 6Le Maitre et al. (2011); 7Silva and Marchante (2012); 8Wisconsin_Department_of_Natural_Resources (2016); 9Marchante et al. (2011); 10Conser and Connor (2009); 11Simmons (2005); 12Oehler (2006); 13Lebo (2007); 14Leach and Dawson (2000); 15Souza-Alonso et al. (2015); 16Weidenhamer and Callaway (2010); 17Evans (2013); 18Haubensak et al. (2004); 19Kulmatiski and Beard (2006); 20Van Driesche et al. (2008); 21Shaw et al. (2014); 22Cory and Myers (2000). 1. Introduction 21 Cristina Vieites Blanco 22 1.2.3.1. Biological control One of the hypothesis to explain invasive species success is enemy release (Keane and Crawley, 2002, Maron et al., 2014). Therefore, introducing enemies of the invasive plants can help counteracting this competitive advantage. Biological control consists mostly on the introduction of potential predators or parasites of the invasive species (Shaw et al., 2014), although it can also involve the increase or maintaining of an existing enemy (McEvoy and Coombs, 1999). It has been done for more than 100 years, especially in developed countries. Experiences in developing countries are often dependent on the research of control agents in developed countries (Shaw et al., 2014). Agents are often looked up for in the area of origin of the invasive species (Shaw et al., 2014). Once found, the safeness of its introduction in the new ecosystem is studied and, if approved, it is released in the invaded area, which should be monitored and evaluated (Shaw et al., 2014). However, there are few studies of the effectiveness of the biological control agent (Shaw et al., 2014, Morin et al., 2009, Myers et al., 2009), whilst most of the funding is spent on looking for potential agents, doing host specificity tests and releasing the agents (Myers et al., 2009). Biological control success depends on the stage of invasion, being more effective on plants in an early phase of invasion (Zimmermann and Neser, 1999). The outcome of biocontrol also depends on the target plant traits, usually being aquatic and asexual plants easier to control (Paynter et al., 2012). Biocontrol tends to also be more successful with plants that are not considered major weeds in their native area (Paynter et al., 2012). Herbivores tend to weaken their host, instead of immediately killing it (Van Driesche et al., 2010). Therefore, unlike mechanical and chemical tools, biological control may take longer to be effective (Van Driesche et al., 2010, McFadyen, 2000). The main advantages of biological control are its cost-effectiveness, its sustainability and its long-term effect, as when it is successful the species will reproduce and spread by itself after its introduction (Shaw et al., 2014, DiTomaso et al., 2017). Biological control is usually applied jointly with mechanical, chemical or other control techniques, allowing a reduction in the use of chemicals (Moran et al., 2013, Van Driesche et al., 2010, Olckers, 2004, Lake and Minteer, 2018). The efficacy of mechanical and chemical control tools without risking re-infestation is limited to small 1. Introduction 23 or isolated areas (Van Driesche et al., 2010). In contrast, biological control can be efficient in larger areas without repeated application of the treatment (Van Driesche et al., 2008). Biological control can affect ecosystems through impact on native species (Van Driesche et al., 2010). Non-target effects on native species due to a lack of specificity are more likely when there are native congeneric relatives of the invasive plant (Van Driesche et al., 2010, Louda et al., 2003). However, non-target effects are rare and the perceived risks seems to be frequently magnified in a subjective way (Seastedt, 2015). Most of these undesirable effects are due to introductions performed decades ago (Strong and Pemberton, 2000, Simberloff and Stiling, 1996), some of them targeting native plants (Seastedt, 2015). To decrease the risk of non-target attacks, it is recommendable to take into account the environmental conditions of the treated area and the dispersal capacities and adaptive change of biocontrol agents, to look for non-target indirect effects, and to do long-term studies and life table analysis (Louda et al., 2003). Also, studying the genetic diversity of control agents can optimize the control, by choosing the most adequate population, which can vary according to the environmental conditions of the area (Gaskin et al., 2011, Rauth et al., 2011). 1.2.3.2. Restoration of invaded habitats To facilitate the recovery of the previous ecosystem properties, several actuations can be performed. According to the Society for Ecological Restoration, restoration can be defined as “the process of assisting the recovery of an ecosystem that has been degraded, damaged, or destroyed” (SER, 2016). Restoration includes abiotic manipulation (Le Maitre et al., 2011) and regeneration with native species by plantation or sowing (Richardson et al., 2007). Once the species is introduced the possibility of restoring the affected area must be considered (Myers et al., 2000) because removing the exotic species might not be enough due to the legacy effects produced in the ecosystem (D'Antonio and Meyerson, 2002, Nsikani et al., 2018). However, the high cost involving restoration makes necessary to select areas that are most affected or where management has the highest probability of success (D'Antonio and Meyerson, 2002). Management success depends on species characteristic such as reproductive biology, size and persistence of the seed bank or propagule dispersal (Panetta, 2009, Myers et al., 2000), as well as habitat properties, Cristina Vieites Blanco 24 decreasing with higher nutrient concentration (Daehler, 2003). Furthermore, the longer the area has been invaded, the more complicated the restoration might be, as the seed bank becomes larger (D'Antonio and Meyerson, 2002), changes in soil properties and microbial community are more important (Marchante et al., 2008, Souza-AlonsoGuisande-Collazo et al., 2015) and invasive species distribution widens (Myers et al., 2000). Restoration efficiency depends as well on the understanding of the processes by which invasive species alter the environment (Le Maitre et al., 2011). Some restoration labours might even enhance exotic species invasion through disturbances of the area, leading to secondary invasions (D'Antonio and Meyerson, 2002, Le Maitre et al., 2011). Consequently, prevention of further invasions must be made (Myers et al., 2000). The complexity of controlling invasion and restoration side effects highlights the importance of taking a preventive approach. 1.3. CARPOBROTUS EDULIS AS AN INVADER 1.3.1. Biology and native distribution of the species The genus Carpobrotus (Phylum Magnoliophyta; Class Magnoliopsida; Order Caryophyllales; Family Aizoaceae; Subfamily Ruschioideae) comprises approx. 13 species of perennial succulent plants original mostly from South Africa and Australia (Waycott, 2016, Malan and Notten, 2006). Carpobrotus species are prostrate plants with creeping branches, extensive root system, sharp 3-angled green (or purple tinged) leaves, solitary flowers with 3-5 petals (purple, pink, yellow or white) on erect and short shoots, and fleshy indehiscent fruits (Hartmann, 2002). Plants from this genus grow mostly in coastal areas (although they can be also found in sandy to marshy areas inland) with Mediterranean or temperate climates (Hartmann, 2002). They can grow in dry and salty soils of preferentially sunny areas, but they cannot tolerate temperatures below -6 °C (GEIB, 2006). In drought or salt stress situations, they can induce crassulacean acid metabolism (CAM) (Robert et al., 2013). Carpobrotus edulis (L.) N.E. Br. (1926) and Carpobrotus acinaciformis (L.) L. Bolus (1927) [commonly called ice plant, Hottentot fig, Cape fig, sour fig (Eng.); uña de gato (Esp.); herba do coitelo (Gal.)] are both South African species widely naturalized in Mediterranean climates. The native 1. Introduction 25 range of Carpobrotus edulis includes both coastal and inland slopes of Northern (from Namaqualand), Western and Eastern Cape (Malan and Notten, 2006). It is easily recognized by its yellow flowers, whilst all other Carpobrotus species have pink/purple flowers (Waycott, 2016). Carpobrotus acinaciformis inhabits coastal sand dunes in Western Cape, from Saldanha to Mossel Bay, where it is native (Malan and Notten, 2006). Besides for the flowers colour, it can be differentiated from C. edulis in the shape of the leaves (the cross section of the middle of the leaves is equilateral in C. edulis and isosceles in C. acinaciformis) (Gonçalves, 1990). Carpobrotus edulis can hybridize with other Carpobrotus species (Weber and D'Antonio, 1999, Waycott, 2016), including C. acinaciformis, both in South Africa (Wisura and Glen, 1993) and in the introduction areas (Suehs et al., 2004), or even with other genera from the Aizoaceae family (Chinnock, 1972). The introgressed hybrid complex between C. edulis and C. acinaciformis is cited as C. affine acinaciformis (Suehs et al., 2004, Delipetrou, 2006). 1.3.2. Introduction and invasiveness Introductions of Carpobrotus species have been mainly of C. edulis and C. aff. acinaciformis (Table 1.3). In Europe, ice plants were firstly introduced in the 17th century (1680-1690) through Holland, Belgium and England (Wisura and Glen, 1993, Mulder, 2003, Robert et al., 2013). Widespread introductions and naturalization along the Mediterranean coast began at the end of the 19th century (Robert et al., 2013, Badalamenti et al., 2016). In Spain, the first record was in the NW coast, in Baiona in 1892 (Lázaro- Ibiza, 1900). In the United States C. edulis is present since its introduction in California in the 1900s (D'Antonio, 1993). References for Table 1.3 (next page): 1Marchante et al. (2014); 2Sanz-Elorza et al. (2004); 3FCBN (2018); 4Vilà et al. (2006); 5Julve (2017); 6Flora_Italiana (2018); 7Seljak (2010); 8Stancic et al. (2008); 9Trinajstic (1998); 10Stešević et al. (2017); 11Arianoutsou et al. (2010);12Washburn and Frankie (1985); 13Robert et al. (2013); 14Dufour-Dror (2012); 15Dufour-Dror (2013); 16Livshits et al. (1988); 17Delipetrou (2006); 18EDIT (2018); 19Vivrette (2012a); 20Vivrette (2012b); 21EDDMapS (2018); 22Wunderlin et al. (2018); 23Gobierno_Mexico (2018); 24Naturalista (2018); 25Tropicos (2018); 26Ríos et al. (2010); 27Sotes et al. (2015); 28Couto and Cardoso (2018); 29Parker (2008); 30Atlas_of_Living_Australia (2018); 31Florence (2004); 32Meyer (2000). Cristina Vieites Blanco 26 1. Introduction 27 Alien Carpobrotus species are present in temperate coastal areas of Eurasia, America, Africa and Oceania (Table 1.3). They colonize dunes, cliffs and disturbed environments (Campos et al., 2004, Maltez-Mouro et al., 2010, D'Antonio, 1993), including burnt areas (D'Antonio et al., 1993). In Europe, they are primarily naturalized in the coastal areas of Southern Europe, both in the Atlantic façade and the Mediterranean basin (see Table 1.3). Carpobrotus edulis is also present in central and Northern Europe, although in a lower extent (Table 1.3). Introductions of Carpobrotus edulis and its hybrid C. aff. acinaciformis can respond to different purposes: a. Ornamental (Campos et al., 2004, Maltez-Mouro et al., 2010, Sanz-Elorza et al., 2004), being this one of its major attractiveness. b. Prevention of erosion by stabilizing embankments (Chenot et al., 2018), dunes and slopes (Campos et al., 2004, Maltez-Mouro et al., 2010), being extensively used in highways cuts in California (Donaldson et al., 1978). However, their shallow roots limit its utility in steep unstable slopes (Pierce, 1994). c. Medical use, using leaves for its antibacterial (van der Watt and Pretorius, 2001, Chokoe et al., 2008) or antifungal (Omoruyi et al., 2014) activity. They are used as a laxative or lotion for scalds and burns, to treat dysentery, as gargle for sore throats or to treat jellyfish stings (Pierce, 1994). d. Culinary use, as jam can be made from its fruits (Pierce, 1994) and leaves can be used to preserve food (Omoruyi et al., 2014). e. Fire prevention, using the plants as firebreaks, which was advised for protecting houses in California by the Fire Department (Pierce, 1994). The ice plant C. edulis and its hybrid C. aff. acinaciformis are considered among the most invasive alien species in Europe (DAISIE, 2018) and Spain (GEIB, 2006). The higher competition capacity of the ice plant compared to natives is attributed to different characteristics: a. Dense mats formation (Maltez-Mouro et al., 2010) of up to 50 cm in depth (Ruffino et al., 2015), which leads to the elimination of native species by competition (Campos et al., 2004). b. Rapid growth (approx. 40 cm per year), which facilitates the colonization of new habitats (Traveset et al., 2008b). Cristina Vieites Blanco 28 c. Inhibition of native plant germination, apparently through release of allelopathic substances during necromass decomposition (Novoa et al., 2012). d. Reproduction both by sexual means, with a production of >1000 seed per fruit for C. edulis and >350 seeds per fruit for C. aff. acinaciformis (Suehs et al., 2004) that have and average viability of 2 years (D'Antonio, 1990b); and asexual means, with rooting nods (Delipetrou, 2006). e. Endozoochory in the invaded areas, being dispersion and germination increased through scarification by rabbits, rats or deers (D'Antonio, 1990b, Novoa et al., 2012, Bourgeois et al., 2005). f. Clonality with capacity of labour division in patchy environments (Roiloa et al., 2014) and putative belowground communication that can promote compensation responses to herbivory (Rodríguez et al., 2018). g. Rapid adaptive evolution in the invaded areas (Roiloa et al., 2016). h. Higher competitiveness with low or moderate soil salinity, which promotes shoot elongation (Varone et al., 2017). i. Different traits than native plants, which can help to explain the ecological effects of the ice plant (Badalamenti et al., 2016). j. Morphological plasticity to light availability (Fenollosa et al., 2017), which allows the ice plant to rapidly colonize heterogeneous habitats (Traveset et al., 2008b). 1.3.3. Effects of the invasion Carpobrotus edulis and its hybrid C. aff. acinaciformis are considered ecosystem engineers, as they can alter both the abiotic and biotic properties of the ecosystem, turning these changes to their favour (Conser and Connor, 2009, Molinari et al., 2007). Carpobrotus edulis invasion seems to be influenced by the habitat (D'Antonio, 1993), so its effects on the ecosystem may be context-specific (Vilà et al., 2006). As the effects of Carpobrotus spp. on the invaded ecosystems can remain even after the removal of the plant, they can significantly influence restoration success (Conser and Connor, 2009). 1. Introduction 29 1.3.3.1. Effects on biodiversity and ecosystems The invasion of ice plants alters the structure of the native plant community towards randomness, similarly to the effect of other disturbances (Santoro et al., 2012). The -species richness and the diversity decrease with the invasion (Badalamenti et al., 2016, Jucker et al., 2013). In invaded dunes, the net productivity of the ecosystem can decrease (Maltez-Mouro et al., 2010). Ruderal plant species appear to be favoured by the invasion (Santoro et al., 2012), whilst the perennial community of transition dunes seems to be the most affected due to the higher propagule pressure of the ice plant in the transition dune (Carboni et al., 2010). Also, plant species related to grasslands are scarce in invaded areas (Badalamenti et al., 2016). Carpobrotus edulis can affect pollination of native plants through competition or facilitation, depending on the plant species and varying with the ecological conditions and time (Moragues and Traveset, 2005, Vilà et al., 2009). Ice plants can also decrease germination, survival, growth and reproduction of native plants (Conser and Connor, 2009, de la Peña et al., 2010). Carpobrotus edulis hybridizes with native Carpobrotus species in Australia (hybrid with Carpobrotus rossii (Haw.) Schwantes) (Waycott, 2016) and in California (hybrid with the supposedly native C. chilensis) (Vilà et al., 1998, Vilà and D’Antonio, 1998). Hybridization with native plants can lead to genetic contamination of the native populations, hybrid vigour or a change in the ecological function of the Carpobrotus plants (Waycott, 2016). The invasion by ice plant can increase both bacterial and fungal biomass, affect the structure of soil microbiota and favor fungal growth (Badalamenti et al., 2016), favouring in turn the invasion (de la Peña et al., 2010). Invasions by Carpobrotus spp. can affect animal communities. It can decrease the occurrence of some reptile species (e.g. Chalcides striatus in NW Spain) due to a lower suitability of the characteristics of the invaded habitat (Galán, 2008), and reduce both abundance and species richness of insects through a decrease in microhabitat heterogeneity (Orgeas et al., 2007). As the ice plants produce fruits when food is scarce for herbivorous animals in California (at the end of the dry season, when there are no annuals, grasses or forbs and the perennial plants are under water stress), Cristina Vieites Blanco 36 1.3.4.3.2. Sclerotinia sclerotiorum as a potential biocontrol agent Sclerotinia sclerotiorum (Phylum Ascomycota, class Discomycetes, order Helotiales, family Sclerotiniaceae) is a cosmopolitan fungus from temperate areas, which has a broad host range. It can infect more than 500 plant species worldwide, including trees, shrubs and grasses (Saharan and Mehta, 2008a). The fungus S. sclerotiorum produces airborne ascospores, which are the primary form of inoculum, and infects plants preferentially through petals, early symptoms of the disease being usually brown lesions and cottony patches of mycelium (Saharan and Mehta, 2008b). In decaying parts of the infected plant, S. sclerotiorum can produce sclerotia (able to survive 5-10 years in the soil), which can form apotecia and produce ascospores (Saharan and Mehta, 2008b) (Fig. 1.3). Many studies have addressed the potential of S. sclerotiorum to control invasive plants, especially in New Zealand (see Table 1.5). To favour the infection, the mycoherbicide should be applied when moisture is high (Saharan and Mehta, 2008c). Also, pathogenicity can be enhanced by increasing the production of oxalic acid in S. sclerotiorum by adding sodium succinate to the growth media (Cessna et al., 2000, Briere et al., 2000). Although in some cases this fungus was able to seriously damage weeds, the high cost of its use discouraged its implementation (Saharan and Mehta, 2008c). The dispersal of S. sclerotiorum ascospores, which is essential to determine the safety zone for its application as a mycoherbicide, is approximately 250 m depending on wind speed (BourdôtBaird et al., 2006, De Jong et al., 2002). From S. sclerotiorum, the mycoherbicide HyakillTM was developed to control the invasive water hyacinth (Eichhornia crassipes (Martius) Solms Laubach), although it was not admitted for commercialization by the European Patent Office (apparently due to the low specificity of S. sclerotiorum) (Dagno et al., 2012). A related species, Sclerotinia minor Jagger, is the base of a commercial mycoherbicide to control dandelion (Taraxacum officinale Webber) (Dagno et al., 2012). Although S. sclerotiorum can infect some Aizoaceae species (Cother, 2000, Saharan and Mehta, 2008a), its potential as a control agent has not been studied yet. One main advantage of its application is that its use as a mycoherbicide would not mean the introduction of a new alien species, as 1. Introduction 37 Fig 1.3. Cycle of Sclerotinia sclerotiorum based on the infection of cultivated plants. The formation of sclerotia has not been found in diseased tissues of Carpobrotus edulis (pers. obs.). Modified and adapted to C. edulis from Johnson and Atallah (2014). Cristina Vieites Blanco 38 the fungus is present worldwide. Also, it has already been widely studied as a biocontrol agent in other species, so the mycoherbicide formulation and the application procedures have already been optimized. 1.3.4.3.3. Pulvinariella mesembryanthemi as a potential biocontrol agent The South African P. mesembryanthemi (Phylum Arthropoda, class Insecta, order Hemiptera, family Coccidae), commonly called cottony pigface scale, is a highly monophagous parasite of Carpobrotus spp. (Washburn and Frankie, 1985). The accidental introduction of infested iceplants outside their native range seems to be the reason of the presence of P. mesembryanthemi in temperate coastal areas from America, Africa, Eurasia and Oceania (Table 1.6). The scale insect P. mesembryanthemi reproduces exclusively by parthenogenesis, although in occasions males are produced (apparently as a relictic feature) (Pesson, 1941, Washburn and Frankie, 1985). Mature females produce ovisacs (waxy sacs with up to 2500 eggs) (Washburn and Frankie, 1985). The new-born scales are mobile and they look for suitable feeding sites, where they settle losing mobility (Washburn and Frankie, 1985). For long-distance dispersal, they can be transported by wind (reaching up to 190 km per generation) or zoochory (e.g. humans, dogs) (Washburn and Frankie, 1981). Immature scales go through three instar before maturing (forth instar) and producing ovisacs (Washburn and Frankie, 1985). In California, P. mesembryanthemi has two cycles per year (bivoltine), although in some favourable conditions (warmer temperatures) it can have 3 cycles per year. In the laboratory its fastest growth is reached at 24.5 °C with 3-4 months per generation (Washburn and Frankie, 1985). Growth rate of P. mesembryanthemi is also favoured by higher nitrogen and water availability for the host plant (Washburn et al., 1987). The South African Pulvinaria delottoi is a semi-cryptic species, with very little morphological differences with P. mesembryanthemi (e.g. the lateral setae, curve and shorter in P. mesembryanthemi and straight and longer in P. delottoi). Both species differ in feeding habits (P. mesembryanthemi feeds preferentially on younger leaves and P. delottoi prefers older leaves) and life cycle duration (P. delottoi only has 1 cycle per year in California) (Washburn and Frankie, 1985, Washburn and Frankie, 1981). Outside its native range, P. delottoi has only been described in the References: 1Cother (2000); 2Mortensen and Hogue (1992); 3Bourdôt and Harvey (1994); 4Green et al. (1993); 5Verkaaik et al. (2004); 6Pottinger et al. (2008); 7Bourdôt et al. (2006); 8Waipara et al. (1993); 9Waipara et al. (2006) 3. 1. Introduction Cristina Vieites Blanco 40 United Kingdom (Salisbury et al., 2011) and California (Washburn and Frankie, 1985). In the 70s in California, both P. mesembryanthemi and P. delottoi produced considerable damages in introduced populations of C. edulis, especially in heavily infested plants (Washburn and Frankie, 1985, Donaldson et al., 1978). Their success can be explained by suboptimal growth-conditions of ice plants (drought and low temperatures) and predator release (Washburn and Frankie, 1985, Donaldson et al., 1978). At the time, the scale insects were considered an economic pest for its capacity of producing decline and death on ice plants and they released predators and parasites to control P. mesembryanthemi (Washburn et al., 1985, Tassan et al., 1982). However, the use of these scale insects as biological control agents was considered for coastal habitats (Washburn and Frankie, 1985) and their potential as biological control agent outside the United States has been repeatedly suggested (ISSG, 2008, Fagúndez and Beiras, 2007). Nevertheless, their suitability as biocontrol agents of Carpobrotus spp. has not been studied so far. Pulvinariella mesembryanthemi has three clear advantages to be used as a biocontrol agent of Carpobrotus spp.: a. It feeds preferentially and almost exclusively on Carpobrotus spp.; b. It has already been introduced accidentally with the plant in areas where Carpobrotus spp. is invasive, so its use as a control agent would not entail the introduction of a new species in invaded ecosystems; c. It can produce massive death in C. edulis, as recorded in the 70's in California. However, the existence of mechanism of defence in Carpobrotus spp. can threaten control success. Death of Carpobrotus plants or plant parts with high densities of scales is considered a major scale mortality factor (Washburn et al., 1985). As scale insects are sessile (except at the crawlers stage), this could result in mortality of them and less available feeding sites for the next generation. Moreover, Carpobrotus avoids the formation of new tissues when supporting high scale densities, which limits population growth (Washburn et al., 1985). 1. Introduction 41 Table 1.6. Countries where the presence of Pulvinariella mesembryanthemi has been recorded outside its native range. 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Objectives Carpobrotus edulis (ice plant) invades dunes, rocky areas and cliffs all along the Galician coast (NW Iberian Peninsula). Extensive invasions of this plant are also found in other temperate coastal areas around the world. Ice plants engineer the invaded ecosystems altering their biotic and abiotic properties, in some cases persistently after plant removal, facilitating their own invasion. Therefore, studying the effects of C. edulis in the invaded ecosystem can help to discern the mechanisms by which this alien plant outcompetes the native vegetation and may help to optimize habitat restoration. Especially variable are the effects of the invasion on soil properties, as they seem to be dependent on the initial characteristics of the invaded habitat. Also, studies on the chemical and physicochemical effects of C. edulis on the soil have been mostly done in dunes and focused on pH and macronutrients. There is an information gap on the effects of the invasion on micronutrients, despite their importance for plant nutrition or toxicity. Furthermore, the effects on the N cycle have not been studied yet. Knowing how C. edulis affects the different soil N fluxes can help to disentangle how this alien plant alters the availability and concentration of one of the most commonly limiting nutrients for plants. Another loose end in the study of C. edulis is the search for alternatives to mechanical and chemical control. Ice plants usually colonizes vulnerable habitats, such as dunes and coastal areas, where it is primordial to avoid eroding the soil and affecting native species (especially those endemic or endangered). Also, invaded cliffs have difficult access for applying mechanical or chemical control to C. edulis and can serve as dispersion sources of the plant. For these reasons, it is necessary to find out an alternative of control which does not produce disturbances in the habitat and that is self-sustainable, such as biological control. Carpobrotus edulis has few enemies on its invaded areas, so introducing a new parasite or predator could counteract this competitive advantage over native plants. The fungus Sclerotinia sclerotiorum is one potential candidate for biocontrol because its use as a mycoherbicide has already been studied for several invasive species. Moreover, it can infect some species of the same family as C. edulis (Aizoaceae), although its suitability to control C. edulis has never been tested. Another potential biocontrol agent is the scale insect Pulvinariella mesembryanthemi, a very specific parasite of Carpobrotus species that can produce the mortality of Cristina Vieites Blanco 70 the plant at high densities. It has been accidentally introduced with C. edulis in different areas worldwide where the plant is invasive. As the abiotic and biotic interactions of P. mesembryanthemi in these alien areas are conditioning its abundance and phenology, their study can give vital information about its suitability as a biocontrol agent in the introduced areas. Characteristics of P. mesembryanthemi like performance under different environmental conditions or impact on the plant can vary across populations due to genetic differences, hence selecting the most appropriate populations would improve the success of P. mesembryanthemi as biocontrol of C. edulis. Taking all these considerations into account, the following objectives were proposed: ‐ To evaluate the effects of C. edulis invasion on the chemical and physico-chemical properties of necromass and soils (0-5 cm and 5- 10 cm depth) of invaded and non-invaded dunes and rocky areas in NW Iberia (Chapter 3). ‐ To determine the impacts of C. edulis invasion on the soil N cycle (gross N fluxes) of invaded and non-invaded soils (0-5 cm and 5- 10 cm layers) of rocky areas from NW Iberia (Chapter 3). ‐ To assess the effects of the fungus S. sclerotiorum and the insect P. mesembryanthemi, separately or combined, on the mortality and performance (growth, biomass and physiological indexes) of native and non-native (NW Iberia) C. edulis plants (Chapter 4). ‐ To study the effect of abiotic (temperature, precipitation, irradiation) and biotic (host, predators, parasites, mutualists) factors on the population dynamics and phenology of P. mesembryanthemi in NW Spain (Chapter 5). ‐ To determine the intra- and inter-populations genetic variability of native and non-native P. mesembryanthemi, as well as the original population and number of colonizing events of the alien populations through a worldwide phylogeographic analysis (Chapter 6). 3. Effects of Carpobrotus edulis invasion on main litter and soil characteristics and on soil gross N fluxes This chapter (text, tables and figures) is based on the following journal articles1,2: Vieites-Blanco, C. & González-Prieto, S. J. (2017) Effects of Carpobrotus edulis invasion on main litter and soil characteristics in backdune and rocky coastal habitats with oceanic climate. Plant and Soil, 425, 363-374. doi: 10.1007/s11104-018-3598-5 Vieites-Blanco, C. & González-Prieto, S. J. (2018) Effects of Carpobrotus edulis invasion on soil gross N fluxes in rocky coastal habitats. Science of The Total Environment, 619-620, 966-976. doi: 10.1016/j.scitotenv.2017. 11.154 1According to the "Copyright Transfer Statement" of Springer: "The author retains the right to use his/her article for his/her further scientific career by including the final published journal article in other publications such as dissertations and postdoctoral qualifications provided acknowledgement is given to the original source of publication." 2According to the "Journal author rights" of Elsevier: "Authors transfer copyright to the publisher as part of a journal publishing agreement, but have the right to: Share their article for Personal Use, Internal Institutional Use and Scholarly Sharing purposes, with a DOI link to the version of record on ScienceDirect (and with the Creative Commons CC-BY-NC- ND license for author manuscript versions)" 72 3. Effects of Carpobrotus edulis invasion on litter and soil 73 3. Effects of Carpobrotus edulis invasion on main litter and soil characteristics and on soil gross N fluxes 3.1. INTRODUCTION Soils are the base of all terrestrial ecosystems and where the different compartments [biosphere, detritosphere (dead organic matter), geosphere, hydrosphere and atmosphere] interact (Voroney, 2007). Therefore, it is essential to study the impact of invasive plants on soil properties. This research is particularly needed for C. edulis, as the changes it produces in soil characteristics can benefit its own invasion (Conser and Connor, 2009) and remain after C. edulis removal, hindering restoration (Novoa et al., 2013) and leading to secondary invasions by ruderal nitrophilous plants (Santoro et al., 2011, Novoa et al., 2013, Santoro et al., 2012). The most studied soil properties in C. edulis invaded areas have been organic matter, salinity and humidity, which increase; as well as pH and inorganic N, which are differently affected depending on the invaded area (Vilà et al., 2006, Novoa et al., 2013, Novoa et al., 2014, Conser and Connor, 2009). Less studied have been other macronutrients, for which increases (P, Na, K), decreases (Ca, Na, Mg) or no significant changes (P, Ca, Na, Mg) have been observed (Novoa et al., 2014, D'Antonio, 1990, Winsemius, 2013). As far as we know, the effect of C. edulis on micronutrients and trace element availability has not been studied until now, despite their importance for the soil-plant system as essential nutrients or toxic elements above a given threshold (Williams and Fraústo da Silva, 2000). While Cr, I, Se, Si, V and W can also be essential for some organisms, the most widespread and important micronutrients, listed in a decreasing order of concentration in living organisms, are: Fe > Mn > Zn, B > Cu >> Co, Mo and Ni (Hoppert, 2011, Grusak et al., 2016, Williams and Fraústo da Silva, 2000). As constituents of metalloenzymes and other proteins, these eight elements are essential in metabolic processes as respiration, photosynthesis or N2 fixation (Hoppert, 2011, Grusak et al., 2016, Williams and Fraústo da Silva, 2000), which play a key role even at the ecosystem level. Likewise, the use of plant and soil 13C and 15N isotopic signatures has been scarce in plant invasions despite being a basic and powerful tool in environmental sciences that provides valuable information on water, C and N cycles (Dawson et al., 2002, Robinson, 2001). Although these analyses have been used in some studies of other invasive species (Li et al., 2017, Msanne et al., 2017), and despite being C. edulis a facultative CAM (Crassulacean Acid Cristina Vieites Blanco 74 Metabolism) species with a 13C isotopic signature different from that of C-3 plants (Herrera, 2009), we have not found references on C and N isotopic signatures in C. edulis invaded areas. Of particular importance are the invaders’ effects on the N cycle, as N is the most widespread limiting nutrient in ecosystems (Vitousek and Howarth, 1991, Galloway et al., 2004). Impacts of invasive plants on the soil N cycle can persist after their removal (Elgersma et al., 2011). This is particularly true when invasions involve changes in microbial communities (Elgersma et al., 2011) or in N stock and availability, affecting the recolonization patterns and the restoration of the ecosystem (Corbin and D'Antonio, 2004), and increasing the risk of secondary invasions by ruderal nitrophilous plants (Novoa et al., 2013, Santoro et al., 2012, Santoro et al., 2011). Invasive plants can modify the soil N cycle through their impacts on soil microbial communities, litter decomposition and soil properties (Wang et al., 2015, Schaeffer et al., 2012, Laughlin, 2011). These changes can be driven by differences in phenology, leaf traits, plant litter composition, N use, N residence time, interaction with herbivores, symbiosis with native or co-introduced N2-fixing bacteria, effects on soil microbiota structure and activity, and effects on the microclimate (Mack and D'Antonio, 2003, Knops et al., 2002, Laughlin, 2011, Laungani and Knops, 2009, Corbin and D'Antonio, 2004, Castro-Díez et al., 2014, Lee et al., 2017). The modifications of the N cycle can be influenced by the characteristics of the invaded site and the invasive species, being often stronger in mild climates and islands (Castro-Díez et al., 2014) and when the traits of the invasive plant differ from those of the native flora (Castro-Díez et al., 2014, Lee et al., 2017). As net N fluxes are the result of several counteracting processes (Murphy et al., 2003), they do not adequately reflect the impacts of invasive plants (Piper et al., 2015). Therefore, studies on the effect of invasive plants on the gross N fluxes are necessary to fully understand how the invasion is disturbing the N cycle. However, these studies are scarce and mostly focussed on annual grasses (Booth et al., 2003, Hawkes et al., 2005, Parker and Schimel, 2010, Piper et al., 2015, Schaeffer et al., 2012, Stark and Norton, 2015), with only a few exceptions for perennial grasses (Thorpe and Callaway, 2011), leguminous plants and trees (Laungani and Knops, 2012). The impact of C. edulis on gross N fluxes has not been studied yet, despite having this alien invasive species site-dependent effects over N compounds (Novoa et al., 2014, Santoro et al., 2011), and being still controversial the interpretation of the processes involved in these changes. 3. Effects of Carpobrotus edulis invasion on litter and soil 75 Carpobrotus edulis can colonize dunes, cliffs and disturbed environments (Campos et al., 2004, Maltez-Mouro et al., 2010, D'Antonio et al., 1993), including burnt areas (D'Antonio et al., 1993). Habitat characteristics seem to condition the impact of C. edulis over soil properties, since induced changes of the invasive plant on some soil characteristics vary between sites (Molinari et al., 2007, Novoa et al., 2014), as it has also been reported for other invasive species. Poorer soils, with low content in nutrients and organic matter, appear to be more sensitive and where soil properties will change more (Santoro et al., 2011, Novoa et al., 2014), and most studies of C. edulis effects on soil properties focus on dunes. However, the other frequently invaded habitats, coastal cliffs and rocky areas, have been less studied. Therefore, the aim of this chapter was to compare C. edulis invaded and non-invaded areas to evaluate the effects of C. edulis on: a) the main characteristics of the necromass and topsoil; and b) the soil gross N rates. For the first part, we measured 24 topsoil (0-5 and 5-10 cm layers) variables in two dunes and two rocky areas from the Atlantic coastline of NW Iberia: water holding capacity (WHC), humidity, pH, electrical conductivity (EC), NH4+-N, NO3--N, 13C, 15N, organic C, organic N, and available Al, B, Ca, Co, Cu, Fe, K, Mg, Mn, Mo, Na, Ni, P and Zn. Also, in necromass, the total content of the latter 16 elements was measured as well as their amounts on a surface basis and the 13C and 15N isotopic signatures. For the second part, we used a paired 15N labelling experiment and the state-of-the-art Ntrace compartment model (Müller et al., 2007) to estimate up to 8 gross N rates in the two rocky areas. 3.2. MATERIAL AND METHODS 3.2.1. Site and sampling description Among 23 sites with well-established C. edulis populations (8 on dunes; 15 on rocks), two back dune and two rocky representative study sites of the areas invaded by C. edulis were selected across the Atlantic coastline of NW Iberia (Table 3.1). In September 2015, for each location, necromass and soil (0-5 and 5-10 cm depth) were separately sampled in 10 randomly distributed 15x15 cm squares under native vegetation and another 10 under C. edulis. Soil subsamples were mixed into a composite sample per site and depth and the same was done with the necromass. In the laboratory, necromass was dried at 55 ºC, weighted, triturated and homogenized. The soils were sieved (< 2 mm), homogenized and then divided into fresh subsamples, which were kept Cristina Vieites Blanco 76 at 4 ºC for inorganic N measurements and the paired 15N labelling experiment, and air-dried subsamples for the other analyses. Sub-samples of dried necromass and soils were finely ground (< 100 μm) for chemical analysis in a planetary ball mill (Retsch PM100, Germany, with cups and balls of zirconium oxide). 3.2.2. Soils and necromass analysis Soil water-holding capacity (WHC) was measured in a Richards' membrane-plate extractor at a pressure of 10 kPa. Soil pH was measured in a 1:2.5 soil:solution ratio, both in water and 0.1 M KCl, with a pH-meter (Metröhm, Switzerland). Electrical conductivity (EC) was measured in soil extracts (1:5 soil:water ratio) with an EC meter (Metröhm, Switzerland). Soil humidity was determined by drying soil samples at 105 ºC for 5 h. 3. Effects of Carpobrotus edulis invasion on litter and soil 77 Total C and total N of soils and necromass, as well as their 13C and 15N isotopic signatures, were measured in ground samples with an elemental analyser (Carlo Erba, Milano, Italy) coupled on-line with an isotopic ratio mass spectrometer (Finnigan Mat, delta C, Bremen, Germany). In the back dune soils, organic C and organic δ 13C were also determined after CaCO3 removal with the 'capsule method' using 20% HCl (Brodie et al., 2011). An elemental reference material (Soil 3 from Eurovector, Milano, Italy) and isotopic standards [IAEA-C-6 and IAEA-CH-7 (for δ 13C) or IAEA-N1 and IAEA-N2 (for δ 15N), alternately, from the International Atomic Energy Agency, Vienna, Austria] were included in each set of 10 samples to check the accuracy of the results; if necessary, drift correction was made against internal standards during the run. Inorganic N species were extracted with 2 M KCl (1:5 soil:solution ratio). The mixture was shaken for 1 h and filtered through glass microfiber filters (Whatman GF/A, Ø 125 mm). Ammonium and nitrate were sequentially liberated with two consecutive microdiffusions (55 ºC, 72 h) from 50 mL aliquots placed in 500 mL glass jars, by adding respectively MgO (0.2 g) and MgO (0.2 g) plus Devarda’s alloy (0.4 g). Both N forms were trapped as NH3 into 10 mL of 0.004 M H2SO4 in a Teflon bottle suspended in the glass jar. Measurement was made by back titration of the H2SO4 excess with 0.004 M NaOH. Three blanks and three standards (NH4NO3) were included in each batch to subtract N from reagents and to check for N recovery. Soil available nutrients and trace elements (Al, B, Ca, Co, Cu, Fe, K, Mg, Mn, Mo, Na, Ni, P and Zn) where extracted with a mixture of 1 M NH4Ac and 0.005 M DTPA (soil:solution ratio of 1:5). The mixture was shaken for 2 h, filtered through cellulose paper (Filter-Laboratory 1242, Ø 90 mm) and analysed with a simultaneous ICP-OES (Varian Vista Pro, Mulgrave, Australia). A calibration curve prepared with certified standards of all elements was measured beforehand and one of the calibration solutions was routinely included in each set of 30 samples as a quality control and, when necessary, the calibration curve was measured again. In order to measure the total nutrient and trace element content of necromass, aliquots of 500 mg were digested in a high performance digestion unit (Milestone 1200 Mega, Sorisole, Italy) for 55 min with 8 mL of 65% HNO3 and 25 mL of 30% H2O2. Blanks and reference materials (hay powder No. 129, Community Bureau of References, EU; apple leaves No. 1515, National Institute of Standards and Technology, USA) were also included in each digestion batch to subtract elements from reagents and to check for element recovery. Once cooled, the Cristina Vieites Blanco 84 Fig. 3.2. Score plots from the principal component analyses (PCA) performed with soil data for each habitat type and deepness. Key: Moledo 0-5 cm layer (MS), Moledo 5-10 cm layer (MD), Nariga 0-5 cm layer (NS), Nariga 5-10 cm layer (ND), Pragueira 0-5 cm layer (PS), Pragueira 5-10 cm layer (PD), Sálvora 0-5 cm layer (SS) and Sálvora 5-10 cm layer (SD). When expressed as amount of nutrients per surface unit (SM Table 3.5), compared to non-invaded areas, necromass under C. edulis accumulated significantly more B in back dunes, more Al and Na in rocky habitats and more Ca and Mn in both habitats, these differences being completely explained by the higher necromass amount. No significant differences were found for the other studied elements. 3. Effects of Carpobrotus edulis invasion on litter and soil 85 The best PCA with necromass properties (KMO= 0.614; Bartlett's test of sphericity P < 0.01) included the concentrations of Al, Co, Cu, Fe, Mg, Ni and Zn (SM Table 3.3b). This PCA extracted two factors that jointly explained 83% of the variance (66.3% and 16.9%, respectively, by Factor 1 and 2). Factor 1 was determined by all trace elements considered, with factor loadings higher than +0.70, whereas Factor 2 was determined by Mg and to a lesser extent by Co and Cu (factor loadings of +0.99, -0.42 and -0.42, respectively). Cristina Vieites Blanco 86 The plane of these two factors discriminate the samples according to both substrate and plant cover (Fig. 3.4): a) except native necromass in Nariga site, necromass over rocky substrates have positive loadings on Factor 2 while those over dunes have ever negative loadings; and b) in all cases, necromass from invaded sites had more negative loadings on Factor 1 and (except for Moledo site) more positive loadings on Factor 2 than native necromass. Fig. 3.3. Mean value ± standard error for some of the most representative studied necromass properties under autochthonous (A) and invasive (I) vegetation. Different letters (a, b) indicate significant differences (paired t-test; P <0.05) within type of vegetation. For Al, the significance (P= 0.052) was close to the considered threshold. 3.3.3. N pools In soils from rocky areas, N pools varied with the invasion. The initial NH4+-N pool was usually higher in the C. edulis invaded soils than in those under native vegetation, differences being wider for the 0-5 cm layer (Figs. 3.5-3.8). While the amount of NH4+-N decreased during the incubation in the uninvaded soils (irrespectively of soil depth), it showed contrasting tendencies in the invaded soils: increase in the surface layer and decrease in the sub-surface one (Figs. 3.5-3.8). Conversely, the initial NO3--N pool was higher in soils under native vegetation than in the C. edulis invaded soils and its size always increased during the incubation (Figs. 3.5-3.8). 3. Effects of Carpobrotus edulis invasion on litter and soil 87 Fig. 3.4. Score plots from the principal component analyses (PCA) performed with necromass data for each habitat type and deepness. Key: Moledo (M), Nariga (N), Pragueira (P) and Sálvora (S) sites. 3.3.4. Gross N fluxes In general, there was a good fit between the measured and the Ntrace modelled data of the size and 15N abundance of the NH4+-N and NO3--N pools (see Figs. 3.5-3.8). Irrespectively of vegetation cover and soil depth, no satisfactory adjustments were obtained for models with two organic N pools (SONlab and SONrec: 1% and 99% of total SON, respectively) and two mineralization rates (MSONlab and MSONrec, with first and zero order kinetic, respectively). In all Cristina Vieites Blanco 88 cases, the best fittings were obtained when only the mineralization of the labile fraction (MSONlab) was considered, because the values yielded by Ntrace for MSONrec were not normally distributed and its inclusion never improved model fitting; consequently, this rate was not included in the finally selected models. Except in the Nariga soil under native vegetation, the MSONlab was significantly higher in the topsoil than in the 5-10 cm layer (Fig. 3.9a). The effect of C. edulis invasion in the MSONlab rate was site- and depth-dependent, not showing a clear trend. The gross immobilization of NH4+ to recalcitrant N (INH4rec) was only modelled for surface soils under native vegetation, accounting for more than half of the total NH4+ immobilization in these soils (Fig. 3.9b). As for MSONlab, except in the Nariga soil under native vegetation, the gross immobilization of NH4+ to labile N (INH4lab) was significantly higher in the topsoil than in the 5-10 cm layer (Fig. 3.9a,b). Carpobrotus edulis invasion increased the INH4lab rate in surface soils (2.81 to 3.07x; 1.19x when considering INH4rec+INH4lab) and it had contrasting effects within sites in the 5-10 cm layer (0.79x for Nariga, 4.48x for Sálvora). Regarding NO3--N production processes, gross oxidation of recalcitrant N (ONrec, i.e. heterotrophic nitrification) was only modelled in the uninvaded surface soil of Nariga site (Fig. 3.9c), the inclusion of this rate in the model being necessary because it improves by 23% the misfit function. Both the ammonium oxidation (ONH4, i.e. gross autotrophic nitrification) and the total gross nitrification rate (ONH4+ONrec) decreased strongly (0.26 to 0.20x) in the 0-5 cm soil layer of the invaded areas, and moderately (0.61 to 0.79x) in the 5- 10 cm layer. These rates decreased with depth in native soils, while the reverse was true in invaded soils (Fig. 3.9c). In most soils, the immobilization of NO3- to recalcitrant N (INO3) was not well modelled by Ntrace and it had to be discarded because its probability density functions (PDFs) did not follow a normal distribution and its sampling accuracy (monitored with the Gelman’s R test) was not acceptable; moreover, the inclusion of INO3 in the model did not improve the misfit function. The INO3 rate was only modelled in Sálvora soils under C. edulis, even being the dominant gross NO3- consumption process in its surface layer (Fig. 3.9d). Conversely, the dissimilatory reduction of NO3- to NH4+ (DNRA) was the exclusive gross NO3- consumption process in the other soils (Fig. 3.9d), strongly decreasing in invaded topsoils (0.07 to 0.04x), and to a lesser extend in invaded deep soils (0.18 to 0.70x). 3. Effects of Carpobrotus edulis invasion on litter and soil 89 Fig. 3.5. Amount and 15N abundance of NH4+-N and NO3--N (mean ± s.e.) in Nariga soils (0-5 cm), under autochthonous vegetation (NAS) and Carpobrotus edulis (NCS), during the aerobic incubation. Lines show the fit of the model to the experimental data. Light grey points or lines refer to the 15NH4NO3 experiment, whilst black points or lines refer to the NH415NO3 experiment. Cristina Vieites Blanco 90 Fig. 3.6. Amount and 15N abundance of NH4+-N and NO3--N (mean ± s.e.) in Nariga soils (5-10 cm), under autochthonous vegetation (NAP) and Carpobro-tus edulis (NCP), during the aerobic incubation. Lines show the fit of the model to the experimental data. Light grey points or lines refer to the 15NH4NO3 experiment, whilst black points or lines refer to the NH415NO3 experiment. 3. Effects of Carpobrotus edulis invasion on litter and soil 91 Fig. 3.7. Amount and 15N abundance of NH4+-N and NO3--N (mean ± s.e.) in Sálvora soils (0-5 cm), under autochthonous vegetation (SAS) and Carpobrotus edulis (SCS), during the aerobic incubation. Lines show the fit of the model to the experimental data. Light grey points or lines refer to the 15NH4NO3 experiment, whilst black points or lines refer to the NH415NO3 experiment. Cristina Vieites Blanco 92 Fig. 3.8. Amount and 15N abundance of NH4+-N and NO3--N (mean ± s.e.) in Sálvora soils (5-10 cm), under autochthonous vegetation (SAP) and Carpobro-tus edulis (SCP), during the aerobic incubation. Lines show the fit of the model to the experimental data. Light grey points or lines refer to the 15NH4NO3 experiment, whilst black points or lines refer to the NH415NO3 experiment. 3. Effects of Carpobrotus edulis invasion on litter and soil 93 Fig. 3.9. Ntrace modelled gross fluxes in the 0-5 cm and 5-10 cm soil layers under native vegetation and Carpobrotus edulis in Nariga and Sálvora sites: (a) mineralization of labile SON to NH4+ (MSONlab); (b) immobilization of NH4+ to recalcitrant (INH4rec) and to labile SON (INH4lab); (c) autotrophic (ONH4) and heterotrophic (OSON) nitrification; and (d) dissimilatory NO3- reduction to NH4+ (DNRA) and NO3- immobilization to SON (INO3). Significant differences - tested by an overlap of the 85% confidence intervals (Payton et al., 2000; Rutting et al., 2010) - is shown by different capital letters for the vegetation types (native or C. edulis), and lowercase letters for soil layers (0-5 or 5-10 cm). Cristina Vieites Blanco 100 3. Effects of Carpobrotus edulis invasion on litter and soil 101 Cristina Vieites Blanco 102 3. Effects of Carpobrotus edulis invasion on litter and soil 103 Cristina Vieites Blanco 104 3. 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(2017) Invasive species’ leaf traits and dissimilarity from natives shape their impact on nitrogen cycling: a meta-analysis. New Phytologist, 213, 128- 139. Li, H., Wei, Z., Huangfu, C., Chen, X. & Yang, D. (2017) Litter mixture dominated by leaf litter of the invasive species, Flaveria bidentis, accelerates decomposition and favors nitrogen release. Journal of Plant Research, 130, 167- 180. Macías Vázquez, F. & Calvo de Anta, R. (2009) Niveles genéricos de referencia de metales pesados y otros elementos traza en suelos de Galicia. Xunta de Galicia, Santiago de Compostela. Mack, M. C. & D'Antonio, C. M. (2003) Exotic grasses alter controls over soil nitrogen dynamics in a Hawaiian woodland. Ecological Applications, 13, 154- 166. Maltez-Mouro, S., Maestre, F. T. & Freitas, H. (2010) Weak effects of the exotic invasive Carpobrotus edulis on the structure and composition of Portuguese sand-dune communities. Biological Invasions, 12, 2117-2130. Molinari, N., D'Antonio, C. & Thomson, G. (2007) 7 - Carpobrotus as a case study of the complexities of species impacts. Theoretical Ecology Series (eds K. Cuddington, J. E. Byers, W. G. Wilson & A. Hastings), pp. 139-162. Academic Press. Msanne, J., Awada, T., Bryan, N. M., Schacht, W., Drijber, R., Li, Y., Zhou, X., Okalebo, J., Wedin, D., Brandle, J. & Hiller, J. (2017) Ecophysiological responses of native invasive woody Juniperus virginiana L. to resource availability and stand characteristics in the semiarid grasslands of the Nebraska Sandhills. Photosynthetica, 55, 219-230. Müller, C., Rütting, T., Kattge, J., Laughlin, R. J. & Stevens, R. J. (2007) Estimation of parameters in complex 15N tracing models by Monte Carlo sampling. Soil Biology & Biochemistry, 39, 715-726. Müller, C., Stevens, R. J. & Laughlin, R. J. (2004) A 15N tracing model to analyse N transformations in old grassland soil. Soil Biology & Biochemistry, 36, 619- 632. Cristina Vieites Blanco 116 Among the native predators of this plant, the South African scale insect Pulvinariella mesembryanthemi (Vallot) is a highly specific parasite of Carpobrotus spp. (Miller and Miller, 2003, Miller et al., 2005). Its potential value as a biological control agent has been reported by several authors (Washburn and Frankie, 1985, Fagúndez and Beiras, 2007). However, to the best of our knowledge, no studies have evaluated this potential yet. This insect may have been introduced accidentally with infested plants outside its native range, as it is currently found in different coastal temperate regions of Europe, Northern and Southern America, Northern Africa and Oceania (Cebeci and Selmi, 2004, Granara de Willink and Claps, 2003, Gómez-Menor Ortega, 1954). When present in high numbers, P. mesembryanthemi can cause considerable damage to C. edulis, as observed in plantations in California (Donaldson et al., 1978, Washburn and Frankie, 1985). It is exclusively parthenogenic (the males that are occasionally produced do not reproduce) and can be univoltine or bivoltine, depending on the environmental conditions (Washburn and Frankie, 1985). Active dispersion of P. mesembryanthemi is limited to the first stages of their life cycle, and long-distance dispersion is mediated by wind (Washburn and Frankie, 1981). Synergistic effects between biocontrol agents are scarcely observed (Xu et al., 2011). However, studying the combined effects of different potential biocontrol agents is important. Different organisms can interact directly or indirectly, with different net effects on the target plants (Morris et al., 2007, Campanella et al., 2009). In the case of C. edulis, the combination of P. mesembryanthemi and S. sclerotiorum appears promising, as each species targets different parts of the plant: thus, the insect prefers to attack younger leaves (Washburn and Frankie, 1985), while the fungus appears to infect more successfully the older leaves (pers. obs.). Invasive plants can undergo rapid adaptive evolution in the new environments, thus altering the interactions they have with antagonists (Müller-Schärer et al., 2004, Prentis et al., 2008). The susceptibility of the plant to the biocontrol agents may depend on the plant genotype, which may vary between the invaded area and the native area (Paterson et al., 2009, Hinz and Schwarzlaender, 2004). In C. edulis, there are some evidences of differences between native plants from South Africa and introduced plants in the Iberian Peninsula regarding division of labour capacity (Roiloa et al., 2016). These differences may be due to an 4. Biocontrol potential of S. sclerotiorum and P. mesembryanthemi 117 evolutionary adaptation of the plant to the invaded area (Roiloa et al., 2016). The aim of this study was to test the individual and combined effects of two potential biocontrol agents - the insect P. mesembryanthemi and the fungus S. sclerotiorum - on C. edulis plants from the native area of distribution and from invaded areas. We hypothesise that the biocontrol capacity of both biological agents together will be greater than that of each individual as a result of synergistic effects. Likewise, we predicted that C. edulis plants from the invaded areas will be more sensitive to both agents as these plants are unlikely to have previously encountered the pathogens. To this end, we evaluated the effects on short-term (physiological indexes) and long-term (survival, growth, biomass allocation) estimators of plant performance commonly used as indicators of fitness (Traveset et al., 2008). To our knowledge, this is the first study investigating the biological control of C. edulis. 4.2. MATERIAL AND METHODS 4.2.1. Plant material Samples of plants from four native populations of C. edulis growing in the Cape Region (South Africa) and another four invasive populations in NW Iberia (Table 4.1), all growing in coastal dune systems, were used in the study. In August 2015, 24 uniformly-sized ramets with 2-3 nodes were removed from the stock plants in each population and transplanted into individual 5 L plastic pots containing a 1:1 mixture of sand and peat as substrate. The initial total fresh weight of the ramets used (measured on November 2015) was 16.13±0.42 g for native plants and 27.69±1.36 g for introduced plants. Plants were maintained in a greenhouse in Santiago de Compostela (Spain; 45.874° N; 8.559° W) under controlled conditions (22 ± 2 °C, natural photoperiod) and were watered regularly throughout the experiment to field capacity. Within each population, plants were randomly allocated to one of four treatments, which were equally represented (six plants per treatment): ‘C’, control (untreated) plants; ‘S’, plants inoculated with the fungus S. sclerotiorum; ‘P’, plants artificially infested with the insect P. mesembryanthemi; ‘PS’, plants treated with both S. sclerotiorum and P. mesembryanthemi. After the application of the treatments, the plants were randomly relocated 3 times during the experiment. Cristina Vieites Blanco 118 Table 4.1. Populations of the plant C. edulis used in this study. Information about the origin of the plants (native or introduced) and coordinates (in decimal degrees) is provided. 4.2.2. Fungal isolate and inoculation The S. sclerotiorum isolate used in the study was extracted from Brassica napus L. growing in Salcedo, Pontevedra, Spain (42.255° N, 8.384° W, 20 m asl) and provided by the Misión Biológica de Galicia (CSIC). Sclerotia were first grown on potato dextrose agar (PDA, Fluka) in Petri dishes (diameter 9 cm), and sub-cultures were established (also on PDA) from 1.5 cm diameter discs of fresh mycelial growth. In October 2016, plants were inoculated with 1.5 mm diameter discs of fresh mycelium from the sub-cultures. Each disc was placed in a recently cut leaf tip of the second node of the main ramet and was covered with a pipette tip to maintain humidity. After one month, a second inoculation was made in the stem (in the first node of the ramet) to ensure the introduction of the fungus in the plants. To this end, a small cut was made in the stem, and fresh mycelium was inserted. The incision was covered with Parafilm® to maintain the humidity. Inoculated leaves that did not exhibit symptoms of infection were reinoculated with fresh mycelium. At the end of the experiment, one year after the first inoculation, small slices of leaves were removed from the surviving inoculated plants to test for the presence of S. sclerotiorum. To isolate the fungus, the tissue was dipped in 2% bleach for approx. 5 seconds and rinsed with deionized water, before being transferred to potato dextrose agar (PDA) medium and incubated at room temperature (20-22 °C). 4. Biocontrol potential of S. sclerotiorum and P. mesembryanthemi 119 4.2.3. Insect source and infestation Individual specimens of P. mesembryanthemi were collected from Punta Nariga (Spain, 43.230° N; 8.910° W) and bred in the greenhouse under controlled conditions (22 ± 2 °C, natural photoperiod). The mean number (±SE) of eggs produced by each female was 333 (± 12). Two mature ovisacs were used to infest each plant in October 2016, by placing them on the first node and on a secondary node of the ramet, respectively. After one month, plants were checked for live crawlers and settlers, and plants in which none of the ovisacs had hatched were infested with another two ovisacs. Only one of the plants (which died after 2.5 months) was not infested with P. mesembryanthemi after the second infestation. 4.2.4. Insect cycle The number of insects on each life stage (instar 1 to 4 and ovisacs; as described by Washburn and Frankie (1985)) per plant was counted every 3-4 weeks. The life cycle of the insect for the treatments ‘P’ (Pulvinariella) and ‘P+S’ (Pulvinariella + Sclerotinia) and the different life stages were represented. Survival was also calculated for each date on which the number of insects was counted. 4.2.5. Physiological measures and visual rating All ecophysiological measures were determined on fully developed and healthy leaves of the main ramet of each plant, before (- 1 day) and after (1week; 3 weeks; 1 month; 1.5 months and then monthly until 1 year) the first inoculation and infestation. Leaf spectral reflectance (300-1100 nm) was measured with a portable spectrometer (UniSpec Spectral Analysis System; PP Systems, Haverhill, MA, USA). The following indexes were calculated from the reflectance data: (1) the photochemical reflectance index ‘PRI’ (R531- R570)/(R531+R570), where R is reflectance and the subscript refers to the wavelength in nanometers, which is inversely correlated with the zeaxanthin content (a photoprotective pigment) and directly correlated with photosynthetic radiation-use efficiency and net CO2 uptake (Gamon et al., 1997), Peñuelas et al., 1995a); (2) the chlorophyll index ‘CHL-NDI’ (R750-R705)/(R750+R705), which is directly proportional to the chlorophyll content of leaves (Richardson et al., 2002, Gitelson et al., 1996); Cristina Vieites Blanco 120 (3) the structural independent pigment index ‘SIPI’ (R800-R445)/(R800- R680), related to the carotenoid/chlorophyll ratio (Peñuelas and Inoue, 1999); (4) the water index ‘WI’ (R970/R900), used to estimate plant water concentration (Peñuelas et al., 1997); and (5) the normalised phaeophytinization index NPQI (R415-R435)/(R415+R435), indicative of chlorophyll degradation to phaeophytin (Barnes et al., 1992). Chlorophyll fluorescence parameters were measured by the saturation pulse method (Screiber et al. 1998) with a portable pulseamplitude-modulated fluorometer (MINI-PAM photosynthesis yield analyser; Walz, Effeltrich, Germany). The leaves were allowed to adapt to darkness for 30 minutes, and the maximum quantum yield of photosystem II (Fv /Fm) was then estimated as the ratio (Fm-F0)/Fm, where Fm and F0 are respectively the maximal and minimal fluorescence yield. Fv /Fm is indicative of the photosynthetic performance of plants (Schreiber et al., 1998) and is correlated with the amount of carbon gained per unit of light absorbed (Bolhàr-Nordenkampf and Öquist, 1993), as well as with photoinhibition of photosynthesis and zeaxanthin production (Baker, 2008, Falbel et al., 1994, Leverenz et al., 1992). In addition, on the same dates that the reflectance and fluorescence measurements were made, the plant performance was assessed using a visual rating based on a 1-9 scale, where 1 indicates a perfectly healthy plant and 9 a dead plant. 4.2.6. Growth rate and dry mass allocation At the end of the experiment, plants were oven-dried at 60 ºC to constant weight to enable the determination of the dry weight of the aerial part (leaves + shoots) and for the roots. The root/aerial biomass ratio was then calculated. A randomly-selected sample of fresh plants was weighted, immediately after being removed from their pots and then after drying to enable calculation of the water content of the plants. The shoot length was also measured, and nodes and leaves were counted both at the beginning and at the end of the experiment, as estimators of plant growth. The increase in the number of nodes (points of leaf insertion in the shoots) has been found to be closely correlated with growth of C. edulis (Traveset et al., 2008). The internodal distance was estimated as the total shoot length of the plant divided by the number of nodes. 4. Biocontrol potential of S. sclerotiorum and P. mesembryanthemi 121 4.2.7. Statistical analysis All analyses were performed with R statistical software (https://www.r-project.org/). Differences in leaf spectral reflectance and fluorescence indexes were analysed by linear mixed models via restricted maximum likelihood (REML) and post hoc tests, using the ‘lme4’, ‘lsmeans’ and ‘multcomp’ packages. The inclusion of random factors in the model (population, origin or subject) was conditioned by the AIC values (Akaike, 1974). The random effect of population nested within origin was therefore never included, while the subject always improved the model. When the analysis of variance (ANOVA) test indicated a significant difference, the following were included as fixed factors in the model (following the minimal adequate model): treatment (‘C’, control plants; ‘S’, S. sclerotiorum inoculated plants; ‘P’, P. mesembryanthemi infested plants; ‘PS’, plants treated with both S. sclerotiorum and P. mesembryanthemi), date, origin (native or non-native populations) and their interactions. The assumptions of linearity, homoscedasticity and normality were checked by plotting the model residuals. Only measurements made after the application of the treatments (from 1 week to 12 months) were used in the analysis. A negative binomial mixed model was applied for visual rating and for the number of insects on each plant (total and per life stage: instars 1st- 4th and ovisacs), with subject as a random factor and treatment or origin as fixed factors. The significance of the factors was checked with the Wald test. For those factors indicated to be significant, post-hoc tests were conducted. The analyses were done using the ‘R2admb’, ‘glmmADMB’, ‘survey’ and ‘multcomp’ packages. A general lineal model was used to analyse insect survival and variables related to plant growth and biomass (growth in shoot length, increase in number of nodes, increase in number of leaves, final dry aerial biomass, final dry root biomass, root/aerial biomass rate, water content), with treatment, origin and their interaction as fixed factors. The assumptions of the model were checked, and when not met, the Tukey’s ladder of powers was applied to the variables (i.e. the square root was applied to root biomass; the decimal logarithm to the root: aerial biomass rate; the square and cube were applied to some of the dates in the survival analysis). Cristina Vieites Blanco 122 The non-parametric Kaplan-Meier procedure (Kaplan and Meier, 1958) was used to construct the plant survival curves. To study the relative effect of the treatments, plant origin and their interaction on plant survival, a Cox proportional hazards regression model (Cox, 1972) was used after checking the assumption of proportional hazard (Grambsch and Therneau, 1994). The log rank test (Harrington and Fleming, 1982) was used to check the homogeneity of survival functions between treatments. For pairwise comparisons, the log-rank test was applied, with Bonferroni correction (Bonferroni, 1936). Survival analyses and representations were conducted using the ‘survival’ and ‘survminer’ packages. Correlations between variables were tested using the Pearson’s product moment correlation coefficient. 4.3. RESULTS 4.3.1. Visual observations and recovery of fungus At the end of the experiment (1 year), the surviving plants of the initial 48 per treatment were: 37 fungus-infected plants (‘S’), 17 insectinfested plants (‘P’) and 18 plants treated with both agents (‘PS’). Within the surviving plants, visual examination revealed death of some plant parts in 3 fungus-infected plants (‘S’), in 9 insect-infested plants (‘P’) and in 4 plants treated with both agents (‘PS’). At the end of the experiment, most of the insect-infested plants (i.e. ‘P’ and ‘PS’ treatments) were found to be affected by black mould (82 and 83 % of the plants, respectively). The inoculated fungus was only recovered from a few plants one year after the first inoculation. 4.3.2. Insect life cycle and survival The insect underwent two complete life cycles and one partial cycle during the experiment, and the duration of the life cycle under the particular greenhouse conditions was 4-6 months (Fig. 4.1). The origin of the plant (i.e. from native or introduced populations) and the treatment (‘P’ or ‘P+S’) did not seem to affect the duration of the life cycle (Fig. 4.1). Some male specimens of the insect were occasionally found on a few plants. 4. Biocontrol potential of S. sclerotiorum and P. mesembryanthemi 123 Fig. 4.1. Mean number ± SE of P. mesembryanthemi insects per plant throughout the year of the experiment (October 2016 - October 2017). The different graphs represent the total number of insects, the insects at the different growth stages (1st, 2nd, 3rd, 4th instars and ovisacs). Black and grey lines represent respectively plants from invasive and native populations. Solid lines represent treatment ‘P’ (plants infested with the insect) and dashed lines represent treatment ‘PS’ (plants infested with the insect and inoculated with the fungus). Cristina Vieites Blanco 124 The total numbers of P. mesembryanthemi, as well as the number of insects on 1st, 2nd and 3rd instar and ovisacs per plant, were significantly higher in insect-only infested plants (‘P’) than in plants in which both agents were combined (‘PS’) (Table 4.2). Significantly greater total numbers of insects per plant, as well as of 2nd and 4th instar and of ovisacs, were found on the introduced plants than on native plants (Table 4.2). Survival of the insect on the plants did not vary depending on the treatment (‘P’, ‘PS’) or the origin of the plant (‘N’, ‘I’), and the interaction Treatment*Origin was only significant after 10 months (F1,42=4.80, p =0.034; non-significant post-hoc interactions, p >0.05). Table 4.2. Results of the Wald test for the mixed negative binomial models of the number of P. mesembryanthemi insects (total and for each of the different stages: 1st to 4th instars and ovisacs) counted on the C. edulis plants in the experiment, with treatment and plant origin as fixed factors. The table includes the numerator (nDF) and denominator (dDF) degrees of freedom, the F-value and p-value of the tests, and the mean ± SE for the total number of insects counted throughout the experiment per plant, for each treatment (P, plants infested with the insect; PS, plants both infested with the insect and inoculated with the fungus) and plant origin (native, introduced). Significant differences (at p ≤0.05) across treatments or origins are indicated by different letters. The mean number of insects per plant from the first life cycle of P. mesembryanthemi was positively correlated with the insect settlement (number of 1st instars) in the subsequent cycles (Pearson’s correlation=0.255, t77=2.31, p =0.023). 4.3.3. Leaf spectral reflectance, chlorophyll fluorescence and visual rating The ‘treatment’ factor was significant for all the indexes included in the analysis, except for WI (Table 4.3). Both treatments with P. 4. Biocontrol potential of S. sclerotiorum and P. mesembryanthemi 125 mesembryanthemi infestation (‘P’ and ‘PS’) yielded lower values of PRI and higher values of SIPI than the control (untreated) plants (‘C’). For the other indexes (CHL-NDI, Fv/Fm, NPQI), only the treatment in which both agents were applied together (‘PS’) produced significantly lower values than the control (no treatment) (Table 4.4). There were no significant differences between ‘P’ and ‘PS’ or between ‘S’ and ‘C’ (Table 4.4). Table 4.3. Mixed-model results from C. edulis experiment describing the variation in the reflectance indexes for the fixed factors and their interactions. The results of a type III ANOVA with Satterthwaite approximation for the numerator (nDF) and denominator (dDF) degrees of freedom are shown. ‘Treat’, treatment applied to the plant; ‘Or’, origin of the plant. For all spectral reflectance and fluorescence indexes, the time since the start of the experiment (‘Date’) was a significant factor (Table 4.3), although no clear trend was observed (Figs. 4.2, 4.3). The variations in the indexes throughout the year of the experiment were similar for all treatments (Figs. 4.2, 4.4). However, for all variables the effect of the treatment was time-dependent (significant ‘Treat*Date’ interaction; Table 4.3). No differences between treatments were found after 1 week for any of the indexes studied (Fig. 4.2, 4.3, SM4.1). The values were lower in plants inoculated with S. sclerotiorum than in the control (untreated) plants Cristina Vieites Blanco 132 However, there was no significant interaction between the plant origin and the treatment (Table 4.9). Final dry weights of both the aerial parts and the roots were significantly lower for the insect-infested plants (‘P’ and ‘PS’) than for the control plants (‘C’), and lower for the native than for the introduced plants (Table 4.7). The root / aerial biomass ratio was also significantly lower in both treatments with the insect than in the ‘Control’ and ‘Sclerotinia’ treatments, although they also differed from each other (the ratio was lower for ‘P’ than ‘PS’: p =0.052) (Table 4.7). The origin of the plant did not significantly affect the root / aerial biomass ratio (Table 4.8). The water content of the plants at the end of the experiment was only affected by the treatment, and it was significantly lower in the fungus-inoculated + insect-infested plants ‘PS’ than in the control plants. The internodal distance was higher in introduced than in native plants (Table 4.7), but did not vary depending on the treatment applied (Table 4.9). The maximum number of insects counted in a plant throughout the trial was positively and significantly correlated with the final dry weight of plant aerial and root biomass, shoot growth, increase in number of nodes and leaves, increase in internodal distance and initial fresh weight of the biomass (Table 4.10). Table 4.10. Pearson’s correlation coefficients between maximum number of insects per plant and the biomass and growth parameters of C. edulis plants (aerial and root dry biomass, root/aerial dry biomass ratio, shoot growth, increase in the number of nodes and leaves, internodal distance and initial fresh biomass). 4. Biocontrol potential of S. sclerotiorum and P. mesembryanthemi 133 4.4. DISCUSSION 4.4.1. Plants treated with the fungus The fungus S. sclerotiorum by itself (treatment ‘S’) had only very brief, short-term effects on the ecophysiological parameters of C. edulis considered, with some decreases in chlorophyll content and photosynthetic-radiation use efficiency (as suggested by the CHL-NDI and Fv/Fm indexes, respectively) in the first half of the year of the experiment. This may indicate that the plants were able to recover quickly from the infection and were only significantly affected for very short periods of time. Fungal pathogens vary in their effects on plants and can produce necrotic or chlorotic diseases (Malthus and Madeira, 1993). In this case, the reflectance values suggest that S. sclerotiorum produces leaf chlorosis, as observed by Lorenzen and Jensen (1989) for mildew-infected leaves and in contrast to the findings reported by Malthus and Madeira (1993) for plants infected with Botrytis fabae. Sclerotinia sclerotiorum did not significantly affect plant survival one year after inoculation, either alone (treatments ‘C’ and ‘S’ did not have significantly different effects) or when combined with the insect (treatments ‘P’ and ‘PS’ did not have significantly different effects). This fungus usually kills plants within a month (Cother, 2000, Waipara et al., 1993, Bourdôt and Harvey, 1994), and therefore it was not expected to affect plant survival even in the longer term. This agent did not significantly affect any of the growth and biomass indicators considered. Overall, the fungus S. sclerotiorum only significantly affected some short-term estimators of fitness at specific times, but it did not alter any of the long-term estimators measured (survival, growth and dry mass allocation). However, the efficiency of S. sclerotiorum infection varies depending on climate conditions and is favoured by high humidity and moderate temperatures of around 20 °C (Berg and Lentz, 1968). Therefore, growth of the fungus may be higher in the field than in greenhouse or laboratory conditions, especially in the warm humid months of spring and autumn. 4.4.2. Plants treated with the insect Considering all dates, the insect P. mesembryanthemi significantly decreased the photosynthetic-radiation use efficiency (as suggested by the Cristina Vieites Blanco 134 PRI) of the plants. However, the effects of the insect on the plant varied over time. A positive and occasional effect of the insect on photosynthetic performance (Fv/Fm) was observed at the beginning of the experiment. This was similar to the observed increase in photosynthetic efficiency in Ilex aquifolium plants infested with scale insects (Retuerto et al., 2004), which was attributed to an imbalance between source and sink tissues in the host plants caused by the insect feeding, promoting compensatory photosynthesis in the plants. The first symptoms of negative effects of the insect P. mesembryanthemi were observed from 5 months after the infestation (with the exception of the lower Fv/Fm observed after 2.5 months), coinciding with the most advanced life stages during the first life cycle (4th instar or ovisac). Infestation with the insect affected all of the indexes considered at some point during the trial. The insect generally took longer to affect the plant than the fungus (2.5 months for Fv/Fm; 5 months for WI and SIPI; 6 months for PRI and 10-11 months for the other indexes), but the symptoms lasted longer. However, the Fv/Fm values were close to 0.8 throughout the year of the trial and for all four treatments. This indicates that the plants were affected by a low level of stress during the experiment (Traveset et al., 2008, Maxwell and Johnson, 2000), even for the treatments including insects (‘P’ and ‘PS’). Like the fungus, the insect seemed to induce chlorophyll degradation (inferred by the decrease in PRI, CHL-NDI and NPQI). Chlorosis has previously been observed in plants infested by other sap-sucking insects such as mites (Peñuelas et al., 1995b) . The reflectance indexes show that the insect P. mesembryanthemi was also able to decrease the water content of the leaves (WI) and the photosynthesis efficiency (PRI) at some points during the trial. Water depletion of the plants could be caused by intense insect feeding (Washburn et al., 1985). In contrast to the effect of the fungus, the insect P. mesembryanthemi affected C. edulis plants in both the short-and the long-term. Survival decreased greatly in the second half of the year-long trial. The increase in mortality in the treatments infested by P. mesembryanthemi (‘P’ and ‘P+S) from the 6th month coincided with the end of the first cycle of the insect, when the insects were fully grown and were probably feeding more intensively on the plant. Furthermore, growth of the plants infested with the insect was lower in terms of shoot length, number of nodes and leaves and aerial and root dry biomass. Although the infestation initially 4. Biocontrol potential of S. sclerotiorum and P. mesembryanthemi 135 enhanced plant photosynthesis (as indicated by the Fv/Fm), the decrease in photosynthetic performance found from 2.5 months after the infestation and the loss of sugars due to feeding seemed to have a detrimental effect on plant growth. Washburn et al. (1985) also observed lower plant growth in plants infested with P. mesembryanthemi, probably due to a reduction in the plant capacity to obtain resources and increased vulnerability to stress (Vranjic and Ash, 1997). In fact, extensive mortalities of P. mesembryanthemi - infested C. edulis plants in the field have been associated with stress conditions of drought and freezing temperatures (Donaldson et al., 1978, Washburn and Frankie, 1985). Although insect feeding sometimes lowered the plant water index, the water content of the plants at the end of the experiment was not significantly different from that of the control (untreated) plants. As plants were watered regularly throughout the trial, they were able to replenish the water lost due to insect feeding. However, in nature, P. mesembryanthemi infestation could increase the vulnerability of C. edulis to drought conditions. The root / aerial biomass ratio was lower in the plants infested with P. mesembryanthemi than in the control plants. The higher proportion of photosynthetic tissue suggests (as indicated by the Fv/Fm index) a measure that compensates for the carbohydrates lost due to insect feeding. The decrease in plant growth produced by the infestation could also be detrimental to the next generations of insects as they would have less material on which to feed (Washburn et al., 1985). Nevertheless, we observed a positive correlation between the mean density of insects in the first life cycle and the number of new settled insects (1st instar) in the subsequent cycles. In contrast to the findings of Washburn et al. (1985) and despite the insect being a deterrent to plant growth, we found that the C. edulis plants supporting the highest density of insects were those with the largest aerial and root dry biomass, shoot growth and increase in number of leaves. The findings may reflect that the density of insects on the plants is limited by the size of the plant because of competition for feeding space. Therefore, bigger plants would support larger numbers of insects. The significant and positive correlation between the initial fresh biomass weight and maximum number of insects per plant corroborates this supposition. The maximum insect density on plants was also positively correlated with the internodal distance, suggesting an exploratory Cristina Vieites Blanco 136 strategy and faster spread of the plants most affected by insects (Traveset et al., 2008, Doust, 1981). In around half of the plants that were only treated with the insect and that survived until the end of the trial, death of the plant parts supporting high densities of P. mesembryanthemi was observed. This could be a defence mechanism of the plant, as it increases the insect mortality and decreases the number of offspring (Washburn et al., 1985). Scale insects are particularly strongly affected by the death of plant tissues as their limited mobility restricts their capacity to spread to healthy plant parts (Washburn et al., 1985). Most of the insect-infected plants were affected by black sooty mould, which has previously been associated with the accumulation of honeydew excreted by scale insects (Bokonon-Ganta et al., 2002). The presence of the mould may lead to a decrease in the level of photosynthesis, as observed in mango trees infested with mealybugs (Bokonon-Ganta et al., 2002). On the other hand, this fungus has also been attributed to increased mortality of scale insects (Collins and Scott, 1982). As ants reduce the accumulation of honeydew (Bach, 1991), the high level of colonization by mould observed in the greenhouse-grown plants might not occur in the field, due to the natural presence of ants. The male specimens of P. mesembryanthemi occasionally observed throughout the trial were probably derived from the same female on each plant, as they occurred close together on the plants (Pesson, 1941). 4.4.3. Plants treated with the combination of the fungus and the insect Only when the insect and fungus were applied together, plant chlorophyll content and photosynthetic performance were significantly lower than in control plants, suggesting that the combination of both agents may be more effective than the application of either agent by itself. This is consistent with the findings of other studies in which the combined use of an insect and a fungus proved more effective than the use of each biocontrol agent alone (e.g. Caesar (2003) in Euphorbia sp.). Nevertheless, the effects of the combination of both agents on the ecophysiological parameters were temporary, and we did not observe any long-term synergetic effects, as treatments ‘P’ and ‘PS’ did not produce any significant differences in survival, biomass or growth of the plants. 4. Biocontrol potential of S. sclerotiorum and P. mesembryanthemi 137 The higher densities of insects observed in treatment ‘P’ than in ‘PS’ indicate a possible detrimental effect of fungal infection on the insect. Insect survival was not affected by the treatment, but the number of 1st instar stages was already higher in plants in treatment ‘P’. This may indicate greater success of the insect colonization in ‘P’ plants than in ‘PS’ plants. The emergence of the insects from the first life cycle (after approx. 4 weeks) coincides with the highest effect of the fungus on the chlorophyll content of leaves (as suggested by the CHL-NDI after approx. 3 weeks). The lower chlorophyll content of the ‘PS’ plants due to the effect of the fungus may negatively affect settlement of the insects, possibly due to lower quality sap, with consequently lower densities of insects. Carpobrotus edulis can produce antifungal compounds (Omoruyi et al., 2014), and inoculation of the plants with S. sclerotiorum may have induced production of such substances. This, in turn, may have affected P. mesembryanthemi, explaining the lower densities in treatment ‘PS’. For instance, saponins, which are present in C. edulis (Omoruyi et al., 2014), are both antifungal and insecticidal compounds (De Geyter et al., 2007). In a study combining an insect and a fungus for plant biocontrol, Campanella et al. (2009) also observed indirect competition between both agents, which was detrimental to the insect. The higher densities of insects found in ‘P’ plants may explain the higher biomass allocation to the aerial parts (and therefore to photosynthetic tissues) than in plants treated with both agents (‘PS’), as they would have to cope with greater loss of sugars through insect feeding. In this case, the combination of both agents did not appear to be beneficial for controlling C. edulis, as the fungus did not have long-term repercussions on the plants and it had a detrimental effect on the settlement of the insects on the plants. 4.4.4. Plant origin (native or non-native populations) Although plant origin affected some of the reflectance indexes, no ‘Treatment*Origin’ interaction was observed. The reflectance indexes therefore did not indicate any differences between native and introduce plants in the vulnerability to the biocontrol agents. The native plants showed lower growth (as estimated by the lower increase in shoot length and node and leaf number) and lower aerial and root dry biomass. However, the native plants were not more susceptible Cristina Vieites Blanco 138 to the fungus or insect attack than the introduced plants, as the ‘Treatment*Origin’ interaction for all these variables was not significant. Native plants also had lower internodal distances than the introduced plants, which suggests a slightly different, more explorative strategy, in non-native plants. The introduced plants were infested by greater total numbers of insects per plant. This did not appear to be the result of any differences in settlement success, as there was no difference in the number of 1st instar stages. The greater success of the insect on the invasive plants may be explained by the larger biomass of introduced plants than of natives plants, which would imply more space for feeding. Also, plant defence is usually lower in exotic populations than in native populations, which can benefit the performance of specialist insects (Hinz and Schwarzlaender, 2004). 4.5. SYNTHESIS In a preliminary attempt to find a biological control of Carpobrotus edulis, we evaluated the potential use of the fungus Sclerotinia sclerotiorum and the insect Pulvinariella mesembryanthemi as biocontrol agents. We carried out a greenhouse-experiment to evaluate the effects of both agents, separately and together, on short-term (physiological) and long-term (survival, growth, biomass) estimators of plant performance. We compared the susceptibility to both agents in plants originating from native and non-native areas. The fungus had immediate and negative, but short-lasting, effects on chlorophyll content and photosynthetic-radiation use efficiency. No significant effects on plant survival, growth and biomass were observed after one year. Artificial infestation with the insect increased photosynthetic performance and decreased the root/aerial biomass ratio of the plants, suggesting a counteractive response to insect feeding. After 5 months, the reflectance parameters were found to be negatively affected. Only half of the infested plants survived for a year, and the growth and biomass were lower in the surviving plants than in untreated (control) plants. In half of the surviving plants, the most heavily-infested parts died. The insectinfested plants were usually also infected by black mould. 4. Biocontrol potential of S. sclerotiorum and P. mesembryanthemi 139 The density of insects was lower in native plants and when both biocontrol agents were used together. Nevertheless, no long-term synergetic effects of the insect and fungus were observed, and the susceptibility of native and introduced plants was not different. Therefore, use of the insect seems to be the best strategy for controlling C. edulis, as it decreases plant growth and increases plant mortality and susceptibility to stress. 4.6. SUPPLEMENTARY MATERIAL SM 4.1. (next pages) Mean ± SE for the different calculated reflectance indexes from C. edulis experiment in control (C), P. mesembryanthemi ‐ infested (P), S. sclerotiorum – inoculated and both infested and inoculated (PS) plants for the 14 times reflectance was measured throughout the experiment. Different letters in a row represent significant differences between treatments for each time of measurement (weeks or months since the beginning of the experiment) according to the post‐hoc test. Significance was set at p ≤0.05. 140 Cristina Vieites Blanco 141 4. Biocontrol potential of S. sclerotiorum and P. mesembryanthemi Cristina Vieites Blanco 244 dunares, lo que podría explicar el diferente efecto de C. edulis en el pH: con incrementos en dunas (posiblemente por una mayor absorción de Ca de zonas profundas del suelo que las especies nativas) y descensos en zonas rocosas (probablemente por una mayor producción de ácidos orgánicos procedentes de la descomposición de necromasa). No obstante, en otros ambientes dunares C. edulis acidifica mayoritariamente los suelos, posiblemente por producción de ácidos orgánicos (D'Antonio, 1990, Santoro et al., 2011, Conser y Connor, 2009). Esta diferencia en el efecto de C. edulis en el pH de suelos dunares y rocosos tiene como consecuencia efectos variables sobre la disponibilidad de nutrientes: en zonas rocosas invadidas aumenta la disponibilidad de Fe y disminuye la de Cu, Mg y Zn; mientras que en dunas invadidas disminuye la disponibilidad de Fe y Co. La reducción en la disponibilidad de estos nutrientes podría conllevar deficiencias en plantas y microorganismos (Williams y Fraústo da Silva, 2000). En dunas invadidas por C. edulis, al contrario que en otros estudios (Novoa et al., 2014, Santoro et al., 2011) así como en la sucesión natural de dunas (Anwar-Maun, 2009, Jones et al., 2008) y a pesar del incremento en necromasa acumulada, el C orgánico y el N total disminuyen en el suelo. La producción por parte de C. edulis de compuestos antibacterianos (van der Watt y Pretorius, 2001) podría estar inhibiendo la descomposición microbiana de su necromasa y reduciendo su humificación e incorporación en la materia orgánica edáfica. Asimismo, la necromasa de esta exótica presenta una mayor relación C/N que la de plantas nativas, lo que podría reducir la mineralización (Packham et al., 2001). La composición de la necromasa varía entre zonas invadidas y no invadidas por C. edulis, presentando la necromasa exótica una menor concentración de Al (elemento potencialmente tóxico) y de algunos micronutrientes (Fe, Cu). Esto podría reflejar una ventaja competitiva de C. edulis, que restringiría la absorción de Al más eficientemente y requeriría menos micronutrientes (o los reabsorbería más eficientemente de tejidos senescentes). Además, la mayor acumulación de Na en la necromasa con respecto a las nativas podría ser una adaptación a la salinidad, al igual que en otras plantas de la misma familia (Weber y D'Antonio, 1999, Delnavaz Hashemloian et al., 2010). La invasión por C. edulis altera los flujos y reservas de N en la capa de 0-10 cm de suelo, especialmente en los primeros 5 cm. La invasión reduce 8. Resumen general 245 las tasas netas de nitrificación y mineralización, lo que podría conllevar un descenso en la disponibilidad de N y, con ello, un cambio en la composición de las comunidades vegetales (Eviner y Chapin, 2003). La mineralización bruta no varía significativamente con la invasión, probablemente por las pequeñas diferencias en la relación C/N de la necromasa y en la forma de vida de la vegetación nativa e invasora, que es perenne en ambos casos. La nitrificación autótrofa desciende en suelos invadidos (posiblemente por inhibición de las comunidades microbianas por exudación de metabolitos secundarios o acidificación del suelo), mientras que la inmovilización de amonio aumenta (probablemente por una mayor recalcitrancia de la necromasa de C. edulis). Consecuentemente, la invasión por C. edulis conlleva que la inmovilización de amonio supere a la nitrificación autótrofa. En los suelos estudiados (tanto invadidos como no invadidos) el mayor consumo de nitrato fue por reducción catabólica [a pesar de que este proceso está habitualmente asociado a suelos anaerobios (Pandey et al., 2016)], aparentemente producida en micrositios anaerobios (Rütting et al., 2011), y siendo ésta superior en zonas no invadidas por C. edulis, donde la disponibilidad de nitrato para substrato de este proceso era mayor. Los resultados obtenidos enfatizan la necesidad de restaurar los ecosistemas invadidos por C. edulis en los primeros estadios de la invasión, para evitar que los efectos en el suelo se intensifiquen, así como la importancia de eliminar la necromasa exótica para evitar efectos heredados en el suelo. 8.3. POTENCIAL DE SCLEROTINIA SCLEROTIORUM Y PULVINARIELLA MESEMBRYANTHEMI PARA CONTROL BIOLÓGICO DE C. EDULIS Se estudió el efecto individual y combinado del insecto P. mesembryanthemi y del hongo S. sclerotiorum en plantas de C. edulis procedentes de su área de distribución nativa (Sudáfrica) y exótica (NO ibérico). Para ello, se evaluaron los efectos en la planta a corto (índices fisiológicos derivados de medidas de fluorescencia y reflectancia foliar) y largo plazo (supervivencia, crecimiento, relación biomasa Cristina Vieites Blanco 246 aérea/subterránea) en un experimento de invernadero de un año con un diseño factorial [8 procedencias de C. edulis (4 nativas sudafricanas; 4 invasoras ibéricas) x 4 tratamientos (control; infestación con P. mesembryanthemi; inoculación con S. sclerotiorum; inoculación e infestación combinados) x 6 réplicas]. Las plantas de C. edulis infectadas con S. sclerotiorum mostraron descensos transitorios y a corto plazo en el contenido de clorofila y la eficiencia en el uso de radiación fotosintética [sugeridos por el índice clorofílico (CHL-NDI) y el rendimiento cuántico máximo del fotosistema II (Fv/Fm), respectivamente], que no conllevaron efectos significativos en la supervivencia o crecimiento. Sin embargo, la capacidad de infección de este hongo está condicionada por factores climáticos (Berg y Lentz, 1968) y en el campo podría verse favorecido por una mayor humedad ambiental. Pulvinariella mesembryanthemi desencadenó en la planta (aparentemente para contrarrestar los efectos de la infestación) un incremento transitorio del rendimiento fotosintético [como sugieren los índices Fv/Fm y NPQI (índice de feofitinización normalizado)] y un descenso de la relación biomasa radicular/aérea. También, como un posible mecanismo de defensa, en la mitad de las plantas con P. mesembryanthemi se observó la muerte de las partes más intensamente infestadas. Esto podría ser especialmente perjudicial para P. mesembryanthemi por su limitada movilidad (Washburn y Frankie, 1985), que podría dificultar la colonización de plantas sanas. La infestación redujo la supervivencia (especialmente después de los 6 meses, cuando P. mesembryanthemi estaba en estadios avanzados de desarrollo) y el crecimiento de las plantas al cabo de un año. El origen de C. edulis (procedencia nativa o exótica) no afectó a su susceptibilidad a los agentes de control estudiados, pese a lo que cabría esperar dadas las evidencias de adaptación evolutiva encontradas en plantas de zonas invadidas con respecto a las nativas (Roiloa et al., 2016). El uso combinado del hongo y el insecto para controlar C. edulis sólo produjo efectos sinérgicos a corto plazo, no teniendo efecto sobre la supervivencia y el crecimiento. Además, la combinación de los agentes fue perjudicial para el establecimiento de P. mesembryanthemi en la planta. Por ello, el uso individual de este insecto parece ser la mejor estrategia de 8. Resumen general 247 control biológico de C. edulis. Como línea de investigación futura, se considera preciso realizar experimentos con P. mesembryanthemi en condiciones de campo para estudiar posibles interacciones bióticas y abióticas de este insecto en condiciones naturales. 8.4. DISTRIBUCIÓN POTENCIAL Y DINÁMICA DE POBLACIÓN DE P. MESEMBRYANTHEMI EN ÁREAS INVADIDAS POR C. EDULIS Para que un agente de control biológico sea efectivo en una zona, es necesario que su distribución potencial y la de la planta invasora que se quiere controlar se solapen ampliamente (Sun et al., 2017). Por ello, se modelizó la distribución potencial en Galicia (NO España), en base a factores climáticos, tanto de P. mesembryanthemi como de su planta huésped Carpobrotus spp. mediante un algoritmo de máxima entropía (MaxEnt). Además, teniendo en cuenta que los estudios de campo son especialmente importantes en el caso de insectos sésiles como P. mesembryanthemi [que se ven particularmente afectados por factores estacionales y bióticos (Jha et al., 2009)], se realizó un estudio de dinámica de poblaciones. Para ello, se siguió la dinámica de 4 poblaciones gallegas de P. mesembryanthemi durante 2 años y se estudió su relación con factores climáticos (temperatura, precipitación, irradiación) y bióticos (planta huésped, depredadores, parásitos y mutualistas). Las distribuciones potenciales de Carpobrotus spp. y P. mesembryamthemi abarcan toda la costa gallega y se solapan ampliamente, por lo que la distribución del insecto parece estar limitada por dispersión o interacciones bióticas y no por factores climáticos (incluidos en el modelo). Ambas especies se ven favorecidas por una mayor temperatura aérea y edáfica. Además, la distribución de P. mesembryanthemi parece estar relacionada con la humedad del suelo, posiblemente porque su crecimiento y fecundidad aumentan con un mayor contenido de agua en la planta huésped (Washburn et al., 1987). En Galicia, el insecto P. mesembryanthemi tiene dos ciclos por año y su abundancia depende de factores climáticos, alcanzando máximos poblacionales en verano y descendiendo drásticamente en invierno. Sólo en una de las poblaciones (situada en un enclave especialmente Cristina Vieites Blanco 248 resguardado) el tamaño poblacional se mantiene más o menos constante a lo largo del año. A pesar de que en condiciones de campo C. edulis se ve sometida a depredación (caracoles Theba pisana Müller y Helix sp.; ratas Rattus rattus L.; conejos Oryctolagus cuniculus L.) y parasitismo por otros organismos (hemípteros Aphis sp., Pseudococcus sp. y Philaenus spumarius L.), sólo las infestaciones por P. mesembryanthemi parecen afectarla significativamente tanto a nivel reproductivo (número y aspecto sanitario de flores y frutos) como vegetativo (vigor y aspecto sanitario de las plantas evaluado a través de una escala visual subjetiva). Sin embargo, las poblaciones de P. mesembryanthemi podrían estar limitadas por parasitismo (Hymenoptera: Chalcidoidea) y depredación (Coleoptera: Coccinellidae, como Chilocorus bipustulatus (L.), Exochomus sp. y Nephus quadrimaculatus (Hbst.)). A pesar de que la presencia de hormigas estaba ligada a la de P. mesembryanthemi [al igual que en otros Hemiptera, de los que se consideran mutualistas (Buckley, 1987)], éstas no parecen evitar la presencia de parásitos y depredadores. 8.5. VARIABILIDAD GENÉTICA DE POBLACIONES DE P. MESEMBRYANTHEMI A ESCALA MUNDIAL Los estudios genéticos de agentes de control biológico permiten optimizar su efectividad al descubrir complejos crípticos de especies (Gaskin et al., 2011), que pueden tener distinta especificidad y agresividad hacia la planta objetivo (Rauth et al., 2011). Por ello, se realizó un análisis filogeográfico [de poblaciones nativas (Sudáfrica) y exóticas (Europa, Oceanía) de P. mesembryanthemi] a partir de un fragmento de gen mitocondrial (citocromo c oxidasa subunidad I, COI) y de otro ribosómico (segmentos de expansión D2-D3 de la subunidad grande del gen 28S), con el fin de calcular la variabilidad intra- e inter-poblacional y determinar el origen y el número de eventos colonizadores de las poblaciones exóticas. Las poblaciones exóticas no muestran variabilidad genética para ninguno de los dos genes estudiados (incluso comparando poblaciones de Europa y Oceanía), mientras que las nativas son muy variables (existiendo incluso variabilidad intrapoblacional en una de las poblaciones estudiadas). Por ello, búsquedas posteriores de la población más adecuada para su uso como control biológico deberían centrarse en las poblaciones 8. Resumen general 249 sudafricanas, donde se encuentra la mayor parte de la variabilidad de la especie. Además, estos resultados sugieren que las poblaciones exóticas proceden de una única población sudafricana de P. mesembryanthemi. La reproducción partenogenética de este insecto (Nur, 1963) junto con la relativamente reciente introducción en las áreas exóticas [las primeras citas de la especie en Europa y Oceanía datan de los años 80 (Collins y Scott, 1982, Vieira et al., 1983)] pueden explicar que se mantuviera esta homogeneidad genética en las poblaciones no nativas de P. mesembryanthemi. La variabilidad encontrada en las poblaciones nativas de P. mesembryanthemi está por encima del límite intraespecífico delimitado en estudios de insectos de la familia Coccidae (Amouroux et al., 2017, Wang et al., 2015), por lo que podría tratarse de un complejo críptico de especies. Adicionalmente, se secuenció un endosimbionte de P. mesembryanthemi, próximo a la familia Rickettsiaceae (-Proteobacteria) y a los endosimbiontes de otros insectos Hemiptera, que presentaba una muy baja variabilidad genética (incluso comparando poblaciones del área nativa y exótica). La presencia de endosimbiontes en Hemiptera es común (Mathenge et al., 2015, Li et al., 2011) y parece estar relacionada con un mayor rendimiento del huésped (Li et al., 2016). 8.6. BIBLIOGRAFÍA Amouroux, P., Crochard, D., Germain, J. F., Correa, M., Ampuero, J., Groussier, G., Kreiter, P., Malausa, T. & Zaviezo, T. (2017) Genetic diversity of armored scales (Hemiptera: Diaspididae) and soft scales (Hemiptera: Coccidae) in Chile. Scientific Reports, 7, 2014. Anwar-Maun, M. (2009) The sand dune environment. The Biology of Coastal Sand dunes (ed M. Anwar-Maun), pp. 23-39. OUP Oxford, Wiltshire, Reino Unido. Badalamenti, E., Gristina, L., Laudicina, V. A., Novara, A., Pasta, S. & La Mantia, T. 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