ANIMAL SCIENCE, ISSUES AND PROFESSIONS
ENDANGERED SPECIES HABITAT, PROTECTION AND ECOLOGICAL SIGNIFICANCE
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ANIMAL SCIENCE, ISSUES AND PROFESSIONS
ENDANGERED SPECIES HABITAT, PROTECTION AND ECOLOGICAL SIGNIFICANCE
MANUEL ESTEBAN LUCAS-BORJA EDITOR
New York
Copyright © 2012 by Nova Science Publishers, Inc. All rights reserved. No part of this book may be reproduced, stored in a retrieval system or transmitted in any form or by any means: electronic, electrostatic, magnetic, tape, mechanical photocopying, recording or otherwise without the written permission of the Publisher. For permission to use material from this book please contact us: Telephone 631-231-7269; Fax 631-231-8175 Web Site: http://www.novapublishers.com NOTICE TO THE READER The Publisher has taken reasonable care in the preparation of this book, but makes no expressed or implied warranty of any kind and assumes no responsibility for any errors or omissions. No liability is assumed for incidental or consequential damages in connection with or arising out of information contained in this book. The Publisher shall not be liable for any special, consequential, or exemplary damages resulting, in whole or in part, from the readers‟ use of, or reliance upon, this material. Any parts of this book based on government reports are so indicated and copyright is claimed for those parts to the extent applicable to compilations of such works. Independent verification should be sought for any data, advice or recommendations contained in this book. In addition, no responsibility is assumed by the publisher for any injury and/or damage to persons or property arising from any methods, products, instructions, ideas or otherwise contained in this publication. This publication is designed to provide accurate and authoritative information with regard to the subject matter covered herein. It is sold with the clear understanding that the Publisher is not engaged in rendering legal or any other professional services. If legal or any other expert assistance is required, the services of a competent person should be sought. FROM A DECLARATION OF PARTICIPANTS JOINTLY ADOPTED BY A COMMITTEE OF THE AMERICAN BAR ASSOCIATION AND A COMMITTEE OF PUBLISHERS. Additional color graphics may be available in the e-book version of this book.
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Published by Nova Science Publishers, Inc. † New York
CONTENTS Preface Chapter 1
Chapter 2
Chapter 3
Chapter 4
Chapter 5
Chapter 6
Chapter 7
vii Population Dynamics of Endangered English Yew (Taxus Baccata L.): Implication for Conservation and Management Amalesh Dhar, Harald Vacik, Herwig Ruprecht and Raphael Klumpp Endangered Plants: A Comparison of Applied Methods for the Assessment of Extinction Risk for Rare Plants in Italy Innangi Michele, Izzo Antonio and De Castro Olga Effect of Climate Change and Deforestation in a Selection of Vertebrate Species and Opuntia in México Patricia Illoldi-Rangel, Víctor Sánchez-Cordero and Miguel Linaje Is Post-Dispersal Seed Predation a Problem for Spanish Black Pine (Pinus Nigra Arn. SSP Salzmannii) Natural Regeneration in Cuenca Mountains (Spain)? Manuel Esteban Lucas-Borja Mediterranean Horny Sponges: How to Drive a Neverending Story of Exploitation toward a Sustainable Management and Conservation Roberto Pronzato, Fabio D. Ledda and Renata Manconi Endangered Mexican Fish Under Special Protection: Diagnosis of Habitat Fragmentation, Protection, and Future – A Review Ricardo Dzul-Caamal, Hugo F. Olivares-Rubio, Cynthia G. Medina-Segura and Armando Vega-López Ecology, Structure and Conservation Status of the Southernmost Natural Populations of Yew Taxus Baccata L. (Taxaceae) in the Iberian Peninsula Juan Carlos Linares
1
17
49
65
77
109
131
vi Chapter 8
Chapter 9
Index
Contents Europe‟s Threatened Species: The Case of Endangered Phengaris (Maculinea) Butterflies (Lepidoptera, Lycaenidae) Paula Seixas Arnaldo, Maria da Conceição Rodrigues and Teresa Fidalgo Fonseca Plant Conservation Biology: The Case of Helianthemum Polygonoides, a Threatened Species from Southern Spain M. A. Copete, J. M. Herranz, P. Ferrandis and E. Copete
143
155 169
PREFACE Natural ecosystems provide the basic conditions without which humanity could not survive. Good and services provided by ecosystems include for example provision of food, fibre and fuel, purification of water and air, cultural and aesthetic benefits, stabilization and moderation of the Earth's climate, generation and renewal of soil fertility, including nutrient cycling or maintenance of genetic resources as key inputs to crop varieties and livestock breeds, medicines, and other products. However, the ability of biological diversity to continue performing these services is seriously threatened since some species are being seriously deteriorated, and in some cases destroyed. While loss of species has always occurred as a natural phenomenon, the pace of extinction has accelerated dramatically as a result of human activity. Ecosystems are being fragmented or eliminated, and innumerable species are in decline or already extinct. The transition to sustainable development requires a shift in public attitudes as to what is an acceptable use of nature. This can only happen if people have the right information, skills, and organizations for understanding and dealing with biodiversity issues. The book is divided into nine chapters, which are all related to endangered species research studies. Different research groups present here scientific studies and research works on the topic of plant conservation biology, World‟s threatened species and climate change effects on population dynamics. Transferring information and knowledge within the society is crucial for fighting biological deterioration. In addition, promoting the sustainable use of biodiversity will be of growing importance for maintaining biodiversity in the years and decades to come. Manuel E. Lucas-Borja.
In: Endangered Species Editor: Manuel Esteban Lucas-Borja
ISBN: 978-1-62257-532-9 © 2012 Nova Science Publishers, Inc.
Chapter 1
POPULATION DYNAMICS OF ENDANGERED ENGLISH YEW (TAXUS BACCATA L.): IMPLICATION FOR CONSERVATION AND MANAGEMENT Amalesh Dhar1,, Harald Vacik2, Herwig Ruprecht2 and Raphael Klumpp2 1
Mixedwood Ecology and Management program, University of Northern British Columbia, Canada 2 Institute of Silviculture, University of Natural Resources and Life Sciences, Vienna, Austria
ABSTRACT Human interventions and land-use changes effected the structure and species composition of the temperate forests in Europe. The English yew (Taxus baccata L.) populations have been negatively affected by this ongoing process which led to a decrease of most of their ranges all over Europe. Yew is distinct from the other evergreen coniferous tree species: it is a dioecious, long living, non-resinous gymnosperm. Although yew is getting priority for conservation activities the knowledge about conservation management is scarce. Several studies on the ecology, genetics and management of yew populations in Austria try to overcome this knowledge gap. The most significant risk factors for the viability of yew populations are light availability, browsing, tree competition, illegal logging, and lack of public awareness. Findings have shown that a shortcoming of certain regeneration height classes is evident, although most of the population indicated abundant number of one-year seedlings. Considering the vitality of the adult yews it reveals that vitality is related to tree height and DBH as well as influenced by the inter-specific competition of the neighbouring tree species. For the analysis of the genetic structure, English yew populations showed a high level of genetic variation (He = 0.274 and
Email address:
[email protected].
2
Amalesh Dhar, Harald Vacik, Herwig Ruprecht et al. Ho=0.238) with a medium level of inbreeding (0.130). The conservation propositions of English yew in the Eastern Alps are discussed in the light of ecological condition, genetic structure and social aspect of this species.
Keywords: Endangered species; forest conservation; forest ecology; regeneration; Stand structure; vitality
INTRODUCTION The English Yew (Taxus baccata L. Taxaceae) is one of the most ancient tree species in Europe [1]. Spjut [2] mentioned that Taxaceae were radiated from southwest China and evolved into several species in Europe by the Tertiary period. He concluded that T. baccata arose from hybridisation between an extinct Russian species (T. contorta) and Tertiary relics. From the Pollen records it was observed that Taxus was present in Europe during previous interglacial periods, starting with the Cromerian (450 000–700 000 yr BP) [3] but was in maximum abundance in the warm, oceanic climate of the Hoxnian, 367 000– 400 000 yr BP [4]. English yew started declining at lower altitudes before 3000 yr BP, probably owing to human influence, shifted to higher altitudes. In the past centuries yew was over exploited for different purposes in the eastern Alps area [5, 6]. A huge amount of yew was exported from Austria to England for preparing crossbows in the middle age to late 18th century and heavily removed in the last century by human due to the risk of domestic animals especially horses being poisoned by eating parts of the yew [5, 6, 7]. In southern Italy it became locally extinct about 2500 years ago [8]. There are many archaeological records of the use of yew from Neolithic to Roman as, for example, spears, axe shafts, or bows [1, 9]. Yew pegs were also used as fastenings for the Neolithic track ways in the Somerset levels, and more esoterically, yew wood was used to sew together the timbers in the „sewn boats‟ [10]. It‟s tough and long lasting timber was used for buildings and the high aesthetic appeal made yew a popular decorative material [1]. Its bark and needles have been used as an anti cancer medicine containing the chemical Taxol which is one of the best natural sources for cancer treatment [11]. The current geographic distribution of English yew is in the north up to 61° N latitude in Scandinavia and in North Africa, vertically up to 2300 m elevation in the Caucasus [12] and about 1400 m in the Alps [13]. It is scattered throughout Europe [14], northern Africa [15] and the Caspian region of southwest Asia [16], isolated individuals grow within beech forests in northern Spain [17]. English yew is often found in mixed deciduous and conifer forests in Europe. According to Svenning, and Magárd [18] the areas of temperate forest in Europe were reduced and species composition of the remaining fragments changed due to climatic change and human interventions. This ongoing process has negatively affected English yew and increased the risk of extinction from most of its ranges [9, 19]. The main reasons for the decline of yew are widespread deforestation, light competition, selective felling and browsing by herbivore [20, 21, 22, 23]. It is very vulnerable to browsing and bark stripping by rabbits, deer and domestic animals such as sheep and cattle [22, 24] in spite of its poisonous properties. In fact, it is one of the most grazing sensitive trees [24] and a densely deer populated area has a strong negative effect on its recruitment and adult endurance [24, 25]. Another most important factor
Population Dynamics of Endangered English Yew (Taxus Baccata L.)
3
is light, although yew is shade tolerant [26, 27]. Seedlings and saplings often die or show poor growth when English yew grows underneath the shelter of beech [28, 29]. Other factors influencing the viability of yew are (1) adverse soil condition, (2) loss of genetic variation [20], (3) illegal cutting (due to high aesthetic appeal) and lack of people awareness [30], (4) unfavourable site conditions ([9] (5) damages caused by fungi [31] and insect pests which restrict the yew recruitment. At present yew is a rare and endangered tree species in the Eastern Alps [5, 6, 7] with restricted occurrence due to human activities, over uses in the past, browsing pressure, unsuccessful regenerations [30, 32, 33]. Genetic diversity is the basis of the overall biological diversity because it is the key component for adaptation, evolution and survival of species and individuals in the changing environmental condition. From different studies [34, 35] it was suggested that a reduction of genetic diversity could predispose forests species, which are therefore more vulnerable. It is important to study the genetic diversity before developing any conservation and management action. An amazing development has been taken place on the studies of genetic variation in last two decade owing to the application of electrophoretic techniques in population genetics. There are numerous publications reporting on genetic variation of forest tree species, but papers addressing yew are very rare. However studies from Northern Germany [36, 37, 38] Switzerland [39] or Poland [40] provide some basic information on this rare dioecious conifer species in Europe. Those first studies reported on a surprising high level of genetic variation even for small relict populations. Populations of slow growing long living plants like English yew typically received little attention in the past. Due to less awareness, this species is now catalogued as a rare and endangered species prone to extinction from all over Europe [9]. There are two general conservation strategies for slow growing long living species, which are rare and endemic in small geographic areas. On the one hand most plans for management of such species have focused on protection measures with the goal of protecting established individuals [41]. Secondly, the conservation efforts have focused on the reinstitution of ecological processes, which are important for recruitment of new individuals. These efforts promote successful regeneration and increase the genetic diversity on the long run [42]. However, these two conservation strategies may not be sufficient if a population of slow growing, long living plants are predicted to be in long term decline [43]. The study of population ecology for endangered species is fundamental for their conservation management [44]. Therefore it is necessary to understand ecological condition and genetic structure of the species before initiating any kind of conservation activities. The main objective of this chapter is to understand the ecological condition, genetic structure and mechanisms responsible for shrinking yew populations in the Eastern Alps, thereby providing guidelines for its conservation management.
REGENERATION STATUS The natural regeneration is valuable as a progenitor of new stands and the major source of canopy replacement [45]. Moreover regeneration processes are the primary concern for endangered species conservation. The total amount of natural regenerations including all tree species in the Eastern Alps comprising English yew populations varied from site to site
4
Amalesh Dhar, Harald Vacik, Herwig Ruprecht et al.
Regeneration [n ha -1]
(Stiwoll, Bad Bleiberg, Mondsee) (158000 to 51541 n ha–1). Similarly English yew seedling density was also highly varied among the sites (14978 to 5209 n ha-1) which are quite different compared to other European sites (see Figure 1) [1, 43, 46]. Based on the study by Hulme [47] and Svenning and Magárd [18] English yew was found to have limited regeneration potential within beech dominant forests but the results of the population studied in the Eastern Alps indicated that seedling abundance is not strictly controlled by beech dominant forest types. Moreover, some other European studies also reported a negative influence by adult English yew on seedlings establishment [9, 21, 26, 48] which was not the case in this area [19, 30, 49]. However, it is interesting to mention that in spite of a large quantity of seedlings, the average number of survived saplings decreases drastically and turns to zero in certain height class at all sites [19, 49, 50, 51, 52] (Figure 2). This might be due to the impact of several factors simultaneously such as; less availability of light on the forest floor [51], inter-specific or intra-specific competition [53] herbivore browsing [30], accumulation beech litter [19] and site conditions [19]. 100000
Stiwoll
14978
Bad Bleiberg
Mondsee
5209 2392
10000 1000
308
137
100
60 4
10 0
0
0
0
0
1 Seedlings 30-50cm
51-150cm
>150cm < 5cm DBH
Growth stage Figure 1. Regeneration status of English yew population at different sites. Stiwoll
Density [n ha-1]
1000
Bad Bleiberg
Mondsee
100
10
1 5-9.9
10-14.9 15-19.9 20-24.9 25-29.9 30-34.9 DBH Classes [cm]
Figure 2. Stem density of English yew by DBH class at three different sites.
>35
Population Dynamics of Endangered English Yew (Taxus Baccata L.)
5
More than 68 % of the total regenerations were found under dense canopy (≥ 90 %) and this indicated that light is not an important factor for English yew seed germination (Table 2). Similar finding were also reported in other European investigations [9, 18, 26, 29, 48]. However, light requirement of yew seedlings is increased with the increase of age [18, 47, 55]. According to Iszkulo and Boratyński [56] below 2 % photosynthetic photon flux density (PPFD) is enough for seedlings survival up to 2-3 years but their light requirement increases up to 10 times when it reaches the sapling stage. This indicates that quantity of relative solar radiation on forest floor is an important component for English yew seedlings growth and light deficiency could be considered as one of the major reason for seedling mortality [9, 26, 56, 57]. Selective silvicultural thinning operations could be an efficient practise to increase solar radiation on the forest floor [19, 51]. Moreover this (selective thinning) practice might lead to a better environmental condition of the mature yews and will increase the seed production [30, 50]. Herbivores can directly affect the yew regeneration as it is very susceptible to both browsing and grazing ([19, 22, 24, 58]. On unprotected localities, herbivores (deer and roedeer, and other mammals) can reduce or even completely eliminate all yew seedlings [29, 46, 47, 57]. Moreover, sometimes repeated browsing can alter the crown form irregularly and reduce the overall growth which may favour pathogen infestation to the plant [46, 59]. Establishment of fence or reduce the number of herbivory may decreases the intensity of browsing pressure and increases the rate of adult survival [51]. While Piovesan et al. [59] reported that establishing a number of watering places further away from yew populations and improving the grass cover around the watering sites can reduce the grazing pressure. Such habitat improvement activity may also be applicable to reduce the browsing pressure in the Eastern Alps. From the above discussion it can be concluded that the impact of different biotic and abitic factors have to be considered before initiating any kind of management strategies for maintaining the English yew regeneration.
STAND STRUCTURE Most of the yew stands in the Eastern Alps are characterized by beech dominant unevenaged forest where English yew occupied the forest understory [53]. The density (492 to 45 n ha-1) and basal area (3.8 to 1.04 m2ha-1) of the overall established yew trees varies from site to site (see Table 1). Density of pole stems (DBH ≥ 5 cm) varied from site to site due to non homogenous site conditions. Site (Stiwoll) characteristics like west exposition, sub montane region, cambisols, and fresh water balance might have a positive impact on the growing conditions for yew population [51, 54, 55]. Depending on the sites, density of English yew declined both linearly and nonlinearly with increase of DBH class (Figure 2). The mean height and DBH of English yew ranged from 6.3 m to 7.6 m and 8.8 to 17.0 cm (Table 2). Iszkuloet al. [55] reported that average height and DBH was lower in shadiest place compared to those of growing in better light conditions. These finding are comparable with Stiwoll because the average canopy closer of this site is 84 %. Moreover, the reproduction (by seed) of yew can be enhanced by better light availability too [18, 60].
6
Amalesh Dhar, Harald Vacik, Herwig Ruprecht et al. Table 1. Stand demographic attributes of three different sites in the Eastern Alps Demographic attribute
Investigated population Stiwoll Bad Bleiberg 4.55 18.4 1500 1335 1008 1239 41.53 31.40 418 276 50.6 68.0 35.6 35.0
Area [ha] Overall tree density [nha-1] Other tree species density[nha-1] Overall basal area [m2ha-1] Overall Tree volume [m3ha-1] Overall Max. DBH [cm] Overall Max. Height [m]
Mondsee 2.6 758 661 36.71 384 108.4 34.5
Table 2. Comparison of English yew tree attributes at three different sites in the eastern Alps Site
Densit y [nha-1]
Volume [m3ha-1]
Basal area [m2ha1 ]
Stiwoll Bad Bleiberg Mondsee
Ave. height [m]
Ave. DHB [cm]
Damag e [%]
Ave. foliage density [%]
Ave. crown [%]
Ave. canopy closure [%]
492
17.3
3.2
6.3a
8.8a
34.2
80.3
62.3
84.3
45
5.1
1.1
7.6b
16.3b
28.3
65.2
65.4
68.2
97
16.7
3.8
7.5b
17.0c
35.3
85.1
68.3
71.1
[Different small letters indicate significant differences (Duncan Multiple Range Test, P ≤ 0.05) among the reserves].
Structural diversity generally describes the species composition, horizontal and vertical variation within the forests stands, which is an important parameter to initiate appropriate forest management actions. Considering the species mixture the population of Stiwoll represent the less inter-specific and high intra-specific competition while Bad Bleiberg population showed high inter-specific and less intra-specific competition. The evidence of illegal logging was observed at all sites across the Eastern Alps and considered one of the important factors related to conservation of English yew populations [30, 49, 50, 52]. Local people are not aware about the importance of this endangered species. Therefore initiating public awareness programmes may help to reduce illegal cuttings and increase knowledge concerning the importance of yew conservation.
HEALTH CONDITION Considering the tree health most of the Eastern Alpine yew population showed good condition and it ranges from 79% to 48 % (Table 3). [19, 53]. The vitality condition of English yew is related to the tree height and DBH. With the increase of DBH and height English yew represent better vitality class (Figure 3). Such relationship is not statistically proven but it shows a trend. The yews with a higher vitality classes represent the higher mean DBH (13.3 cm) and height (7.9 m) compare to lower vitality class (see Table 3). The inter-specific competition
Population Dynamics of Endangered English Yew (Taxus Baccata L.)
7
also plays an important role on yew vitality. For example, the lowest vitality (53 % less vital to least vital) site has the highest number (1239 nha-1) of other trees species (Table 1). Ruprecht et al. [53] reported that vitality of each individual yew was influenced by the inter-specific competition of the neighbouring tree species and yews themselves did not affect the vitality of neighbouring yews. They also reported that vitality of English yew is increasing with the increasing of tree-tree distance and decreases with negative tree height differentiation. According to USDA Forest Service [61] 500 healthy mature individuals per stand is the minimum amount currently considered necessary to safeguard population genetic variability. Most of the yew populations in the eastern Alps maintain this limit (Table 5).
Figure 3. Box plots representing the over all relationship between vitality class with DBH and tree height (all three sites combined).
8
Amalesh Dhar, Harald Vacik, Herwig Ruprecht et al. Table 3. Tree height, diameter at breast height (DBH) of each vitality classes and overall vitality percentages of three sites
Vitality class
Total number of yew
Very vital Vital Less vital Least vital
870 1417 738 235
Tree height
DBH
Overall vitality in different reserve [%]
7.9(±1.9)d 6.7(±2.0)c 6.0(±2.0)b 5.3(±2.2)a
12.3(±5.0)c 10.9(±5.4)b 10.6(±4.9)b 9.9(±3.7)a
Stiwoll 32.4 46.8 17.4 3.4
Bad Bleiberg 5.7 41.7 34.9 17.8
Mondsee 49.0 14.3 30.6 6.1
[Numbers in parenthesis indicate standard deviations; different letters indicate significant differences (Duncan Multiple Range Test, P ≤ 0.05)].
Almost one third of the total yew individuals have some kind of visible damage (ranges from 35.3 to 28.3%) ranges from 35.3 to 28.3% across the different sites in the Eastern Alps. These indicate that vitality of the individual yew is not strongly related with stem damage. From different studies it was reported that English yew is damage tolerant species [26, 27] and there was no significant interaction between yew vitality and damages [51, 53].
Sex Ratio The sex ratio of mature fertile individual is an important parameter because formation of viable seeds requires both sexes co-exist on the landscape and only a balanced ratio of sexually mature individual could provide a good reproductive ability as well as future progeny production. This would permit the preservation of the natural population dynamics and long-term evolution at the landscape level. Eastern Alpine yew populations showed mixed result up to now. Two populations showed a balanced sex ratio whereas stiwoll showed female-biasness (Table 4). Svenning and Magárd [18] reported that the population of Munkebjerg was female biased whereas Willamson [48] mentioned male biased at Kingely valley in England. On the other hand Hilfiker et al. [39] noted that a small population showed a female biased sex ratio in Switzerland. This is not the case for the Eastern Alpine populations (Table 4), where two small populations represent a balanced sex ratio. Moreover, in case of the large population like Stiwoll, “sex biasness” is not a general one, as more than 1/3 population were unidentified up to now due to its dioecious sexual system. So the final sex ration could be changed after completion of full assessment. A similar observation like Eastern Alpine was reported by Iszkulo et al. [62]. According to them sex biasness in yew is not related to population size, but a positive correlation between female frequency and precipitation were observed. The latter explanation leads to the hypothesis of reduced drought tolerance of female yew in comparison to male individuals [62]. Table 4. Sex ratio of selected English yew populations in the eastern Alps Population Stiwoll Bad Bleiberg Mondsee
Total no. of individuals 2236 828 253
Female 835 392 78
Male 535 432 73
Undefined 866 4 98
Ratio 1.56 0.91 1.07
Population Dynamics of Endangered English Yew (Taxus Baccata L.)
9
GENETIC VARIATION AND DIFFERENTIATION Knowledge regarding the distribution patterns of genetic variation in plants and among the populations is an important key factor for developing conservation strategies. It is mentionable that English yew has a widespread geographic distribution throughout Europe and occurs rather sporadically in most of its range. The genetic structures of Eastern Alpine yew population were described based on the isozyme analysis. In total 9 isozyme gene loci and 32 alleles were identified. The average number of allele per locus (A/L) varied from population to population (2.8 to 2.4 A/L) whereas the average effective number of allele per locus was very close among the populations (Table 5). Considering the heterozygosity Eastern Alpine yew populations showed wide range of variation. The range of expected (He) and observed (Ho) heterozygosity was 0.230-0.304 and 0.178-0.272 respectively. The average expected and observed heterozygosity for the seven populations was estimated to be (Ho) 0.238 and (He) 0.274 respectively (Table 5). This indicated that the analysed yew populations in the Eastern Alps showed a high level of genetic variation with some differences compared to others studies in eastern and central Europe with respect to observed (Ho = 0.238) and expected (He 0 0.274) heterozygosity [37, 63] and showed a similar trend like Asian [64]and American species [65]. From the different studies it was assumed that the needle or apical meristeme based analyses showed lower values of Ho in comparison to He [64, 65] and balanced the value with respect to microgametophyte based studies [66]. Besides these, some other factors such as breeding system, mechanisms of seed dispersal, geographic distribution, and ecological amplitude could influence the heterozygosity [67]. The level of inbreeding which was described by Wright‟s fixation index (F) showed an overall considerable deficiency of heterozygotes relative to Hardy-Weinberg expectation. The range of F values was 0.066 to 0.228 with an average of 0.131. This indicated that yew populations in the Eastern Alps have a moderate excess of homozygotes [68]. This moderate level of inbreeding can result from a variety of causes such as positive assortative mating (mating among the similar genotype) [69]; selection for homozygotes; family structure within a restricted neighbourhood causing mating among relatives [70] and finally the Wahlund effect caused by the artificial grouping of individuals from different breeding populations [71] or due to the sampling procedure and size. Therefore further intensive investigations with different DNA marker could help to find the most relevant influential factor. Table 5. Genetic diversity of selected English yew populations in the eastern Alps (compiled from Dhar [54]) Name of the Province
Population
Steiermark [WG 5.3] Känten [WG 6.1] Upper Austria [WG 4.1] Upper Austria [WG 4.1] Upper Austria [WG 4.1] Lower Austria [WG 5.3] Lower Austria [WG 5.1] Average
Stiwoll Bad Bleiberg Mondsee Almtal Losenstein Hirschwang Piesting
Elevation [m] 580-700 900-1300 480-530 460-490 540-680 700-900 400-650
Total number of English yew 2236 828 253 >2000 1815 ~ 80 >100
Sample size 109 75 95 121 122 40 61 89
Parameters A/L Ho 2.8 0.232 2.8 0.178 2.4 0.257 2.8 0.228 2.7 0.272 2.7 0.242 2.6 0.260 2.7 0.238
He 0.267 0.230 0.304 0.267 0.299 0.270 0.279 0.274
[F] 0.124 0.228 0.155 0.149 0.089 0.105 0.066 0.131
[Mean number of alleles per locus (A/L); Observed heterozygosity (Ho); Hardy-Weinberg expected heterozygosity or genetic diversity (He); Wright‟s Fixation index (F)].
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Amalesh Dhar, Harald Vacik, Herwig Ruprecht et al.
RECOMMENDATIONS FOR CONSERVATION MANAGEMENT Based on the recent studies and inventories in the Eastern Alps the recruitment of English yew was least or absent [19, 49, 50, 52, 54] which is one of the most important concerns from the conservation perspective. The least or absent recruitment could shrink the future effective population size which has direct consequence on genetic variation within population, gene flow and potentiality of adaptive traits in the changing environmental condition [68]. Moreover the studied populations at these sites showed a medium level of inbreeding although the pattern of genetic variation is high compared with other conifer species [68]. This medium level of inbreeding can create detrimental effects for in situ management and adaptive traits in the changing environmental condition of these populations in the future. In the light of conservation management English yew populations in the Eastern Alps have to ensure the recruitment for maintaining the population size which also promotes the natural selection and adaptations to minimise genetic drift in the long term. Based on the above discussion following points need to be considered with regard to conservation and management of English yew populations in Eastern Alps area:
During Planning Activities
A multi level approach (e.g. game management, silvicultural treatments, rising people awareness) is required to meet the demands of sustainable conservation strategies for yew populations. A Population Viability Risk Management (PVRM) approach could be helpful in order to identify courses of action for improving the viability of endangered populations see Vacik et al. [50] and Dhar et al. [30]. Public awareness programmes will help to enhance the knowledge about the ecological importance of yew for the local people. Regular information and publications might help to increase the level of awareness and improve the overall knowledge about this species and have an effect on the reduction of illegal logging. Beside this, the presence of research activities will have positive effects on the public awareness too.
During Monitoring Activities
The establishment of fences in the studied forests will help to record and quantify the impact of browsing on the regeneration. Additionally the reduction of the roar and red deer population is a prerequisite for further conservation activities. Even without a clear quantification of the severity of the browsing impacts it seems not rationale to plan conservation activities ignoring the negative effects of ungulates. It is very important to monitor the reproduction rate and sex ratio of the mature yew individuals because formation of viable seeds requires both sexes co-exist on the landscape. This would help to understand natural population dynamics and support the long-term evaluation at the landscape level.
Population Dynamics of Endangered English Yew (Taxus Baccata L.)
11
The methods of seed dispersal are an important parameter for dioeciously plants. So, more attention has to be paid on a better documentation of the role of birds and mammals that are responsible for seed dispersal. Within a multi-level planning approach such research activities could be included. The regeneration status needs to be evaluated in 5-10 years cycles. The analysis of the survival rate of the young yew individuals will allow sound recommendations for conservation activities in the future.
Operational Guidelines for Management Activities
Continuous selective thinning will enhance the light availability on the forest floor and in the tree layer. This improves the overall health condition for the pole stand and increases the rate of survival of English yew seedlings. It also helps to minimise the intra-specific and inter-specific competition as well as improve the horizontal and vertical structure of the forest. The thinning will increase pollen disperse within the forest and reduces the risks of inbreeding as well. However sound and professional harvesting operations is recommended during selective tree felling to minimize fungal or biotic damages.
ACKNOWLEDGEMENT We would like to thank Mr. Schuster and the owners of the gene conservation forests for their interest in conservation management and their lively support during the field studies. We also thank the Österreichische Bundesforste AG (ÖBF) and the Forest Province Office of Styria for financial support as well as the Österreichische Orient-Gesellschaft (ÖOG) for OneWorld Scholarship.
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In: Endangered Species Editor: Manuel Esteban Lucas-Borja
ISBN: 978-1-62257-532-9 © 2012 Nova Science Publishers, Inc.
Chapter 2
ENDANGERED PLANTS: A COMPARISON OF APPLIED METHODS FOR THE ASSESSMENT OF EXTINCTION RISK FOR RARE PLANTS IN ITALY Innangi Michele1, Izzo Antonio2 and De Castro Olga1, 1
University of Study of Naples Federico II, Dept. Biological Science, Sect. Plant Biology, Via Foria, Naples, Italy 2 University of Study of Naples Federico II, Botanical Garden, Via Foria, Naples, Italy
ABSTRACT In this chapter we discuss how to assess the risk of extinction for plants using two different methods and six different species as case of study from the Mediterranean, namely from Campania, an administrative region of Southern Italy. Italy falls within one of the most important hotspots for biodiversity of the world: the Mediterranean region. In Italy, plant biodiversity is particularly rich, with some 7500 species with a high rate of endemics (13.5%). A great number of these species is currently exposed to various threats in spite of the existence of international, national and regional laws to protect them. Institution of protected areas or measures for species conservation is often ineffective without an assessment of the actual risk of extinction of a taxon. The most acknowledged method to evaluate this risk is the IUCN protocol for the creation of Red Lists, which is widely used worldwide for both animals and plants. Other approaches have been developed as well, like the creation of the Red Numbers in Israel, which is designed specifically for plants. Both procedures point out the importance of an evaluation as much quantitative as possible in order to avoid excessive subjectivity when measures are taken to preserve one species rather than the other. In this chapter, even though these two different methods are not meant to be alternatives, we will juxtapose them using as case study the plants growing in Campania in order to allow a comparison of both methods at a regional scale and to detect possible flaws in either methodology, in terms of conservation of plants, on
Corresponding author:
[email protected].
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Innangi Michele, Izzo Antonio and De Castro Olga the basis of detailed data which have been collected during field surveys. The applied approach highlights the challenges and complexity of evaluating the risk of extinction for plants, which is the basis for conservation of both species and habitats.
INTRODUCTION In this chapter we will discuss about how to assess the risk of extinction for plants using two different methods and six different species as a case study from the Mediterranean, in particular from Campania, an administrative region of Southern Italy. We will see that, even at a regional scale, evaluating which plant is in need of more urgent conservation. How much it is endangered and when it is likely going to be extinct is no easy assessment, even at a local scale. The goal of this chapter is to juxtapose these two methodologies on the basis of the same regional data, confronting the results and discussing further needs in the development and application of risk assessment methods for plants and their conservation. It is not the aim of this chapter to assess which method is better than the other, as they are conceived to be neither complementary nor in opposition with each other, but only to present a case study with both approaches and discuss the results, which can be very useful and indicative to all people dealing with conservation of plant species and habitats, especially at a regional level. In the following introduction paragraph, we will outline why plant biodiversity is important and why it is threatened, especially in the Mediterranean. We will also summarize the threats to biodiversity in the Mediterranean and what are the main initiatives to repel them. As every organism is important in its ecosystem, it is very well known that without primary productivity in the sea and on the mainland, life on Earth would not exist, at least not as we know it. Accordingly, protection of plants should be of the highest priority for all scientists and organizations dealing with conservation when a plant species or, worse, a plant community is threatened with extinction. Extinction is a natural process. Species can survive, evolve or go extinct, there is no other way. In the long history of life on this planet, the vast majority of taxa are now extinct, and entire groups of plants and animals are known only by fossil records. Competition between species, climatic changes and catastrophes are all natural phenomena that can result in extinction events [1]. In the last millennia, humans have shaped landscapes, tamed animals, cultivated plants, moved seeds or livestock from one continent to the other, annihilated species for hunger or sport, causing extinctions that are no longer considered fully “natural” [2]. Indeed, according to the International Union for Conservation of Nature (IUCN), almost 800 species have gone extinct after the year 1500 AD (an arbitrary date chosen to define the beginning of the “modern age”) and most of these extinctions are directly or indirectly related to human activities [3]. Not-sustainable impact of humans on wildlife has been almost ignored until recently, in the 20th century, when scientists became more and more aware of the selfdestructive behavior of local governments regarding the environment. Thus, national and international initiatives exist nowadays in order to preserve wildlife and habitats, including plants, from the risk of extinction driven by human impact.
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The cradle of modern Western society is the Mediterranean basin; therefore here the impact of humans on the environment is very high and very old within one of the most important hotspots for biodiversity in the world. The Mediterranean has a peculiar climate and its plant species diversity is comparable to that of tropical rainforests [4], with more than 25,000 plant species and almost half of them endemic to the region [5]. The climatic features of the Mediterranean also exist outside of Europe defining the Mediterranean biome, which is a global conservation priority [6]. The overall extent of this biome covers just 2% of the World‟s surface, but it hosts 20% of the known vascular plant diversity [4]. The level of threat to this biome is impressive: by 2100, the Mediterranean biome will likely experience the largest proportional loss of biodiversity of all terrestrial biomes due to its significant sensitivity to multiple biodiversity threats and interactions among these threats [7]. Considering only the Mediterranean in Europe [as we will do for the rest of this chapter], this hotspot for plants is very important especially for endemic and often rare plants. Endemics are roughly 13,000 species, existing in this part of the world because of the significant number of distinct elevation belts, the high geological diversity, the sharp latitudinal gradients, the broad oceanic‐continental gradients from the coastal areas to the inner mountain regions, and the frequent isolation of mountains that all contribute to the high diversity of the Mediterranean mountain and island flora [8]. In point of fact, the complex system of Mediterranean mountains is extremely rich in endemic species because of the climatic events in relatively recent geological times. Above all, climatic cycles of the Pleistocene drew most of the patterns of species richness and endemism that we see nowadays [9]. This led to the existence of so-called refugia in the mountains of the Mediterranean, where species could survive the dramatic climate changes [10]. Undoubtedly, climate change is a major threat in the Mediterranean, especially for plants. According to Giorgi and Lionello, given the most recent global and regional model simulations, we can suppose :«a collective picture of a substantial drying and warming of the Mediterranean region, especially in the warm season [precipitation decrease exceeding - 2530% and warming exceeding 4-5°C]. The only exception to this picture is an increase of precipitation during the winter over some areas of the northern Mediterranean basin, most noticeably the Alps. Inter-annual variability is projected to generally increase, as is the occurrence of extreme heat and drought events» [11]. What is really going to happen is hard to forecast, but it is not arguable that climatic change along with the relentless expansion of invasive alien species is changing not only the flora of the Mediterranean, but also its vegetation. As outlined by Heywood: «the Mediterranean region plays a unique role in the context of climate change and its effects on biodiversity as it acts as a barrier to migration of many plants from south to north within the time-scale of concern. Because of the lack of a comparable Saharan hinterland that characterizes the corresponding North African climatic belt, a novel climate will develop in Southern/Mediterranean Europe as a result of climate change and it is difficult to envisage the kind of vegetation that will occupy this in the absence of large-scale migration of species from North Africa although long-distance dispersal will allow some species to migrate. It will be vulnerable to weedy or invasive species: existing ones will be expected to persist or expand their distributions and new ones take hold» [12].
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In light of this knowledge, the need for effective and immediate conservation strategies is evident. A recurring dilemma among scientists is whether it is best to protect species or ecosystems, although there is a strict interdependence between these two approaches. In recent years, ecosystem-based approach has been implemented as a strategy for integrated management of land, water and living resources that promotes conservation and sustainable use of these resources in an equitable way [13]. However, when dealing with conservation of a particular species, this can be achieved in situ [by protecting the plant in its natural habitats, mostly using protected areas] or ex situ (typically in gene banks or botanical gardens). Sometimes, plants can be reintroduced in locations where they were historically present but erased by human action; this approach is frequently called inter situs [14]. Protected areas are at the present time the main way to dam the ongoing expansion of human not-sustainable exploitation of the environment. In an international context, they were defined as «an area of land and/or sea especially dedicated to the protection and maintenance of biological diversity and of natural and associated cultural resources, and managed through legal or other effective means» by the Fourth World Congress on National Parks and Protected Areas in 1992 [15]. According to the 2003 United Nations List of Protected Areas, there are more than 100,000 Protected Areas covering more than 11% of the Earth‟s terrestrial surface [16]. In the Mediterranean, there are many areas area which are protected by international conventions, programs or directives such as:
Ramsar sites; Council of Europe Biogenetic Reserves; Council of Europe European Diploma of Protected Areas; UNESCO Man and Biosphere Reserves; World Heritage sites: Natural sites; Barcelona Convention: Special Protected Areas; Habitats Directive Sites of Importance; EU CORINE Programme Biotopes; WWF Forest Hotspots.
Nonetheless, even though there are many international initiatives dealing with conservation, not all countries have the same concern about biodiversity or, albeit there is a real policy to safeguard wildlife through a system of protected areas, actual knowledge of species‟ status is far from being complete. Only for endemics to the Mediterranean, it is not clearly known how many of these species are endangered. National Red Books for some European countries exist, even if they are not always up to date. Just for Italy, Greece, Spain and France, there are probably some 4000 species threatened [5]. In-depth knowledge of biology, ecology and population size is a mandatory source of information to understand which species are in need of more protection and where they are. The where factor is particularly important, because a species can be rare and/or endangered in some political/geographical region while being widespread elsewhere. The species which are more likely at risk of extinction in a few years are those that require immediate protection measures, in situ and ex situ.
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But how do we assess if and when a plant is heading towards extinction? A real forecast, as we will see, requires a huge amount of information, mostly via surveys on the field. Moreover, a reliable estimate of a plant‟s risk of extinction cannot be achieved by using the same methodology you would use for animals without adequate modifications. Probably the best-known and widespread methodology to assess if a species is endangered with extinction is the one from IUCN in order to prepare Red Lists of threatened species. It is a methodology that can be applied to all life on Earth, including animals and plants, with specific guidelines if the Red List is meant for the whole world, or for a single country or a specific region [17]. Additionally, a methodology developed specifically for plants on a regional basis is the one for the construction of Red Numbers for Israeli flora, which is not meant to be an alternative or in opposition to IUCN, but a complementary instrument at a regional scale to assess conservation priorities [18]. We will apply these two methods on a regional level, using as case studies six plants which are considered rare and/or endangered in Campania, a region in southern Italy in the Mediterranean basin. These plants all have peculiar habitats and/or interesting biogeographical features for the region and have been chosen for the case study because a lot of detailed field data are available due to intensive field surveys in this region. An outline of the origin of flora and vegetation of Italy will be followed by a synthetic explanation of the aforementioned methodologies. Then we will present the available data for the six plants from Campania and apply the methodologies. In conclusion, we will have a comparison and a discussion with remarks on advantages, disadvantages and flaws in the two methods for our case study.
PLANT DIVERSITY IN ITALY Italy plant diversity is undoubtedly one of the richest and most interesting in the World, surely the most diverse flora in the European Union. The vascular flora consists of 7,634 entities, namely 6,711 species and 2,125 subspecies. Overall, 196 families are represented and 1,267 genera. It should also be mentioned that among these species, 751 are non-native, 11% of the total flora, but this is a value still much lower than in other European countries [19]. More and more taxa are enlisted in the flora every year. This species richness is explained by the peculiar geology, ecology and biogeography of the Italian peninsula, which is a core area in the Mediterranean. As a matter of fact, the flora of Italy is so special because, on the one hand, there are a great variety of environments with complex lithological, topographical and climatic conditions. On the other hand, paleogeographic and paleoclimatic history made the peninsula a place of convergence for plants coming from wide ranges of the World. The sole geographic location of Italy could be enough to explain this high biodiversity; however, it is primarily the geological history of the territory that has contributed to the increase in the number of species [20]. As we already mentioned in the previous paragraph, over millions of years the Mediterranean has undergone major changes because of the transformation of the Tethys basin, the different orogeny phases, the long stages of emergence of land and the ice ages. Italy, consecutively, is centrally located in the Mediterranean and the Tyrrhenian basin, which represents its geometric center, and is surrounded by the peninsula and islands. This position
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has no doubt helped, and continues to facilitate, the phenomenon of colonization by species from surrounding lands, particularly from the West, from the South and East [2]. All the complex phenomena of orogeny and tectonics leading to the current configuration of the Italian territory began in the Tertiary. The Alps mountain range originated about 50 Ma, while the Apennines rose just 5 Ma. In some circumstances, floristic differences that are found in cases of intrageneric vicariance reflect very precisely the edges of tectonic limits, especially in the Alps [20]. Peninsular Italy, as was said, did not have the Apennines up to 5 million years ago. Before their advent, geological and biological evidence suggests the presence of two land areas that made a bridge between the Sardinian-Corsican area to the West and the GarganoApulian zone to the East, while in the middle there was a large sea. From this sea, as a result of tectonic movements, first arose an archipelago and then the Apennines [21]. As a consequence, on the mountain peaks of the Apennines there are several neoendemics and many species are shared between the coast of the Tyrrhenian and the Adriatic. During the Tertiary, the coasts of the Mediterranean were very different from today. At that time it was connected with the Atlantic Ocean, which at that moment was just forming, and the Thetis, a large ocean that covered much of Eurasia. Subsequent climate changes transformed the Mediterranean climate from equatorial to subtropical. In the Messinian age (5-6 Ma) the threshold of Gibraltar rose, and the Mediterranean, along with the Tethys, became a closed basin [21]. A Saharan-like climate resulted in an impressive evaporation of the sea, evidenced by the large salt deposits virtually ubiquitous in the levels of the Messinian. This period facilitated the entry of Italy and in the Mediterranean of plants adapted to the windy ridges as Astragalus spp. and Genista spp. along with many plants adapted to salty or brackish soils. Above all, in this period, sea cliffs species such as Limonium spp. arrived, plants showing clear similarities with species of desert mountainous environments [20]. After the Messinian crisis, the end of the Tertiary is characterized by a cycle of marine transgression, which at times led to an increase in sea level of nearly 300 m. The floristic richness in the smaller islands appears to be closely related to these phenomena, because it is much higher in those islands that exceed an altitude of 270-280 m than in the lower ones, where the original flora was probably deleted by the sea and regenerated only after the emersion. The Quaternary is characterized by a very harsh climate, with peaks of cold that go under the name of glaciation or ice ages. At least four of these periods had a particular impact on Italian biogeography, both animal and plant. During the glaciations, the Alps were characterized by a continuous ice cap (Inlandsis) which reached the plains below as well. The Apennines presented numerous glaciers, quite large ones even if more fragmentary than the Alps. The vegetation, in these times of crisis, continued to survive in small patches only on the highest peaks jutting from surrounding glaciers, a peculiar kind or refugia known as nunatakker. In these areas, even when they are not recognizable by geomorphological evidence, it is very common to find the presence of paleoendemics testifying the ancient climate and vegetation. The cool climate allowed the advance of many species associated with this type of climate along the Apennine ridge, with the formation of vicariances, paleoendemics and neoendemics [20]. As a final point, it cannot be overlooked that during the last glacial period, the Würm, about 50,000 years ago, Homo sapiens made their appearance in Italy and began to severely
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affect the biological evolution of both plants and animals. In the post-glacial period, between 15,000 and 10,000 years ago, humans introduced grazing livestock and its impact on the flora and vegetation became far from negligible, driving radical and frequently irreversible changes in the landscape [2]. The geological, biogeographical and climatic history of Italy is the source of a very diverse pattern of landscapes, environments, habitats, soils and biodiversity. Plants, as we have said, offer an incredibly high diversity in species and communities even in small territories [19]. In terms of the phytogeography, Italian vascular flora can be attributed to nine main chorotypes. The largest groups are the Eurasian (20.93%) and Steno-Mediterranean species (16.65%). A remarkable figure is that of the Endemic species (13.50%). At regional scale, as might be expected, the two island regions of Sicily and Sardinia have the highest values of endemism [22]. Undoubtedly, the most interesting group are endemic species, because it not unusual that they are at a high risk of extinction. Most of the time, these species are very rare and/or with a scattered distribution because they are usually tied to specific and often delicate habitats. These habitats can be precarious for both ecological factors and for human impact. Therefore, the importance of protection of endemic plant species is essential for the conservation of habitats and ecosystems as well. To close this section, we will now mention just a few particularly interesting examples of rare and/or threatened species in the Italian territory, also describing their peculiar habitats, within or outside Italian protected areas (Figure 1). These few cases are not meant to be exhaustive, but only indicative, because there are countless examples in such a marvelous territory and it is not the aim of this chapter to describe in full detail plant biodiversity in Italy or in the Mediterranean. Once again, it is important to highlight that, even though endemic species are usually very important in terms of conservation, even widespread species can be, locally, very rare and worthy of protection. Besides, endemism is an important word to identify a taxon living in a definite geographical region, but the word does not imply how large this region is. For instance, we can say that Pistacia lentiscus L., the Mastic, is endemic to the Mediterranean but it cannot be said it is an endangered plant whereas, if a natural stand of Mastic should be found in Sweden, that hypothetical spot of Mediterranean flora found so north of its natural range would certainly be worthy of the highest priority for protection. This is just a fictional example, but many similar cases occur in reality for species very common in their natural range but somewhere unusual, sometimes witnessing important biogeographical phenomena that happened ages ago. Given these preliminary remarks, mountain flora and vegetation of Italy, as with all the Mediterranean, offers the vast majority of endemic and/or rare species. Several taxa are surely quite common and widespread, but many others, even if not so rare, are of great interest for conservation, ecology and biogeography. A widespread species in Europe and Northern Italy is the silver fir, Abies alba Mill., which cannot be enlisted as a rare plant. Nonetheless, this plant, which grows in cold and wet climates at altitudes between 900 and 1800 m, has a patchy distribution in Italy, becoming increasingly rare as you move down from the Alps to reach the southern Apennines, where stands of silver fir are significantly few. All this is linked to the cycles of Quaternary ice ages, which initially favored the first advance of A. alba in southern Italy which was later decimated by the changing climate, reducing it to survive only in nunatakker environments.
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Most likely, by some fir population reaching Sicily, through a striking geographical and genetic isolation, a neoendemic fir species originated in the Madonie Mountains. This species, A. nebrodensis (Lojac.) Mattei, is currently one of the rarest and most threatened plant in the World, because the population is extremely small (less than 30 mature individuals) which are fenced and protected in every possible way [23].
Figure 1. Protected Areas in Italy.
The Madonie Mountains are one of the many examples of areas that are home to unique habitats occupied by many species of particular interest in Italy. Besides the aforementioned A. nebrodensis, species extant only in the Madonie Mountains include Petagnaea gussonii (Spreng.) Rausch., a rhizomatous plant surviving only in circa 20 locations in damp and shaded habitats [24] and Peucedanum nebrodense (Guss.) Strobl., which lives in only one location at 1,800 m on a rocky slope [25]. After mountains, smaller islands and archipelagos of the peninsula are very interesting for their rich and diverse flora. The Tuscan Archipelago, the Pontine Islands, the archipelago of the Aeolian Islands and the Campania Archipelago, just to name a few, are plentiful in endemic species of considerable value. We could mention, for instance, Silene hicesiae Brullo et Signorello or Centaurea gymnocarpa Moris et De Not., both included in the Top 50 most endangered species of the Mediterranean islands according to IUCN [26]. The first, which is endemic to the Aeolian Islands, grows on rocky slopes of two volcanic islets; Panarea (where the population covers an area of 3-4 hectares, with nearly 400 individuals) and Alicudi (with a population of less than 30 specimens covering 60 m2). The second one is endemic to the island of Capraia in the Tuscan Archipelago and is a species that colonizes cliffs‟ cracks and crevices.
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Rock faces and screes, no matter if on a high mountain range or islands, are the richest areas in terms of endemic and/or rare flora of the Italian peninsula. Each region hosts many interesting species associated with these environments, habitats that can be either precarious and subject to human impact or isolated and unreachable by humans [19]. It is not thinkable to list all of these species, but it is worth mentioning a few. Dianthus rupicola Biv. grows on coastal cliffs in some areas of southern Italy and is an important species among the Community included in Annex II of Habitats Directive [27]. Included in the aforesaid Directive is also Primula palinuri Petagna, found on some cliffs overlooking the sea of Campania, Basilicata and Calabria. At higher altitudes we can mention Globularia incanescens Viv., endemic to the Northern Apennine growing mainly on calcareous rocks between 400 and 2,000 m, and Aquilegia champagnatii Moraldo, Nardi et La Valva, a plant with beautiful flowers endemic to Campania and growing on very few vertical limestone cliffs in the Picentini and Lattari Mountains. Given the geology, geomorphology and the complex hydrographic network of the peninsula, wetlands, swamps, lagoons and aquatic habitats are very abundant. These ecosystems are often among the most threatened on account of their delicate ecosystem balance. The issue of damp habitats is certainly one of the most worrying for the preservation of the national floristic diversity, as they have been in constant and rapid regression for several decades, sometimes causing the disappearance of many plant species. In such environments are often spread many interesting ferns like Osmunda regalis L., Marsilea quadrifolia L., Pteris cretica L., P. vittata L. and Woodwardia radicans (L.) Sm., the latter dating back to the Tertiary age for this region. Additionally, more interesting and sometimes unusual species are found in these places, such as the elegant Sagittaria sagittifolia L. or some remarkable insectivorous species of the genera Drosera, Utricularia and Pinguicula.
ASSESSING EXTINCTION RISK: TWO DIFFERENT METHODOLOGIES The first methodology to assess the extinction risk we will describe is the one developed by the IUCN (International Union for the Conservation of Nature and Natural Resources), an international organization founded in 1948 in France, which aims to preserve nature and promote the sustainable development of human societies. As for conservation, one of the main initiatives of IUCN has been the creation of Red Lists of plants and animals since 1963, in order to highlight taxa threatened with extinction in need of conservation actions. IUCN protocol has been gradually improved by updating the threatened categories, drafting more objective criteria and defining numerical thresholds for certain parameters to evaluate quantitatively, like the number of individuals, the distribution area and the area occupied [28] [29] [30]. Thanks to this continuous evolution and implementation of both qualitative and quantitative assessments, the IUCN system has become the World's most used in science and conservation. The latest and currently in use version of IUCN criteria and categories dates back to 2001 and is accompanied by integrations and guidelines containing detailed methodological guidance on how the assessment of a taxon should be done [17] [31] [32]. In 2003 IUCN proposed and published guidelines for the application of criteria, which were originally designed for global Red Lists, at the regional level, such as nations or other administrative regions [33] [34] [35] [36].
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We will now summarize and highlight the main features of the latest available version of the protocol IUCN and suggested guidelines for its application. The methodology poses great attention on terms that assume a different meaning than in biology, as it can cause errors in the process of risk assessment [37]. For this terminology and for the sake of completeness, please refer to the original documents that can be easily obtained on-line. The ultimate goal of IUCN protocol is enlisting taxa to one of ten categories that reflect the level of risk of extinction in the wild according to the degree of knowledge for listing.
Figure 2. IUCN Categories and Criteria (from http://www.iucnssg.org/index.php/ categories-andcriteria).
In fact, species may be not evaluated or, if evaluated, can be data deficient to be enlisted in any category. If there is enough data for a correct assessment, the species can be categorized as Least Concern [LC] or Near Threatened [NT] if it has no risk or a minor risk of extinction. Otherwise, if a species is threatened it can be Vulnerable [VU], Endangered [EN] or Critically Endangered [CR]. Species which have already undergone extinction are enlisted as Extinct in the Wild [EW] or Extinct [EX] (Figure 1). The protocol consists of five quantitative criteria that can be used to evaluate if a taxon is at risk or not, and when at risk, what risk category can be assigned among "CR", "EN" or "VU". The criteria are based on biological indicators related to the risk of extinction of the populations, like processes such as rapid decline or the presence of small and fragmented populations. In summary, the five criteria cover the following aspects: A. B. C. D. E.
Declining populations (past, present and/or projected for the future); Geographical range width reduction, fragmentation and decline or fluctuations; Small populations, fragmented, subject to decline or fluctuations; Very small or very restricted populations; Quantitative analysis, e.g. Population Viability Analysis [38] or other models.
Most of the criteria also include options that are used to assign a taxon to a specific risk category. For instance, a taxon described as "VU C2a (ii)" has been assigned to the Vulnerable category because their population is less than 10,000 mature individuals (criterion C), subject to a continuous decline and all mature individuals are concentrated in one single
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subpopulation (option (ii) of sub-criterion C2). The details of these options can be retrieved in the original documents as well. The definition of these criteria is derived from the analysis of the most significant risk factors for the greatest number of organisms. Similarly, the quantitative thresholds set for different categories of risk are considered as appropriate and consistent with the current knowledge on conservation of wildlife. Moreover, even though a taxon should be evaluated with all the criteria that can be applied on available data, to assign it to any risk category it is sufficient that it meets the requirements of only one of the criteria [31]. Whatever the case may be, when it is possible it is be best to list all of the criteria applied in the evaluation procedure of the risk of extinction; for example: [CR]: A2cd; B1+2de; C2a (i). If different criteria provide dissimilar responses, according to the precautionary principle it should be given priority to those that lead to the highest risk category. It should be noted that, compared to criterion E, criteria A, B, C and D allow a more frequent inclusion of taxa in risk categories. In actual fact, these criteria have been purposely formulated to be more inclusive, because they can be based even on partial or incomplete information; they have larger margins of error than rigorous quantitative analysis. Of course, the more detailed the data are the more precise and definite the assessment will be. In conclusion, it should be pointed out that the quantitative nature of the criteria does not imply a total lack of flexibility in the protocol. Actually, the use of inference procedures and projections based on valid and documented assumptions is possible as it provides the opportunity to evaluate even taxa on which little information exists. We will introduce a second methodology to assess extinction risk developed in recent years in Israel for the local flora. This flora comprises some 800 genera and 3,000 species in 121 families, with a rate of endemism of 7%. Although not as rich as other Mediterranean floras, due to the great extent of deserts, Israel is particularly important as it is a crossway of different phytogeography regions [39]. The conservation methodology known as Red Numbers is implemented primarily by the Israel Plant Information Center (Rotem) of the Hebrew University of Jerusalem and the main output of this research has been publishing the Red Data Book, a two volumes release containing detailed information on endangered flora of Israel and Palestinian Authority and how this red list has been created [39][40]. The basis for the conception of this methodology resided in an inadequate and not up-todate protection and knowledge of plants in Israel. As a matter of fact, plants are still now protected in Israel by the law of “Protected Natural Values” dated 1964. An annexed list recorded 268 species in need of protection throughout Israel [41]. This list has been frequently updated over the years, last time in 1992 and it is now in need of revision [42]. After a checklist of Israel flora in 1999, it was pointed out that 67 taxa were extinct and that they were not attractive plants but species occurring in very few locations and/or endangered habitats [43]. Thus, Shmida et al. criticized the list annexed to the law because its main goal was to protect only plants with large, nice-looking flowers while totally ignoring other taxa that are now extinct in Israel. Moreover, the authors state «the old list includes many ambiguities. It includes plant species that are highly abundant, whole genera without specifying the species, plants that are not natives in Israel and general groups such as "all the trees in the Negev". On top of it, picking of flowers had decreased throughout the seventies, thus the attractiveness is no longer essential for determining vulnerability. Since the species were not selected on the basis of any quantitative data, it is highly insufficient for conservation» [42].
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In the outset of this methodology, an entrance threshold for a plant to enter the list is rarity, defined by the area of species occupancy, in order to provide the highest protection to the rarest species [42]. To assess rarity, sites of occurrence of the species must be counted in relation to the overall extent of the territory. A site is defined as 1 km2 grid cell on the map [44], and the number of sites is considered as a good indicator of a species‟ rarity and the dimension of its populations [45]. The next point that gave birth to the Red Number system in Israel is how to set priorities of species conservation once it has been established if they are endangered. Sapir et al. suggested that «while considerable effort has been invested in conservation programs around the world, the conflict of how to distribute limited resources is still not completely solved. The main challenge is to identify the species and the ecosystems that are a high priority for conservation. The main problem is to define the importance of a certain species for investment of conservation efforts in a situation of limited resources» [18]. A rather complex prioritization system exists on the basis of data at population, taxon and ecological community levels, genetic structure and population ecology [46]. This amount of data is not always available, especially when dealing with many species, thus Red Numbers have been conceived as a fast, pragmatic method to determine conservation priorities at a regional and/or national scale, flexible and adjustable to any other region rather than Israel and not meant to be as alternative or opposition to the IUCN criteria, but complementary in the regional scale [18]. This system is regarded as advantageous by its authors because «calculating red numbers demands minimum data gathering. It can be calculated even if only preliminary data are available. For example, if only herbarium records exist for a certain country, red numbers will be calculated based on existing data, and then a successive approximation will be conducted during or after the field survey» [18]. The Red Number is an additive index, calculated by summarizing the values of six parameters; a parameter that does not contribute any value to the checked species is scored as zero. The first parameter (rarity) is set as a threshold: plants that do not meet the minimum score for this parameter are de facto considered not in danger and the Red Number is not calculated. The advantage of the linear summary is the practical use and the possibility to compare red lists between different regions [18]. Entry threshold of a species to the calculation is the rarity parameter and then the other parameters are added. The highest possible score for a species can be up to 19, when all the parameters get the highest score, but this value is never reached. Besides, to be closer to the decimal system, the output Red Number is subsequently divided by two resulting in a Red Index, e.g. a score of RN 16 becomes RI 8 [40]. The results of this method led to enlist in the Israel Red List all taxa with a Red Number equal or above six, namely 414 taxa corresponding to almost 14% of Israeli flora [40]. The parameters and their possible scores to calculate the Red Number are as follows: I. Rarity [0-6]: this parameter is calculated in relation to the overall area of the region [for Israel and the Palestinian Authority it is circa 33,000 km2] and the number of sites a species occupies. The percentages of territory, the corresponding sites and the resulting score for the parameter are as follows: a. > 0.5% [>110 sites] 0 b. 0.5%-0.1% [110-22 sites] 1
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c. 0.1%-0.05% [21-11 sites] 2 d. 0.05%-0.01% [10-4 sites] 3 e. 0.01% [3 sites] 4 f. 0.005% [2 sites] 5 g. Single site 6 II. Habitat Vulnerability [0-4]: this parameter is calculated on the decrease of sites for a species in Israel since 1964, year of “Protected Natural Values” law. The amounts of decrease for a resulting score for the parameter are as follows: a. No known change since 1964 0 b. No. of sites decreased by 1-30% 1 c. No. of sites decreased by 31-50% 2 d. No. of sites decreased by 51-80% 3 e. No. of sites decreased by >81% 4 III. Attractivity [0-4]: this parameter is closely linked to the main principle of the 1964 law as it suggests that plants with showy and large flowers are more at risk than those with small flowers. Moreover, human exploitation of the plants, such as food, medicine or handicraft, is taken into account as high risk similar to very large flowers. Dimensions of the flowers for a resulting score for the parameter are as follows: a. Flowers not showy 0 b. Flowers showy [< 2 cm diameter] 1 c. Flowers showy [2-4 cm diameter] 2 d. Flowers showy [4-6 cm diameter] 3 e. Flowers showy [> 6 cm diameter] and/or human use 4 IV. Endemism [0-4]: this biogeographical parameter considers the endemism to the region as an important factor for the assessment of the vulnerability of the species. The extent of the area of endemism and its corresponding score for the parameter are as follows: a. Not endemic 0 b. Endemic to the Levant (Israel to Southern Turkey) 1 c. Endemic subspecies 2 d. Endemic mostly to Israel with few other locations outside Israel 3 e. Endemic to Israel and/or the Palestinian Authority 4 V. Peripheriality [0-1]: a species is considered to be peripheral where it occurs on the margins of its natural range. Peripheral populations have a great importance in terms of conservation [47]. Of course, a species can only be or not be peripheral in a region, thus there are only two possible scores for this parameter. VI. Disjunctivness [0-1]: if a species has a very scattered and severely separated distribution in different provinces of the region, then it is disjunctive, otherwise it is not. Scattered and fragmented populations are very important when considering the conservation status of a taxon [10]. For the aims of this chapter, the Red Number system has to be adapted to the region we will analyze, namely the administrative region of Campania in Italy (Figure 3). In particular, parameter I, II and IV should be slightly modified while the others are perfectly adaptable without modifications. The modifications to the method are as follows:
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I. Rarity [0-6]: this parameter is calculated in relation to the overall area of the region; Campania has a surface of 13,590 km2, thus the corresponding sites and the resulting score for the parameter are as follows: a. >0.5% [≥ 68 sites] 0 b. 0.5%-0.1% [67-14 sites] 1 c. 0.1%-0.05% [13-7 sites] 2 d. 0.05%-0.01% [6-4 sites] 3 e. 0.01% [3 sites] 4 f. 0.005% [2 sites] 5 g. Single site 6 II. Habitat Vulnerability [0-4]: in Campania, as in the whole Italian territory, there is no immediate equivalent to the 1964 law of Israel or a local comprehensive flora. Thus, in order to use a comparable parameter as habitat vulnerability and declining rate, we will refer to the known historical locations for our species in the local floras available for Campania, namely flora of Monti Picentini [48], floras of Monti Lattari, also known as Penisola Sorrentino-Amalfitana [49] [50] [51], floras of the island of Capri [52] [53] and flora of the area around the city of Naples [54]. In addition, the two main floras of Italy have been considered as historical distribution data as well [55] [56]. Other historical data, when available and geographically identifiable, have been collected from herbarium collections of Michele Tenore (1780-1861) and Giovanni Gussone (1787-1866) at Herbarium Neapolitanum (NAP) in the Botanical Garden of Naples. Accordingly, the amounts of decrease for a resulting score for the parameter are as follows: a. No known change from available floras 0 b. No. of sites decreased by 1-30% 1 c. No. of sites decreased by 31-50% 2 d. No. of sites decreased by 51-80% 3 e. No. of sites decreased by >81% 4 III. Endemism [0-4]: this biogeographical parameter considers the endemism to the region as an important factor for the assessment of the vulnerability of the species. If we consider Southern Italy as the union of Abruzzo, Molise, Campania, Calabria, Apulia, Basilicata, Sardinia and Sicily regions, then the extent of the area of endemism and its corresponding score for the parameter are as follows: a. Not endemic 0 b. Endemic to Italy 1 c. Endemic subspecies 2 d. Endemic to Southern Italy 3 e. Endemic to Campania 4
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Figure 3. Campania (Southern Italy) with its main geomorphologic units.
MODEL OF SIX DIFFERENT SPECIES FROM SOUTHERN ITALY As we mentioned in the previous sections, we will use as a case study six different species occurring in Campania, an administrative region of Southern Italy. It is worth highlighting once again that the resulting IUCN risk category and Red Number/Red Index are meant to be at a regional level. Classification and taxonomy are according to the Angiosperm Phylogeny Group [57]. The species are:
Chamaerops humilis L. (Arecaceae); Globularia neapolitana O. Schwarz (Plantaginaceae); Medicago arborea L. (Fabaceae); Parnassia palustris L. (Celastraceae); Pinguicula hirtiflora Ten. (Lentibulariaceae); Simethis planifolia [L.] Gren. et Godr. (Xanthorrhoeaceae).
We will now present a report for each species giving the main descriptive and ecological features, the distribution range, the available historical data for Campania and the data gathered during field surveys during 2009-2011 (Figure 4) [58]. In conclusion, we will present the resulting IUCN risk category and the computed Red Number and Red Index with a few remarks, where necessary, on the methods and choices for each case. Conclusive comparisons and remarks on the two methodologies will be in the final section of this chapter with references to the practical examples of these species when needed.
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Figure 4. Overall historical and verified distribution of case study species in Campania (Southern Italy). Green dots: Chamaerops humilis, Black dots: Globularia neapolitana, Yellow dots: Medicago arborea, Red dots: Parnassia palustris, Blue dots: Pinguicola hirtiflora, White dot: Simethis planifolia.
Chamaerops Humilis L. (Arecaceae) Biology and Ecology: The European fan Palm or Mediterranean dwarf Palm is a shrublike palm native to the Mediterranean. It is a nano or meso-phanerophyte, usually 1-2 m tall, with very small yellowish flowers pollinated mainly by Coleoptera. In Italy, this palm lives in typical Mediterranean maquis or macchia ecosystems, favoring steep cliffs facing the sea, between 0 to 600 m of altitude [56]. Distribution range: The plant is native to the Western part of the Mediterranean, from Portugal up to Malta [59]. In Italy, the plant is widely distributed in Sardinia and Sicily, but it is also present along the coast of the Tyrrhenian Sea from Calabria to Tuscany, yet with scattered and less dense populations [56]. Available historical data for Campania: The plant was historically present only in the island of Capri and in the neighboring Penisola Sorrentina. As for Capri (Figure 4), the plant was well known in the 18th century in the most remote and inaccessible cliffs of the island while further studies proved it diminished along the coast in the 19th and 20th century [52][53]. A very similar phenomenon happened to the coast of Penisola Sorrentina (Figure 5), where the plant in the past was reported to reach the city of Salerno [49]. Survey data: Two populations have been found in Campania, one in the Northern side of the Island of Capri and the other along the Southern coast of the Penisola Sorrentina between the small towns of Maiori and Erchie, covering an area of four sites, with an overall estimated population of about 220 mature individuals.
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IUCN risk category: EN (endangered) according to criterion C2a(i), meaning the population size is estimated to number fewer than 2,500 mature individuals and a continuing decline, observed, projected, or inferred, in the numbers of mature individuals with a population structure of no subpopulation estimated to contain more than 250 mature individuals [17]. Red Number and Red Index: I (3) + II (3) + III (0) + IV (0) + V (0) + VI (1) = RN (7) RI (3.5).
Figure 5. Habitat of Chamaerops humilis on the cliffs of Capri.
Figure 6. Details of Chamaerops humilis along the coast of Penisola Sorrentina.
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Globularia Neapolitana O. Schwarz (Plantaginaceae) Biology and Ecology: G. neapolitana is a small chamaephyte endemic to Campania. The plant has rather small, azure-tinged flowers and grows as a tiny shrub climbing onto steep cliffs and crevices of limestone walls between 400 and 1400 m [56]. It is very similar in habitus and habitat to G. meridionalis (Podp.) O. Schwarz and G. cordifolia L., with taxonomic distinguishing features still in need of investigation [58]. Distribution range: The plant is endemic to Campania [56].
Figure 7. Habitat of Globularia neapolitana.
Figure 8. Detail of Globularia neapolitana flower.
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Available historical data for Campania: On the Island of Capri, the plant is reported to grow in the highest part of the island, around 400 m of altitude, a growth height fairly uncommon among other Globularia species [52] [53]. In the Penisola Sorrentina, the plant grows in the highest part of the Monti Lattari, on the vertical cliffs around 1400 m [50]. For the mountain range of Monti Picentini, the plant is reported to grow on two different peaks between 1400 and 1800 m [48]. Survey data: Only one population has been confirmed to be present in Campania during field surveys, namely in Penisola Sorrentina, around Monte S. Angelo a Tre Pizzi, between 1200 and 1450 m, covering 2 sites (Figures 6 and 7). Except for the one in the Monti Picentini which has not been investigated, the other historical populations have not been found again. The overall population can be estimated in more than 10,000 mature individuals [58]. IUCN risk category: CR (critically endangered) according to criterion B2ab(iv), meaning geographic range in the form of area of occupancy is estimated to be less than 10 km2, known to exist at only a single location with a continuing decline, observed, inferred or projected in the number of locations or subpopulations [17]. Red Number and Red Index: I (5) + II (3) + III (1) + IV (4) + V (0) + VI (0) = RN (13) RI (6.5).
Medicago Arborea L. (Fabaceae) Biology and Ecology: This species, commonly known as moon trefoil, shrub medick or tree medick, is a shrub-like phanerophyte in the large genus Medicago characterized by its silvery leaves, bright yellow flowers and disc-shaped legumes. This plant grows in typical Mediterranean maquis or macchia ecosystems, favoring cliffs facing the sea, between 0 to 300 m of altitude, on both tuff and limestone [56].
Figure 9. Medicago arborea growing close to a suburban area near Naples.
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Figure 10. Medicago arborea growing on the Island of Capri.
Distribution range: The plant grows in the Southern part of the Mediterranean, from Greece to Spain [59]. In Italy, the plant has several, often scattered populations between the Central and the Southern part of the peninsula, including the two major islands of Sicily and Sardinia [56]. Available historical data for Campania: The plant is reported to grow along the coast of the Western side of the city of Naples, on tuff cliffs [54]. Historical herbarium specimens of both M. Tenore and G. Gussone document the species in Pozzuoli and Ischia, north of Naples. For the Island of Capri, the plant is reported from several localities (Figure 8) [52] [53]. Survey data: A very large and rather continuous population still grows in the area of Naples known as Posillipo (Figure 9), while few individuals have been found on ancient Roman ruins not far away from Pozzuoli. In Capri, the species grows in the Northern part of the island between the villages of Capri and Anacapri and in a few spots on the Southern side of the island. The overall population counts some 2,000 mature individuals in 12 sites [58]. IUCN risk category: VU (vulnerable) according to criterion D2, meaning the population has a very restricted area of occupancy (typically less than 20 km2) or a small number of locations (typically five or fewer), such that it is prone to the effects of human activities or stochastic events within a very short time period in an uncertain future, and is thus capable of becoming Critically Endangered or even Extinct in a very short time period [17]. Red Number and Red Index: I (2) + II (3) + III (1) + IV (0) + V (0) + VI (1) = RN (7) RI (3.5).
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Parnassia Palustris L. (Celastraceae) Biology and Ecology: This plant, known as Marsh Grass-of-Parnassus, Northern Grassof-Parnassus or Bog-star, is a hemicryptophyte characterized by a basal rosette of leaves and a single leaf on the middle of the flowering stem, holding a nice-looking white flower, between 2-3 cm wide, with a distinguishing set of five nectaries. The typical habitats where this species is found are all damp and marshy ecosystems, bogs or swamps, usually between 200 and 1900 m [56]. Distribution range: This species is a boreal element of European flora, widely distributed in Northern and Central Europe [59]. In Italy, the plant is quite diffused along the Alps, the Northern and Central Apennines but is rarer and rarer due south [56]. Available historical data for Campania: Campania has historically always been the southernmost outpost of this species in Italy. A historical decline for the plant in this region is already documented in local floras. In Penisola Sorrentina it was present at the beginning of 20th century [49], yet it is reported to have disappeared already in the 80s [50]. On the mountains of Monti Picentini the plant is reported as very rare, growing into a few gorges [48]. Survey data: Former locations have been confirmed as no longer hosting this species. Even in the sites known as still holding a small population of P. palustris, the plant has not been found again, except for very few individuals in a deep gorge of Monti Picentini known as Valle della Caccia (Figures 10 and 11), covering just one single site [58].
Figure 11. Habitat of Parnassia palustris inside the Valle della Caccia.
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Figure 12. One of the few ruined specimens of Parnassia palustris found in the Valle della Caccia.
IUCN risk category: CR (critically endangered) according to criterion D, meaning the population size is estimated to number fewer than 50 mature individuals [17]. Red Number and Red Index: I (6) + II (3) + III (2) + IV (0) + V (1) + VI (0) = RN (12) RI (6).
Pinguicula Hirtiflora Ten. (Lentibulariaceae) Biology and Ecology: This plant is a very small insectivorous hemycriptophyte with sticky leaves and very beautiful violet, blue or whitish flowers. As with many other Pinguicula species, the plant is mainly present on dipping carbonate or serpentine rocks, usually in mountain ecosystems but also not very far away from the sea whenever the habitats are available [56] [60].
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Figure 13. Valle delle Ferriere, a typical Pinguicola hirtiflora habitat in the Penisola Sorrentina.
Figure 14. Pinguicola hirtiflora blossoming in the peaks of Penisola Sorrentina.
Distribution range: The plant is distributed along Southeastern mountains of the Mediterranean, mainly in Turkey, Greece, and Albania with the most western outpost in Italy [59]. In the latter country, the plant is present mainly in Campania and in a single site in Calabria [55] [56].
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Innangi Michele, Izzo Antonio and De Castro Olga
Available historical data for Campania: This species was described by Michele Tenore on the mountains of Penisola Sorrentina, namely on Monte S. Angelo a Tre Pizzi [61]. It was later found in several gorges and valleys of this peninsula during many surveys throughout the years [49] [50] [51] [60]. In the mountains of Monti Picentini the plant has been discovered more recently and it is present only in two locations [48]. Survey data: Many locations for Penisola Sorrentina have been confirmed, whereas a few have been not been found again and others discovered as new (Figures 12 and 13). On the Monti Picentini, the known stands have been verified, even though in one site the population presents interesting phenotypic features with probable taxonomic relevance [62]. The overall population for this taxon in Campania is some 12,000 specimens, although there are great differences in the extension of the subpopulations within the 8 sites [58]. IUCN risk category: VU [vulnerable] according to criterion B2ab(ii+iii+v), meaning the geographic range in the form of area of occupancy is estimated to be less than 2000 km2, and estimates indicating a population severely fragmented or known to exist at no more than 10 locations with a continuing decline, observed, inferred or projected in area of occupancy, quality of habitats and number of mature individuals [17]. Red Number and Red Index: I (2) + II (1) + III (1) + IV (0) + V (1) + VI (1) = RN (6) RI (3).
Simethis Planifolia [L.] Gren. Et Godr. (Xanthorrhoeaceae) Biology and Ecology: The Kerry Lily is a geophyte apparently looking as some Allium species, distinguished by small white flowers with hairy stamens. The plant usually grows in the Mediterranean in macchia-garrigue ecosystems or, in the rest of Europe, in coastal, dry and acidophilus grasslands, between 0 and 500 m of altitude [56]. Distribution range: This species has a Mediterranean-Atlantic distribution, with a rather peculiar and scattered range throughout Southwestern Europe, reaching Ireland in the North [59]. In Italy, the plant grows in some Tyrrhenian regions, with the most Southeast outpost in Campania [55] [56]. Available historical data for Campania: The plant is long known for a single site in Penisola Sorrentina close to the coast on the tip of the peninsula (Figures 14 and 15) [50]. Survey data: The known single site for this species has been verified with a very large population of some 100,000 individuals in a rather restricted area along the coast [58]. IUCN risk category: VU (vulnerable) according to criterion D2, meaning the population has a very restricted area of occupancy (typically less than 20 km2) or a small number of locations (typically five or fewer), such that it is prone to the effects of human activities or stochastic events within a very short time period in an uncertain future, and is thus capable of becoming Critically Endangered or even Extinct in a very short time period [17]. Red Number and Red Index: I (6) + II (0) + III (1) + IV (0) + V (1) + VI (0) = RN (8) RI (4).
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Figure 15. Habitat of Simethis planifolia.
Figure 16. Detail of Simethis planifolia flower.
CONCLUSION Although the two methods for risk assignment are not conceived as alternative or antagonist, the results, on the basis of the same data for the same species, are largely overlapping and comparable, with some distinctions. As a matter of fact, the species considered as Critically Endangered according to IUCN have the highest resulting Red Number/Red Index, like those considered as Vulnerable in relation to IUCN have the lowest Red Number/Red Index as well (Figures 17 and 18).
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Innangi Michele, Izzo Antonio and De Castro Olga
Figure 17. IUCN criteria and risk category for case study species from Campania (Southern Italy).
Figure 18. Red Number and Red Index for case study species from Campania (Southern Italy)1.
Less correspondence between IUCN and Red Number is found only for C. humilis, and this is probably due to the extent of the populations which cover more than one site per population but with very few and scattered individuals. Moreover, G. neapolitana results as more endangered than P. palustris in terms of Red Number whereas, as a matter of fact, P. palustris can be considered as almost extinct for Campania. This happens mostly because G. neapolitana is endemic to Campania while P. palustris is, elsewhere, widely distributed. Thus, these two methods offer comparable results but, as noticed by Sapir et al., the Red Number is more prone to local investigation and can used as an entry instrument to decide which plants need urgent investigation, perhaps using IUCN method [18]. Nevertheless, regional and local situations always require detailed field investigation in order to obtain information that can more closely represent the reality, regardless if IUCN or Red Number is applied. Admittedly, even the IUCN method sometimes offers results that do not exactly match the real status of a plant in a local territory. Once again, if we consider G. neapolitana and P. palustris, they both result as CR according to IUCN, but this is far from be true or at least “comparatively critical” in Campania. It is important, in this case, to consider growth form and habitats for G. neapolitana that usually lives undisturbed on the many inaccessible mountain peaks of Campania where it is rarely seen. Indeed, notwithstanding several guidelines provided by the IUCN according to different cases, criteria and options do not consider the differences between various growth forms. Moreover, other important biological data for plants, e.g. pollination or seed dispersal, are ignored in the criteria. This and other small flaws in the IUCN method are all due to the fact that the criteria are meant for all living things on Earth, no matter if they are animal,
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fungi, algae or plants. IUCN offers a powerful and comprehensive tool to assess the risk of extinction of biota on Earth, but maybe the criteria, at least at option level, should be different if you are dealing with plants or animals. Notwithstanding some minor issues with IUCN protocol and guidelines, this method is undoubtedly one of the most important tools for conservation strategies [63]. The method is flexible and dynamic, and it is developed to be as much objective and quantitative as possible without denying the experts making subjective evaluations. One expert based addendum is the list of threats which aids understanding which are the more urgent conservation initiatives to enterprise in order to salvage a species, although sometimes this subjectivity can be misleading or erroneous [64]. Probably these threats, when correctly and thoroughly assessed, should be somehow quantified as figures to complete and perfect the risk assessment. In conclusion, the IUCN method is complete and precise with few weaknesses when few data are available and almost flawless when a thorough data collection is achieved. Admittedly, when the assessment is meant to be done on the basis of data gathered on the field it is not a fast way to evaluate whether a species is threatened or not, but it is imperative to implement and improve its application for conservation at regional, national and global scales. A faster approach is the one provided by calculating Red Numbers. In actual fact, even when few data are available, Red Numbers can be computed to decide which species need more urgent investigation in order to choose which ones deserve more pressing conservation actions, especially when few economic resources are accessible. This is particularly true at regional scale, where often local flora are not always up to date and not abundant in information, thus calculating Red Numbers can be an useful instrument to quickly evaluate a great number of species and select which ones require investments in terms of human and economic action. In comparison with IUCN method, which has only three degrees of threat, namely Vulnerable, Endangered and Critically Endangered, Red Numbers, being figures, have a wider range of risk evaluation. For instance, in our case studies, M. arborea, P. hirtiflora and S. planifolia are all Vulnerable according to IUCN while we can see different figures as Red Numbers, somehow “quantifying” the differences in risk between the species. On the other hand, being this a fast method that can offer results even on the basis of few data, not always the outcome is closely responding to reality, although no human investigation can actually understand the full complexity of natural phenomena. The parameters proposed to calculate Red Numbers for Israeli flora are imagined as to be adaptable to other local floras as well. But some of this parameters, mostly the number II (Habitat Vulnerability) is not conceived as to be immediately corresponding to any other flora because it is based on a local law and on the assumption that data on the decline of a species are available. Attractivity, which was greatly evaluated in the Israeli 1964 law, has been preserved as an important parameter in Red Numbers. It cannot be denied that large, colorful flowers like lilies or big orchids can be more threatened than anonymous-looking grasses or nettles, but most of the times threats to the habitats, population size or global range are much more serious matters for a plant‟s survival than the size of its flowers. Thus, even if all parameters of the Red Numbers have to be evaluated, they should have a coefficient according to the relevance of the parameter in terms of conservation. For instance, Habitat Vulnerability, Attractivity and Endemism all can score up to 4 but, according to us,
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Innangi Michele, Izzo Antonio and De Castro Olga
given two species with the same rarity, a plant which is suffering a loss in its habitat of more than 81% should have more urgent need of conservation than one with flowers larger than 6 cm with no known decline. Additionally, endemism is considered by IUCN only in terms of geographic range because this method is more centered on evaluating risk of extinction whereas it is very important for Red Number computing as this method is also more prone to evaluate priorities for investigation and/or conservation. Besides, this parameter could be, in our opinion, more in a wider biogeographical sense, giving 0 score to cosmopolitan species and higher figures the more the biogeographical range goes toward endemism. Again, Endemism is a relative thing, thus for instance, in the Mediterranean, Steno-Mediterranean species should have a higher score than Euro-Siberian species. In conclusion, assessing risk of extinction for plant species and prioritize conservation initiatives it is not easy and it is far away from being a definitive process. Given field survey data for a local flora of an administrative region of Italy, quantitative methods such as IUCN protocol and Red Numbers give reliable assessments of extinction risk. Red Numbers, with adequate adaptations, can be a very useful instrument when evaluating large numbers of species, while IUCN, even though requiring much more data and evaluations for a reliable and thorough evaluation, give in depth analysis of actual risk and threats for a species. Both methods can be implemented and perfected especially at regional or local level. But then again, the real necessity to safeguard biodiversity is first of all improving our knowledge of biology, ecology and taxonomy of species, creating a real sustainable development of natural and human-shaped ecosystems and protecting habitats, especially the frailer ones which host the vast majority of rare and endemics species.
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[35] V. Keller and K. Bollman, Conserv. Biol. 18, 1636 (2004). [36] E.J.Milner-Gulland, E. Kreuzberg-Mukhina, B. Grebot, S. Ling, E. Bykova, I. Abdusalamov, A. Bekenov, U. Gärdenfors, C. Hilton-Taylor, V. Salnikov and L. Stogova, Biodiv. Conserv. 15, 1873 (2006). [37] IUCN, Analysis of the Application of IUCN Red List Criteria at a National Level, Report from the National Red List Advisory Group Workshop.Villa Majagual, Venezuela(2005). [38] S.R. Beissinger and D.R. McCullough, Population Viability Analysis, University of Chicago Press, Chicago (2002). [39] A. Shmida and G. Pollak, Red Data Book: Endangered Plants of Israel, vol. I, Israel Nature and National Parks, Jerusalem (2007). [40] A. Shmida and G. Pollak, Red Data Book: Endangered Plants of Israel, vol. II, Israel Nature and National Parks, Jerusalem (2011). [41] U. Paz, The Nature Reserves of Israel, Nature Reserves Authorities, Massada, Tel Aviv (1981). [42] A. Shmida, O. Fragman, R. Nathan, Z. Shamir and Y. Sapir, Ecologia Mediterranea 28, 55 (2002). [43] O. Fragman, R. Nathan and A. Shmida, Ecol. Environ. 5, 207 (1999). [44] K.J. Gaston, Rarity, Chapman and Hall, London (1994). [45] K.J Gaston, T.M. Blackburn, J.J. Greenwood, R.D. Gregory, R.M. Quinn and J.H.Lawton, J. Ap. Ecol. 37, 39 (2000). [46] D.J. Coates and K.A. Atkins, Biol. Conserv. 97, 251 (2001). [47] D. García and R. Zamora, J. Veg. Sci. 14, 921 (2003). [48] B. Moraldo, V. La Valva, M. Ricciardi and G. Caputo, Delpinoa 27-28, 59 (1986). [49] M. Guadagno, Bollettino dell'Orto Botanico della Regia Università di Napoli V, 133 (1918). [50] G. Caputo, V. La Valva, R. Nazzaro and M. Ricciardi, Delpinoa 31-32, 3 (1989-1990). [51] G. Caneva and L. Cancellieri, Il Paesaggio Vegetale della Costa d'Amalfi, Gangemi, Roma (2007). [52] M. Guadagno, Flora Capraearum Nova. Tipografia Valbonesi, Forlì (1932). [53] M. Ricciardi, Annali di Botanica LIV, 7 (1996). [54] A. De Natale and V. La Valva, Webbia 54, 331 (2000). [55] A. Fiori, Flora Analitica d'Italia, Edagricole, Bologna (1923-1929). [56] S. Pignatti, Flora d'Italia, vol. II, Edagricole, Bologna (1982). [57] Angiosperm Phylogeny Group, Bot. J. Linn. Soc. 161, 105 (2009). [58] M. Innangi, A. Izzo and V. La Valva, Delpinoa 49(2007), 77 (2011). [59] T.G. Tutin, N.A. Burges, A.O. Chater, J.R. Edmondson, V.H. Heywood, D.M. Moore, D.H. Valentine, S.M. Walters and D. A. Webb, Flora Europea, vol. III, Cambridge University Press, Cambridge (1972). [60] M. Innangi, A. Izzo and V. La Valva, La biodiversità vegetale in Italia: aggiornamenti sui gruppi critici della Flora Vascolare Italiana, Società Botanica Italiana, Firenze (2010).
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[61] M. Tenore, Flora Napolitana 1 (Ad Florae Neapolitanae Prodromus) VI, Napoli (1811). [62] M. Innangi, O. De Castro, A. Izzo, Geometric morphometrics of Pinguicula hirtiflora Ten. (Lentibulariaceae) in Southern Italy: taxonomic and biogeographical evidence. Manuscript in preparation. [63] A.S. Rodrigues, J.D. Pilgrim, J.F. Lamoreux and M. Hoffmann, Trends Ecol. Evol. 21, 71 (2006). [64] M.W. Hayward, Cons. Biol. 23, 1568 (2009).
In: Endangered Species Editor: Manuel Esteban Lucas-Borja
ISBN: 978-1-62257-532-9 © 2012 Nova Science Publishers, Inc.
Chapter 3
EFFECT OF CLIMATE CHANGE AND DEFORESTATION IN A SELECTION OF VERTEBRATE SPECIES AND OPUNTIA IN MÉXICO Patricia Illoldi-Rangel, Víctor Sánchez-Cordero and Miguel Linaje Departamento de Zoología, Instituto de Biología, Universidad Nacional Autónoma de México, México
ABSTRACT México has great biological diversity, but deforestation and climate change threatens its conservation. The loss of biodiversity has been inferred, quantifying deforestation of the principal vegetation types, relating that loss with the reduction in natural habitat. Nonetheless, this scope does not evaluate the impact of deforestation at species level. Climate change is causing changes in climatic regimes that are already impacting in different aspects of biodiversity, like alteration in geographic distributional ranges of species. In this chapter, we integrated habitat loss and climate change to evaluate the effect in the geographic distribution range of a selection of vertebrate species and species from the Opuntia genera in México. Ecological niche models were generated and then projected as distributional ranges. The actual distribution for each species was estimated according to the loss of vegetation types to which these species are associated, using a land use and vegetation map for México. Projections on two different climate change scenarios were done, using scenario A2 (severe or “pessimistic”) and B2 (conservative or “non pessimistic”) for 2020, 2050 and 2080 in order to anticipate its effect in the distribution of the species selected.
E-mail:
[email protected];
[email protected].
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Patricia Illoldi-Rangel, Víctor Sánchez-Cordero and Miguel Linaje
INTRODUCTION México is a mega diverse country in terms of terrestrial vertebrates worldwide [1, 2], but a high deforestation rate has put at risk its conservation. The loss of biodiversity has been generally inferred by quantifying deforestation of the principal types of vegetation –i.e. tropical forests, pine forests, cloud forest, etc.- associating the loss of biodiversity to the reduction in such vegetation types. This method has been used at a worldwide scale [3, 4, 5], as well as in México [6, 7, 8, 9]. Nonetheless, this approach does not predict the impact of deforestation at a species by species level, one or the essential components of biodiversity. In other aspects, it is well documented that today, global weather is changing in an unprecedented manner [10] and that this climatic alterations are having effects within biodiversity at different levels. It has been detected in the physiology of some mammals [11], in migration patterns and hibernation of some species of birds and insects [12, 13], and in movements of the distributional limits towards the artic zones or high altitudes of some varieties of butterflies, birds and some plants [14, 15, 16, 17, 18].
Potential and Actual Distribution of Endemic Vertebrate Species and Species from the Opuntia Genera Selected, Using Ecological Niche Models (ENM) Scientific collections harbor primary information to generate knowledge for biological diversity and its geographical distribution. The specimens kept in scientific collections and their associated information (i.e. locality and date) contains basic information for the knowledge of the biological diversity in a certain region. For these reasons, scientific collections and biological inventories are a fundamental part of biodiversity studies [19, 20, 21, 22, 23, 24]. An ideal model would take advantage of the extensive information contained in the scientific collections and would resolve the disadvantages that present other statistical models. A promising area is the use of algorithmic models [25]. The use of these algorithms has been developed to the modeling of species distribution, which gives the option of using information contained in scientific collections with the advantage of a more adequate management of the available data. Bias, geographic and taxonomic within biological inventories, normally exclude areas in which species are potentially present [26, 27, 28, 29]. For this reason, ecological niche modeling represents a good alternative given that they extrapolate from climatic, geologic and vegetation parameters using the known occurrence of the species, with the purpose of identifying the habitat where a species has not been registered but is probable to occur. This approximation, implemented in the computational package MaxEnt [30], has demonstrated the capability of providing accurate predictions, in which each cell gives a reference of accumulative values, represented as a proportion of the probability value of each cell [30]. Ecological niche modeling provides a framework to determine the impact of deforestation and climate change on the distribution of species; hypothesis are provided for the actual and future distributions of the species, considering the potential impact of both factors [22, 23, 31]. The actual distribution hypothesis is based on the assumption that
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conversion (at a “coarse” scale) of the natural habitats to agro-systems and human settlements results in ecological conditions not appropriate for the species [32, 22, 23, 24]. If climatic conditions suffer important changes because of global warming, species populations that inhabit areas with strong alterations could tend to disappear from those sites. In the same manner, if sites that actually are inappropriate for a species but by effect of climate change become adequate, it is possible that populations of that species colonize new areas.
Climatic Projections for Endemic Vertebrate Species and Species from the Genera Opuntia Selected Under Climate Change Scenarios Ecological niche models for selected endemic vertebrate species were projected to actual conditions and climatic conditions for 2020, 2050 and 2080 using climate change scenarios generated by the Intergovernmental Panel on Climate Change (IPCC), SRES A2 y B2 from the Canadian Centre for Climate Modeling Analyst (CGCM2) (http://ipcc-ddc.cru.uea.ac.uk/ sres/cgcm2_download.html).
Selection of Conservation Priority Sites in Actual Distribution Scenarios and Climate Change Scenarios One approximation to the problem of area conservation is priority sites techniques. The goal of prioritizing areas within conservation biology consists in arranging a series of places based on their biodiversity content [33, 34, 35, 36, 37]. Given that it is almost impossible to achieve the complete conservation of all biodiversity at any scale, conservation strategies often focus in choosing some aspects of an ecosystem that will work as “surrogates” in the most efficient possible way, and from there, prioritize areas [34, 36, 38, 39]. These areas must be prioritized before selecting areas in which any conservation action will be taken [34, 36]. One of such algorithms is found in the ResNet program [35]. ResNet implements a hierarchical selection controlled algorithm, based on rarity and complementarity. ResNet uses an iterative process that selects places based on a rarity criterion; in case of conflicts between the areas selected (i.e. they have the same geographical rarity), the algorithm uses a complementarity criterion (i.e. selects the area that contains the greatest quantity of surrogates that have not yet been represented according to the established goal). If there exists a tie between areas after those two criteria have been used, ResNet optionally uses adjacency (adjacent cells are preferred over non adjacent cells) and finally the algorithm selects areas in a random manner [39, 40]. Usually, two types of goals are utilized: (i) a level of representation for the coverage expected of each one of the surrogates (i.e. the mean number or expected number of occurrences of the species used) within a conservation area; and (ii) conjunct such representation with the maximum area that can be conserved [37]. The main objective for the present chapter is to quantify the potential distribution areas for a selected group of terrestrial vertebrate species and species from the genera Opuntia, considering scenarios for deforestation and climate change, to identify the mayor areas of risk for regional extinction and the areas of major importance for conservation.
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METHODS Database and Criteria for the Selection of Species Species of terrestrial vertebrates and species for the genera Opuntia were selected under the following criteria: (1) the species must be in the list of the Norma Oficial Mexicana (Mexican Official Norm; NOM-059-SEMARNAT-2001), under a risk, threatened and/or endangered category; (2) they must be endemic to México, and (3) they must have more than 10 different localities (geographical coordinates). The total number of species that compelled to these criteria was 93 species (see Table 1). Table 1. Terrestrial vertebrate species list and species from the genera Opuntia selected for this chapter MAMMALS Cryptotis magna Cynomys mexicanus Dipodomys gravipes Dipodomys margaritae Dipodomys phillipsii Geomys tropicalis Lepus flavigularis Liomys spectabilis Megasorex gigas Microtus oaxacensis Microtus quasiater Musonycteris harrisoni Myotis vivesi Nelsonia neotomodon Neotoma phenax Peromyscus bullatus Peromyscus zarhynchus Procyon insularis Rheomys mexicanus Romerolagus diazi Sciurus oculatus Sorex macrodon Sorex milleri Spilogale pygmaea Xenomys nelsoni Zygogeomys trichopus OPUNTIAS Opuntia excelsa BIRDS Aimophila sumichrasti Amazona finschi Anas platyrhynchos Ardea herodias Campylopterus excellens Cyanolyca mirabilis Cyanolyca nana Dendrortyx barbatus Dendrortyx macroura
Number of ocurrences 40 28 33 18 168 11 37 17 51 75 110 51 13 26 58 13 30 34 11 42 24 15 10 56 14 18 Number of occurrences 46 Number of occurrences 48 86 63 193 26 21 22 19 43
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Number of occurrences
Doricha eliza Eupherusa poliocerca Euptilotis neoxenus Forpus cyanopygius Geothlypis speciosa Nyctanassa violacea Nyctiphrynus mcleodii Passerina rositae Progne sinaloae Rhynchopsitta terrisi Thalurania ridgwayi Toxostoma guttatum Vireo bairdi Vireo nelsoni Xenospiza baileyi
25 29 58 94 47 151 19 56 14 22 19 24 43 30 23
AMPHIBIANS AND REPTILES
Number of occurrences
Ambystoma amblycephalum Ambystoma granulosum Bolitoglossa platydactyla Bufo cristatus Chiropterotriton chiropterus Chiropterotriton chondrostega Chiropterotriton dimidiatus Chiropterotriton lavae Chiropterotriton magnipes Chiropterotriton multidentatus Duellmanohyla chamulae Duellmanohyla schmidtorum Eleutherodactylus angustidigitorum Eleutherodactylus berkenbuschi Eleutherodactylus decoratus Hyla plicata Lineatriton lineolus Plectrohyla acanthodes Plectrohyla lacertosa Pseudoeurycea belli Pseudoeurycea cochranae Pseudoeurycea gadovi Pseudoeurycea galeanae Pseudoeurycea juarezi Pseudoeurycea leprosa Pseudoeurycea robertsi Pseudoeurycea scandens Pseudoeurycea werleri Rana brownorum Rana dunni Rana montezumae Rana neovolcanica Rana pustulosa Rana sierramadrensis Thorius dubitus Thorius narisovalis Thorius pennatulus Thorius pulmonaris Thorius troglodytes
15 33 169 70 290 68 13 19 16 58 16 24 67 57 76 135 32 22 23 179 20 32 17 19 228 33 36 16 122 19 256 30 847 22 22 30 34 20 25
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Potential and Actual Distribution of Species Using Ecological Niche Modeling We used 21 climatic layers with a 0.01o 0.01o resolution, derived from WorldClim (http://www.worldclim.org/), to model the ecological niche model (ENM) of the 93 species selected. To generate the actual distribution models the natural remnant habitat was considered within the projection of the potential distribution, based on the land use and vegetation map generated by the National Institute of Geography and Statistics (INEGI, Serie III; www.inegi.gob.mx); that is, areas that only contained remnant natural habitat within the potential distribution of the species was considered as the actual distribution of the species.
Species Distribution Under Climate Change Scenarios A2 Y B2 For this part of the analyses, the climatic variables included were: mean annual temperature (°C), daily oscillation of temperature (°C), isothermality (°C), maximum mean temperature of the hottest period (°C), minimum mean temperature of the coldest period (°C), mean temperature of the most rainy quarter (°C), mean temperature of the driest quarter (°C), annual precipitation (mm), precipitation of the rainiest period (mm), precipitation of the driest period (mm), precipitation seasonality, slope and topographic index, all with a pixel size of 0.01° (1 kilometer resolution), and obtained from the United States Geological Service (http://edcdaac.usgs.gov/gtopo30/hydro/), giving a total of 19 climatic variables. All distributions for the selected species were projected into four time periods (actual, 2020, 2050 and 2080). For each species, 7 models were generated: one for the potential/actual distribution; two (A2, B2) for the 2020 climatic scenario; two (A2, B2) for 2050, and two (A2, B2) for 2080 (figure 1).
Figure 1. Regions of high risk for the selected endemic species. The impact of deforestation (D) is more relevant in the Neovolcanic Belt and the Neotropic (Gulf of México Plain), while the main impact of climate change (CC) can be seen in the Altiplano and north of México, as well as in the Gulf of México Plain.
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Selection of Priority Sites for Conservation of the Selected Species under Actual and Climate Change Scenarios All the models that were generated using MaxEnt were exported as a data matrix (in ascii format) to be incorporated to ResNet. ResNet requires a target for each of the species, that is, the minimum number of occurrences that a surrogate must be represented in the selected cells. For the present analysis, the selected target was 100% of the potential distribution for the species selected, given that all of them are endemic to México and included in the Mexican Norm (NOM). Also, the area that could be conserved can be selected, assigning a percentage of the total area used in the study. For the present chapter three analyses were made, selecting 15, 20 y 25% of the total area of México for its conservation.
RESULTS AND DISCUSSION We modeled 93 distributions for endemic terrestrial vertebrate species and species of the genera Opuntia, all projected as potential and actual distributions. Percentages of natural remnant habitat were calculated for each one within the potential distribution.
Impact of Deforestation on the Distribution of Species The evaluation of the impact of the deforestation on the distribution of the selected endemic species was based on the quantification of the loss of area of distribution of the potential distribution models compared to the actual distribution models. This assumption is in accordance to the precautory principle in the sense that prioritization for conservation is made only for natural remnant habitats within their distribution area. Therefore, we consider this results and tendencies as viable and of high predictive power. The results obtained indicate two tendencies on the impact of deforestation on the distribution of the selected endemic species. There was no observed correlation between the loss of natural remnant habitat and the assignation of a risk category between the selected species (rs = 0.28, P > 0.1). This suggests that there are species that, from the loss of natural habitat perspective, must be considered as priority in the assignation of a risk category, according to the NOM. Also, no correlation was observed between the percentage of remnant natural habitat and the size of the distributional area of the endemic species (r = 0.16, P > 0.1), which indicates that the percentage of loss of natural remnant habitat is independent of the size of the distributional areas of these species. This shows the second relevant tendency, which implies that the loss of natural remnant habitat is more associated with the geographic region than with the area of distribution per se (figure 1). The latter demonstrates that the Mexican Neotropic located east-southeast of México and the Transvolcanic Belt are priority regions for conservation for these endemic species, given the high percentage of natural remnant habitat loss they showed, independently of the size of their distributional area [41]. In contrast, the species distributed in the northern region of the country, including the Mexican Altiplano, northern state and Baja Peninsula showed the highest values of natural
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remnant habitat in their distribution, independently of their distributional area. Consequently, the second important result derived from this study reflects that the Neotropical region located east and southeast of México (i.e. Gulf of México Plains) and the Transvolcanic Belt must be considered as priority for conservation given that they present a high loss of natural remnant habitat (see priority sites for conservation below).
Climate Change Impact Scenarios on the Distribution of Endemic Species The scenarios of the impact of climate change on the endemic species showed different tendencies than those of deforestation. For climate change, the Altiplano region (northern states of México) was the one that showed the greatest reduction in the distributional area of the species, in both scenarios. These results suggest a tendency toward a greater change in the reduction of the distribution of the species distributed north of México compared to less reduction in the distribution towards the species located in the Transvolcanic Belt. The species with Neotropical distribution (Gulf of México Plain) presented intermediate changes, although important ones (figure 2). We will describe the comparative changes in the distribution of endemic vertebrate species selected for the A2 and B2 scenarios for 2020, 2050 and 2080, respectively:
Scenario A2 Under climate change scenario A2 (CC A2), a low proportion of the endemic species selected showed an increment in their distribution for the year 2020; a lesser proportion showed an increment of less than 50% of their distribution for 2050 and a minimum part of the endemic species showed an increment for 2080 (figure 2). In contrast, the majority of endemic vertebrate species selected showed a decrease in their distribution for the year 2020; this tendency was even greater for 2050. In 2080, the proportion of species in which a decrease in their distributional area predicted greater than 80% is reached by almost 40% of the total number of species selected (figure 2). Scenario B2 Under the more conservative climate change scenario, B2 (CC B2), a similar tendency was observed to the one presented for the CC A2, that is, the majority of species selected decrease their distributional area and the proportion of species with the greatest decrement in incremented in the 2050 and 2080 scenarios, respectively (figure 2). Given that, some interesting differences are also noted; for example, the proportion of species that showed and increment in their distributional area was less for 2020, 2050 and 2080 in the CC B2 compared to CC A2. So, the proportion of species that decreased their distribution was greater for the CC B2 scenario compared to the CC A2 scenario for 2020, 2050 and 2080, respectively (figure 2). This means that the CC A2 scenario puts a greater proportion of the endemic vertebrates selected at risk of extinction, compared to the CC B2 scenario. It is probable that these differences result from the ecological requirements of the selected species. For example, it is expected that under the CC A2 scenario (most severe scenario), the temperature will be greater and the precipitation lower than actual, in comparison to the CC B2 scenario (more conservative scenario).
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Figure 2. Percentage of the change in the size of the distributional area for the endemic species selected and species of the genera Opuntia selected for this chapter. Green color indicates the proportion of the area incremented in the specie‟s distribution, while the red colors indicate area loss under the climate change scenarios (A2 and B2 respectively).
It is known that terrestrial vertebrates, and particularly the selected endemic species, depend heavily on primary productivity, since they are primary consumers, being mainly herbivorous, frugivorous, insectivorous, seed consumers, etc. [14, 15]. Primary productivity is directly related, on the other side, to precipitation. Under these order of ideas, it is expected that under the CC B2 scenario a greater primary productivity occurs than under CC A2, which would bring as consequence a greater impact in most of the distributions of the endemic species selected for this chapter. Given the above, it is indispensable to include a greater number of faunistic and floristic species with different ecological requirements, where a broader spectrum of responses can be analyzed in response to both CC A2 y B2 scenarios. Finally, when comparing the results of the impact of deforestation and climate change an unfavorable tendency was observed regarding the conservation of these endemic species. On one side, endemic species with distributions in northern México (Mexican Altiplano) showed a lesser effect regarding deforestation, but an important effect for climate change. Species with distributions in the Transvolcanic Belt and the Neotropic (Plains of the Gulf of México) showed a significant reduction in their distribution as a consequence of deforestation, but the expected impact of climate change is less than that of deforestation. These tendencies show that a significant proportion of the country, i.e. north México, the Transvolcanic Belt and the Mexican Neotropic, must consider relevant actions of conservation in virtue that deforestation and climate change affect in a geographical differential manner the endemic species included in this chapter (figures 1 and 2). These results are the first ones to appear in a conjunct manner in a research regarding the whole country. The expected impact of deforestation and climate change on the selected endemic species is significant and requires of mitigation measures to diminish their negative effect. Even if the
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problem requires a more profound analysis, some immediate measures can be considered as viable proposals to establish a national level strategy. For example, facing the growing rate of deforestation in the Transvolcanic Belt region, it has been proposed a network of priority conservation areas that connect decreed natural protected areas; this network could minimize the conservation area needed, and maximize the inclusion of biodiversity [41].
Selection of Priority Sites for Conservation for the Selected Species Under Actual and Climate Change Scenarios Priority sites of conservation, when considered actual distribution were generated including three different goals: 15%, 20% and 25% of the area of México (see Methods; figures 3 A-C).
Figure 3A. Selection of sites considering the actual distribution of the species selected, with a 15% area for the whole country.
Figure 3B. Selection of sites considering the actual distribution of the species selected, with a 20% area for the whole country.
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Figure 3C. Selection of sites considering the actual distribution of the species selected, with a 25% area for the whole country.
It is notorious that priority sites coincide, in general terms, for the three established goals, highlighting the shoreline of the Gulf of México, as well as important areas of the states of Tabasco, Chiapas, and Yucatán. It also strikes as important the Gulf of California area as a priority site of conservation under actual distribution scenarios for the selected species. It is interesting to note that in the scenario that considers 25% of the territory important areas of the Transvolcanic Belt and the Mexican Altiplano were selected, while in the 15 an 20% goal scenario this regions were not selected (figures 3 A-C). Using the same techniques, priority sites for conservation under climate change scenarios were modeled, considering 15%, 20% and 25% of the land of México. In this case, no significant differences were observed (figures 4 A-D). It is interesting to note that the areas northwest of México and the Baja Peninsula were identified as priority areas for conservation in all three goals (i.e., 15%, 20% y 25%; figures 4 A-C). Nonetheless, in the scenario considering 25% of the country areas in the Gulf of México plains and the Yucatán Peninsula were also selected (figure 4D). It is important to notice that it is the first analysis of this kind done for México.
Figure 4A. Selection of priority sites with a 15% restriction of the total area of the country, under climate change scenarios.
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Figure 4B. Selection of priority sites with a 20% restriction of the total area of the country, under climate change scenarios.
Figure 4C. Selection of priority sites with a 25% restriction of the total area of the country, under climate change scenarios.
Figure 4D. Comparison of the selected sites for conservation, with the mentioned restrictions, under climate change scenarios.
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CONCLUSION Analysis on the State of the Art of the Techniques and Their Implementation in Conservation Actions Up until now, some multi-taxa studies have been conducted addressing the impacts of deforestation and climate change on different groups of plants and animals in the country. An interesting scope derived from some of these studies involves the use of biological information deposited in scientific collections (and available through data bases; see www.conabio.gob.mx and www.ibiologia.unam.mx), thematic digital maps for environmental variables and a series of geographic information systems [21]. All of these offers a unique opportunity to include a considerable amount of flora and fauna groups in a multi-taxa evaluation of the impact of climate change over the distribution of biodiversity in México. However, maybe the most notorious advance in the matter is the theoretical background in which these analyses are based. For example, ecological niche modeling projected as their potential distribution allows the projection of these distributions in deforestation and climate change scenarios –as was done in the present chapter. Under the assumption of the niche conservadurism [42], robust hypotheses can be generated regarding expected changes in the distribution of species under climate change scenarios. Finally, we can state that if the biological information exists (databases in scientific collections), environmental information (thematic digital maps, for example A2 CC and B2 CC scenarios) and available software, multi-taxa analyses of the kind presented here can be expanded for all México. Our proposal is that a strategy to “fill-in” the gaps in information that exists today is to incorporate all available information with the information of the impact of deforestation and climate change in the distribution of biodiversity in México. Once generated the distributional models under these scenarios, networks of priority areas for conservation can be introduced, based on optimization algorithms, as the case presented here for the Transvolcanic Belt area. This perspective is, perhaps, the best option at this time to approach this complex problem.
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[22] Sánchez-Cordero, V., M. Munguía and A. T. Peterson. 2004. GIS-based predictive biogeography in the context of conservation. In: Frontiers in Biogeography. M. Lomolino and L Heaney (eds.). Sinauer Press. Capítulo 16. [23] Sánchez-Cordero, V., P. Illoldi-Rangel, M. Linaje, S. Sarkar and A. T. Peterson. 2005. Deforestation and extant distributions of Mexican endemic mammals. Biological Conservation 126:465-473. [24] Sánchez-Cordero, V., P. Illoldi-Rangel, M. Linaje, T. Escalante, F. Figueroa and S. Sarkar. 2009. Deforestation and biodiversity conservation in Mexico. In: A. Columbus and L. Kuznetsov, eds.). Endangered Species: New Research. Nova Science Publishers. New Haven, USA. [25] Holland, J. H. Adaptation in Natural and Artificial Systems: An Introductory Analysis with Applications to Biology, Control, and Artificial Intelligence. Ann Arbor, MI: University of Michigan Press, 1975. [26] Stockwell, D. R. B. and D. Peters. 1999. The GARP modelling system: problems and solutions to automated spatial prediction. International Journal of Geographical Information Science, 13:143-158. [27] Dennis, R.L.H. and Thomas, C.D. 2000. Bias in butterfly distribution maps: the influence of hot spots and recorder‟s home range. Journal of Insect Conservation, 4, 73–77. [28] Peterson, A. T. 2003. Predicting the geography of species' invasions via ecological niche modeling. Quarterly Review of Biology 78:419-433. [29] Soberón J, Peterson AT. 2005. Interpretation of models of ecological niches and species‟ distributional areas. Biodiversity Informatics 2: 1-10. [30] Phillips SJ, Dudìk M, Schapire RE. 2004. A maximum entropy approach to species distribution modeling. In: Bratko I, Dˇzeroski S, editors, Proceedings of the 21st International Conference on Machine Learning, New York: ACM Press. pp. 655-662. [31] Parra-Olea G., M. García-París, T. J. Papenfuss, and D. B. Wake. 2005. Systematics of the Pseudoeurycea bellii species complex. Herpetologica 61: 145-158.41. Fuller, T., M. Mayfield, M. Munguía, V. Sánchez-Cordero and S. Sarkar. 2006. Incorporating connectivity into conservation planning: A multi-criteria case study from Central Mexico. Biological Conservation. 133(2):131-142. [32] Ortega-Huerta, M. A., and A. T. Peterson. 2004. Modelling spatial patterns of biodiversity in northeastern Mexico. Diversity and Distributions, 10:39-54. [33] Margules, C. R., Nicholls, A. O. and Pressey, R. L. 1988. Selecting networks of reserves to maximise biological diversity. Biol. Conserv. 43, 63-76. [34] Margules CR, Pressey RL .2000. Systematic conservation planning. Nature 405: 245253. [35] Sarkar, S. 2003. Conservation area networks. Conservation and Society 1: v-vii. [36] Sarkar, S., Margules CR. 2002. Operationalizing biodiversity for conservation planning. Journal of Biosciences 27(S2): 299-308. [37] Sarkar, S. 2005. Biodiversity and Environmental Philosophy: An Introduction to the Issues. Cambridge, UK: Cambridge University Press. [38] Garson, J., Aggarwal, A., Sarkar, S. 2002. Birds as surrogates for biodiversity: an analysis of a data set from southern quebec. Journal of Biosciences 27, 347–360.
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[39] Kelley, C., Garson, J., Aggarwal, A., and S. Sarkar. 2002. Place Prioritization for Biodiversity Reserve Network Design: A Comparison of the SITES and ResNet Software Packages for Coverage and Efficiency, Diversity and Distributions 8: 297 – 306. [40] Sarkar S, Aggarwal A, Garson J, Margules CR, Zeidler J. 2002. Place prioritization for biodiversity content. Journal of Biosciences 27(S2): 339-346. [41] Sarkar, S., Pressey, R.L., Faith, D.P., Margules, C.R., Fuller, T., Stoms, D.M., Moffett, A., Wilson, K.A., Williams, K.J., Williams, P.H., and Andelman, S. 2006. "Biodiversity Conservation Planning Tools: Present Status and Challenges for the Future," Annual Review of Environment and Resources 31: 123 - 159. [42] Peterson, A. T., J. Soberón, and V. Sánchez-Cordero. 1999. Conservatism of ecological niches in evolutionary time. Science 285:1265-1267.
In: Endangered Species Editor: Manuel Esteban Lucas-Borja
ISBN: 978-1-62257-532-9 © 2012 Nova Science Publishers, Inc.
Chapter 4
IS POST-DISPERSAL SEED PREDATION A PROBLEM FOR SPANISH BLACK PINE (PINUS NIGRA ARN. SSP SALZMANNII) NATURAL REGENERATION IN CUENCA MOUNTAINS (SPAIN)? Manuel Esteban Lucas-Borja Castilla La Mancha University, School of Advanced Agricultural Engineering, Department of Agroforestry Technology and Science and Genetics, Campus Universitario, Albacete (Spain)
ABSTRACT Black pine (Pinus nigra Arn.) forests appear the EU endangered habitats list of natural habitats which require specific conservation measures (Resolution 4/1996 by the Convention on the Conservation of European Wildlife and Natural Habitats) due in part to the lack of basic understanding about the regeneration biology of this long-lived species. Seed predation has been lately described as important factor in relation to Spanish black pine natural regeneration. Ants, birds and rodents have been described as the main post-dispersal seed predators although the impact varies depending on forest location, seed characteristics, phenological cycles of plants and animals, feeding habits of predators and habitat characteristics. The objective of this study was to assess the Spanish black seed postdispersal seed predation under different forest stand densities over several years in two different sites, one located in a continuous homogeneous Spanish black pine forest and other at the ecological limit of their distribution in the Cuenca Mountain area (Spain). The importance of each predator group (ants, birds and rodents) in relation to seed predation was also evaluated. Our study showed that the overall predation rates of Spanish black pine seeds by the three groups of predators were quite high in both experimental sites (the central population and at the upper altitude limit of the species distribution) during low seed rain years (almost 90% of predated seeds in 2005 and 2007) when compared to the mast year of 2006 (less than 25% of predated seeds). Moreover, the contribution of each predator group on the overall predation showed temporal variations in the rich-year of seed production (2006).
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Manuel Esteban Lucas-Borja Thus, Spanish black pine regeneration felling and management plans should take into account the mast year factor but also post-dispersal seed predation rates since in non masting years, seed predation can avoid natural stand regeneration as about 90% of seeds could beaten.
1. INTRODUCTION European black pine is a tertiary relictual species belonging to the group of Mediterranean pines and it is an important pine species in forest management of Mediterranean areas (Bogunic et al. 2007) and particularly in Spain where sub-specie salzmannii (Spanish black pine) can be found (Lucas-Borja 2008). Their forests are included in the EU endangered habitats listing of natural habitats requiring specific conservation measures (Resolution 4/1996 by the Convention on the Conservation of European Wildlife and Natural Habitats) due in part to the lack of basic understanding about the regeneration biology of this long life species (Kerr 2000; Tíscar 2003). Most of the information regarding this topic is based on opinions and observations derived from the authors‟ own experience during their professional years, but they are not backed by studies which are intended to quantify it (Serrada et al. 1994). The natural regeneration of forest stands is an important subject in both conservation biology and management being constrained by abiotic factors such as physiographic and climatic variables, and by the effect of biotic factors (Silvertown and Lovett-Doust, 1993; Zackrisson et al., 1995; Castro et al., 1999). Furthermore, these negative effects may dramatically increase when the species is restricted to a cluster of isolated populations instead of a continuous homogeneous population found at the most common site habitat (Castro et al., 1999). Differences in climatic characteristics and soil conditions can be important in modifying seed production and seed germination trends at high or low altitudes and at isolated populations (Mencuccini et al., 1995; Del Cerro et al., 2009). The seed, seedling and sapling stages have been determined as the most important stages in natural regeneration since seed dispersal, predation, germination and survival have been considered as a major force governing the regeneration, structure and succession of vegetation (Clark et al., 1999; Dalling et al., 2002; Tíscar, 2003). Thus, information on plant seed dispersal, natural loss dynamic of seeds (mainly due to post-dispersal seed predation) and germination are critical for understanding the natural regeneration mechanism of the focal species. Seed predation can limit population recruitment not only reducing seed availability (Schupp 1995; Tíscar 2003) but also changing the spatial distribution of seeds in the initial seed rain (Schupp 1995). This implies that the availability of safe sites for seed depends on the interaction between seed rain distribution, habitat structure and preferences of postdispersal seed predators. Pardos et al. (2005) summarized that predation of Quercus ilex acorns can affect up to 50% of the annual crop. To Pinus jeffreyi and Pinus ponderosa seeds, Vander (1994) estimated a half-life on the ground of 120 hours and predicted that less than 1% of the seed crop would go undetected by seed predators. Ants, birds and rodents have been described as the main post-dispersal seed predators (Hulme and Hunt 1999; McShea 2000). The predation of these groups of animals may be very variable, depending on seed characteristics (Kollmann et al. 1998), phenological cycles of plants and animals (Denslow 1987) and feeding habits of predators or habitat characteristics (Hulme 1997).
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Information on plant seed dispersal, natural loss dynamics of seeds (mainly due to postdispersal seed predation) and germination are critical for understanding the natural regeneration mechanism of the species and to manage for its sustainability. Therefore, the objective of this study was to follow the Spanish black pine post-dispersal seed predation by ants, rodents and birds under different forest stand densities. The study was conducted in a Spanish black pine central population and at the upper altitude limit of the species distribution in the Cuenca Mountain area (Spain). We hypothesized that post-dispersal seed predation percentage depends on forest location (central or peripheral Spanish black distribution) and is affected by stand density, predator group and season of the year.
2. MATERIAL AND METHOD The study area is located in the region of Castilla La Mancha, Central-eastern Spain. A continuous homogeneous Spanish black pine experimental forest was selected at the most common and representative location of this species in the Cuenca mountain area (Los Palancares y Agregados (PAL), 1200 m above sea level, 40º01´50´´N; 1º59´10´´W). According to Del Cerro et al. (2009) Spanish black pine naturally occurs in this area between 1000 and 1500 m and dominates the forest stand composition. A second experimental forest was selected at the ecological limit of the Spanish black pine distribution in the Cuenca mountain area. (Ensanche de las Majadas (MAJ), 1490 m above sea level, 40º 14´ 30´´ N; 1º 58´ 10´´ W). Here, this species is located in its altitudinal limit and Scots pine (Pinus sylvestris) dominates the forest stand composition in certain locations, relegating Spanish black pine to isolated or relict populations, which in turn are fragmented into smaller stands. Both experimental sites present even-age stands of about 100 years old, which have been naturally regenerated by the shelterwood system with a 100 year rotation and a 20 year regeneration period. Other forest stand characteristics and more details of each experimental site can be observed in Table 1. Table 1. General characteristics of the forest stand at each experimental sites Site 1a 1b 1c 2a 2b 2c .
-1
Experimental site “Los Palancares y Agregados” “Los Palancares y Agregados” “Los Palancares y Agregados” “Ensanche de las Majadas” “Ensanche de las Majadas” “Ensanche de las Majadas”
N.ha-1
G
Dm
Hm
Ho
Principal forest Shrub Stand composition cover
295
15-20
23.16
12.33
12.78
100% Pn
56 %
568
25-30
22.16
13.36
13.71
100% Pn
53 %
789
35-40
21.19
12.98
13.27
100% Pn
48 %
287
15-20
24.44
11.72
12.23
70% Pn; 30% Ps
54 %
503
25-30
22.56
11.73
12.71
75% Pn; 25% Ps
51 %
767
35-40
20.14
11.02
12.54
72% Pn; 28% Ps
49 %
2
N ha : Number of trees per ha; G: Basal area (m /ha); Dm: Average stand diameter (cm); Hm: Average stand height (m); Ho: Dominant stand height (m). Pn: Pinus nigra Arn. subsp. salzmannii; Ps: Pinus sylvestris
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The experimental sites are separated by 50 km. According to Allué (1990), the climate characteristics of the two experimental forests can be classified as Mediterranean humid climate, although in general and within this level, Los Palancares y Agregados present lower annual precipitation (mean annual precipitation of 1137 mm and 595 mm for Ensanche de Las Majadas and Los Palancares y Agregados, respectively) and higher temperatures (mean annual temperature of 9.6 ºC and 11.9 ºC for Ensanche de Las Majadas and Los Palancares y Agregados, respectively). Calcareous soils are dominant in the Cuenca mountains area and the soil found in both experimental sites can be classified as Leptosol, according to the Soil Atlas of Europe (2005). In both experimental sites, the slope is less than 5º and the exposure is north. According to the Cuenca Mountain Forest Service, the granivorous species considered as potential predators of Spanish black pine seeds can be classified into three groups: 1) Aphaenogaster subterranean, Messor capitatus and M. structor for ants; 2) Fringilla coelebs, Parus caeruleus, Parus major, Passer domesticus, Columba palumbus and Serinus serinus for birds and 3) Apodemus sylvaticus and Mus musculus for rodents. All predatory groups were allowed in all experimental conditions, although the effect of each predatory group was tested by excluding the other two with specific experimental devices (Ordoñez and Retana, 2004). The device for ants consisted of a transparent plastic tube 6 cm in length and 0.5 cm diameter, in which we placed one Spanish black pine seed. The device for rodents was a wire cylinder of 0.5 cm grid, 4 cm diameter and 20 cm length open at both ends covering a 3 x 3 cm piece of wire with a Spanish black pine seed glued in the centre. The device for birds was a 3 x 3 cm piece of fine wire mesh with a Spanish black pine seed glued on in the centre of the piece, thus avoiding ant predation.
Figure 1. Pictures of the seed predator‟s devices (modified from Ordoñez and Retana, 2004).
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In field experimental tests we established that ants were able to find seeds relatively easily within these tubes and that birds (tests were carried out with sparrows, Passer domesticus) managed to eat seeds from these devices without problems caused by the glue. Rodents (tests were carried out in laboratory with Apodemus sylvaticus, the most common rodent in the study areas) were able to eat seeds from these devices also without problems caused by the glue. To monitor ant predation, ant devices were randomly distributed on each plot. Devices were fixed to the ground to prevent rodent and bird manipulation. In order to monitor rodent predation, rodent devices were also randomly placed on each plot. To test bird predation, we fenced a 3 x 3 m subplot with 0.5 cm wire mesh grid inside where ant and rodent devices had been previously placed. This mesh, 1 m high and buried 25 cm into the ground, prevented rodent access to seeds. Bird devices were randomly distributed around each plot. Seed removal was monitored on a weekly basis in all experimental devices, between 2005 and 2007 from early January to the end of May (in a total of 7 dates/year), which corresponds to the period of the year when Spanish black pine seeds could be present in the soil bank between seed fall in winter and seed germination in spring. Nine seeds/devices per predator group and basal area interval were monitored in each survey date and once the seed was removed by predators, the devices were substituted. Different sampling methods were used to evaluate the abundance of the three seed predator groups in the different experimental sites and forest stand density in the four seasons of years 2005, 2006 and 2007. In each sample, all species were identified but only those likely to predate Spanish black pine seeds were considered for further analyses. The dietary spectrum of the different species was determined following the criteria of experts. Rodent populations were sampled on each site using two-live trapping transects at intervals of 15 × 20 m along a whole area of 1 ha. Seven trap stations with a Sherman live-trap each were placed at 10 m intervals in each transect. Traps were baited with a mixture of toasted bread and canned tuna, and were checked twice a day during two days per sampling period. The total trapping effort accounted for 102 trap-nights per site. The captures obtained in each plot were used to calculate an abundance index, as the number of individuals caught per trap and night. Bird abundance was estimated using point counts (Bibby et al. 1992). Each 1 ha site, was surveyed per season by a two 20 minute fixed point counts, early in the morning and on the afternoon of the same day. The number of individuals heard (singing or calling) or seen for each species was recorded. The surveys where carried out only with appropriate weather and moment in time conditions by the same ornithologist (Ordoñez and Retana 2004). The number of granivorous birds recorded within the 1 ha surface during the two counts was used as a measure of bird abundance. Pitfall traps (plastic vials of 6 cm diameter and 7 cm deep partially filled with water, ethanol and soap) were used to measure the ground ant abundance of each site, as they provide good estimates of the relative abundance of ant species (Ordoñez and Retana 2004). In each site, ten traps were laid over an area of 3 × 3 m during ten days. After that period, ants that had fallen into the traps were collected, sorted and identified in the laboratory to the species level. Analyses of variance were performed according to split-plot analysis models. All statistical analyses were performed using the JMP 7.0® statistical software (SAS Institute Inc. Cary, North Caroline, USA).
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3. RESULTS There were significant differences in seed removal among years since the percent of removal was lower in the seed-rich year of 2006 (mean±se 25.15±3.7%) than in the seed-poor years of 2005 (98.71±9.0%) and 2007 (97.29±11.6%). Removal rates also differed between experimental sites and survey dates for 2006, while experimental site had not significant effect in 2005 and 2007. Stand density was not an important source of variation during the experiment. With respect to predator group, significant differences were found only for the seed-rich year where, both the interactions of experimental site versus predator group and predator group versus survey date were significant. There were not significant differences in the predation rates of birds and rodents between the two experimental forests (birds: 37.4±4.9% in Los Palancares y Agregados and 34.1±3.6% in Ensanche de Las Majadas; rodents: 21.2± 3.4% in Los Palancares y Agregados and 23.0± 3.9% in Ensanche de Las Majadas), whilst the predation rates of ants were significantly highest in Los Palancares y Agregados. At the same time, birds had different values in the tested period, ants showed the highest removal rate in spring and rodents had a peak in late winter during 2006 (Figure 3). In seed-poor years, the percent of removal was higher the late winter or spring season, decreasing in early winter (Figure 4).
Figure 2. Predated seeds found in the field experiments.
In relation to abundance of three seed predator group there were significant differences in the abundance of ants among season of the year, experimental sites and years. Abundance of ants was higher in Los Palancares y Agregados than in Ensanche de las Majadas and in spring and summer than in autumn and winter. The abundance of both birds and rodents showed significant different among season and years but not for the other two factor, experimental site and stand density. The maximum values for rodents were found in winter and autumn and
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the minimum ones in summer or spring, while for birds the high abundance percentage was estimated to occur in winter and spring and the lowest ones in summer or autumn. Stand density was not an important factor related to abundance of seed predator groups. Seed-rich year propitiates a higher abundance of the three seed predator group in comparison to 2005 and 2007, where seed rain was lower for two experimental sites (Figure 5).
Figure 3. Predated seeds (%) by the three predator groups in the different sampling season tested in 2006. Vertical bars are the ±se.
Figure 4. Predated seeds (%) in the monitoring dates (Month-day) in 2005 and 2007. The three predator groups were gathered for each year since no statistical differences were found between them. Standard error is indicated by vertical bars.
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Figure 5. Mean (+se) abundance of the three predator groups in the sampling season of 2006 and experimental sites. Abundance units of the three groups are: ants, number per 10 pitfall traps; bird, number recorded within the 1-ha surface during two counts of 40-min each; rodents, number per 14 traps and night.
4. DISCUSSION AND CONCLUSION Although the literature on post-dispersal seed predation is voluminous (Crawley, 1992), studies usually provide little more than a list of the guilds of animals responsible with few details of their relative importance. In most studies of post-dispersal seed predation, seeds are completely removed, leaving little sign of the predator‟s responsible (Hulme, 1999). At the present work, the overall predation rates of Spanish black pine seeds by the three groups of predators were quite high for relict and typical location during poor-year (years 2005 and 2007) than in rich-year seed production (year 2006). Result are very similar to other studies
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reported previously in Spanish black pine (Ordonez and Retana, 2004) and lower than those summarized for other pine species in the Mediterranean area (Castro et al., 1999). Seed predation may obstruct recruitment dynamics although in good years of seed production this effect is less important as predation rates reach lower percentages. Janzen (1978) argued that masting is an evolutionary response to seed predation. Under this “predation satiation” hypothesis, proportional seed survival is expected to increase with seedfall density as predator are overwhelmed by the abundance of seed (Schupp, 1990). Results evidence that years of heavy seedfall disproportionately satiate the postdispersal seed predation of Spanish black pine seeds, which is consistent with others studies of post-dispersal predation (Nilsson and Wästljung, 1987). On the other hand, forest stand density has not influenced the predations rates. Other authors have reported that small rodents and granivorous birds tend to concentrate their activity under or near shrub, where they find protections from predators (Simmonetti, 1989) thereby augmenting seed predation in those microhabitats (Herrera, 1984). However, in our study shrub coverage was similar in all cases and stand density effects was not as important as could be in areas without shrub cover, where seed predators could not easily find shelter. The contribution of each predation group to overall predation showed temporal variations in richyear seed production. No conclusion can be obtained for 2005 or 2007, where seed fall was almost non-existent and all seeds were removed. For 2006, the first seeds dispersed in winter were mainly predated by rodents, which also registered their highest abundance in this season of the year. Thus, the high rodent abundance corresponded in time with the offer of an attractive resource such as Spanish black pine seeds in the least seed productive period of the year in the Mediterranean area (De Lillis and Fontanella, 1992). Ants became more important predators, coinciding with their increased abundance in spring, at the end of the natural dissemination period of Spanish black pine seeds. The highest outside activity of ants occurs in summer (Cros et al., 1997), when temperatures are higher than the others season of the year. In fact, ants are a more important predator group in typical Spanish black pine forest than in relict areas, where temperatures and climate conditions are not favourable for these insects. Birds showed the highest predation values, which is probably related to their feeding habits, since, birds have larger feeding areas than rodents and ants in general (Peters, 1983). Although differences were not significant, the highest values of seed predation by birds in spring might be related to the fact that many bird species have their nesting and breeding periods in this season, and the food requirements of adults increase to supplement the nourishment of the brood (Ordoñez and Retana, 2004). With respect to the survey date, predation rate were always higher in the middle of the dissemination period, when the animals have assimilated the new source of food. In late spring, others plants and fruits can increase the availability of food for each predator groups, decreasing Spanish black pine seed predation rates. Post-dispersal seed predation do not predict good expectancies for typical or relict Spanish black pine forest in non-masting years, being especially problematic in those areas where this specie reaches their altitudinal limit. Good year‟s recruitment has to be preceded by a good seed production, in which case post-dispersal seed predation is not as important factor as if low seedfall is noted. Predators present different seed removal rates depending on the season of the year and forest ecosystem location, being seed predation an important limitant factor in natural regeneration of Spanish black pine in poor-year seed productions.
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ACKNOWLEDGMENTS The author thanks the Cuenca Mountain Forest Service for site access. This contribution has been supported by the Castilla La Mancha Government (project code POII10-0112-7316). In Memoriam to Prof. Antonio Del Cerro Barja, who established, guided and developed many different research projects about Spanish Black pine natural regeneration in Cuenca Mountain and also educated many forest engineers during his work at Castilla La Mancha University.
REFERENCES Allue Andrade, J.L., 1990. Atlas fitoclimático de España. Taxonomías. Ministerio de Agricultura, Pesca y Alimentación. Instituto Nacional de Investigaciones Agrarias, Madrid. Bibby, C. J., Burgess, N. D., Hill, D. A. 1992. Bird census techniques. Academic Press. London Bogunic, F., Muratovic, E., Ballian, D., Siljak-Yakovlev, S., Brown, S., 2007. Genome size stability among five subspecies of Pinus nigra Arnold s.l. Environ. Experim. Bot. 59, 354-360. Castro, J., Gómez, J. M., García, D., Zamora, R., Hódar, J. A., 1999. Seed predation and dispersal in relict Scots pine forests in southern Spain. Plant Ecology 145, 115-123. Clark, J.S., Beckage, B., Camill, P., Cleveland, B., Hille Ris Lambers, J., Lichter, J., McLachlan, J., Mohan, J. and Wyckoff, P. 1999 Interpreting recruitment limitation in forests. Am. J. Bot. 86, 1–16. Crawley, M. J. 1992. Seed predators and plant population dynamics. In: Fenner, M. (ed.), Seeds, the ecology of regeneration in plant communities. CAB Int., New York. pp. 167182. Cros, S., Cerdá, X., Retana, J., 1997. Spatial and temporal variations in the activity patterns of Mediterranean ant communities. Ecoscience 4, 269-278. Dalling, J.W., Muller-Landau, H.C., Wright, S.J., Hubbell, S.P., 2002. Role of dispersal in the recruitment limitation of neotropical pioneer species. Journal of Ecology 90, 714-727. De Lillis, M., Fontanella, A., 1992. Comparative phenology and growth in different species of the Mediterranean maquis of central Italy. Vegetatio 99-100: 83-96. Del Cerro, A., Lucas-Borja, M.E., Martínez García, E., López Serrano, F.R., Andrés Abellán, M., García Morote, F.A., Navarro López, R., 2009. Influence of stand density and soil treatment on the Spanish black pine (Pinus nigra Arn. ssp. salzmannii) regeneration in Spain. Invest. Agrar.: Sis. Recur. For. 2009 18(2), 1-14. Denslow, J. S., 1987. Fruit-removal from aggregated and isolated bushes of the red elderberry, Sambucus pubens. Can. J. Bot. 65, 1229-1235. Herrera, C., 1984. Avian interference of insect frugivory: an exploration into the plant-birdfruit pest evolutionary triad.Oikos 42, 203-210.
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Hulme, P. E., 1997. Post-dispersal seed predation and the establishment of vertebrate dispersed plants in Mediterranean scrublands. Oecologia 111, 91-98. Hulme, P. E., Hunt, M. K., 1999. Rodent post-dispersal seed predation in deciduous woodland: predator response to absolute and relative abundance of prey. J. Anim. Ecol. 68, 417-428. Janzen, D. H., 1978. Seeding patterns of tropical trees. In Tomlinson, P. B., Zimmermann, M. H. (eds.),Tropical Trees as Living Systems, Cambridge University Press, Cambridge, pp. 83–128. Kerr, G., 2000. Natural regeneration of corsican pine (Pinus nigra subsp. laricio) in Great Britain. Forestry, 73 (5), 479-487. Kollmann, J., Coomes, D. A., White, M., 1998. Consistencies in post-dispersal seed predation of temperate fleshy-fruited species among seasons, years and sites. Funct. Ecol. 12, 683-690. Lucas Borja, M.E., 2008. La regeneración natural de los montes de Pinus nigra Arn. ssp salzmannii en la Serranía de Cuenca; Bases para la gestión forestal del monte "Los Palancares y Agregados" (CU). PhD Thesis dissertation, Castilla la Mancha University, Spain. McShea, W. J. 2000. The influence of acorn crops on annual variation in rodent and bird populations. Ecology 81, 228-238. Mencuccini, M., Piussi, P., Zanzi Sulli, A.,1995. Thirty years of seed production in a subalpine Norway spruce forest: patterns of temporal and spatial variation. Forest Ecology and Management 76, 109-125. Nilsson, S.G., Wästljung, U. 1987. Seed predation and cross-pollination in mast-seeding beech (Fagus sylvatica) patches. Ecology 68, 260-265. Ordóñez J.L., Retana, J., 2004. Early reduction of post-fire recruitment of Pinus nigra by post-dispersal seed predation in different time-since-fire habitats. Ecography 27, 449458. Pardos, M., Ruiz del Castillo, J., Cañellas, I., Montero, G., 2005. Ecophysiology of natural regeneration of forest stand in Spain. Invest. Agrar.: Sist. Recur. For. 14(3), 434-445. Peters, R. H., 1983. The ecological implications of body size. Cambridge Univ. Press, London. Schupp, E. W., 1990. Annual variation in seedfall, postdispersal predation and rcruitment of a neotropical tree. Ecology 71, 504-515. Schupp, E. W., 1995. Seed-seedling conflicts, habitat choice, and patterns of plant recruitment. Am. J. Bot. 82, 399-409. Serrada Hierro, R., Dominguez Lerena., S., Sánchez Resco, Mª. I., Ruiz Ortiz, J., 1994. El problema de la regeneración natural del Pinus nigra Arn. Rev. Montes 36, 52-57. Silvertown, J. W., Lovett-Doust, J. 1993. Introduction to plant population biology. Blackwell. Scientific Publications, Oxford, UK. Simonetti, J.A., 1989. Microhabitat use by small mammals in central Chile. Oikos 56, 309318. Soil Atlas of Europe 2005 European Soil Bureau Network. European Commision, Luxembourg.
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Tíscar Oliver, P. A., 2003. Condicionantes y limitaciones de la regeneración natural en un pinar oromediterráneo de Pinus nigra subsp. Salzmannii. Invest. Agr.; Sist. Recur. For. 12(2), 55-64. Vander Wall, S. B., 1994. Seed fate pathways of antelope bitterbrush: Dispersal by seedcaching yellow pine chipmunks. Ecology 75, 1911-1926. Zackrisson O., Nilsson, M.C., Steijlen, I., Hörnberg, G., 1995. Regeneration pulses and climate-vegetation interactions in nonpyrogenic boreal Scots pine stands. Journal of Ecology 83, 469-483.
In: Endangered Species Editor: Manuel Esteban Lucas-Borja
ISBN: 978-1-62257-532-9 © 2012 Nova Science Publishers, Inc.
Chapter 5
MEDITERRANEAN HORNY SPONGES: HOW TO DRIVE A NEVERENDING STORY OF EXPLOITATION TOWARD A SUSTAINABLE MANAGEMENT AND CONSERVATION Roberto Pronzato1,*, Fabio D. Ledda1,2 and Renata Manconi2 1
Dipartimento di Scienze della Terra, dell'Ambiente e della Vita, Università di Genova, Corso Europa, Genoa (Italy) 2 Dipartimento di Scienze della Natura e del Territorio, Università di Sassari, Via Muroni, Sassari (Italy)
ABSTRACT The most recent biodiversity assessment on Mediterranean sponge fauna indicates a value of species richness higher than 600 with over 30% of endemicity. This data supports the status of the Mediterranean Sea as a precious hotspot of biodiversity. The Mediterranean sponge taxonomic richness is under pressure and losses are possible because of epidemic diseases and over-fishing of bath sponges, together with a huge human impact due to both discharge of pollutants and alteration of ecosystems since ancient times. International agreements on the marine fauna conservation focus on 12 Mediterranean sponge species, including bath sponges. This list urgently needs a deep revision and enlargement. During the last decades, the number of species that could be actively harvested is dramatically increasing because new trends in the industrial exploitation of bioactive metabolites produced by sponges; this tendency, in the absence of any knowledge on sponge population distribution and density, could lead to local extinctions. The present overview deals with the state of the art on the historical exploitation of Mediterranean horny sponges. It also provides an update on trends in the utilization of these sponges seriously endangered by both long-term harvesting and habitat loss. Attention is also focused upon the species of recent commercial interest as sources of metabolites with biomedical or cosmetic potential, such as Dysidea avara, *
E-mail address:
[email protected].
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Roberto Pronzato, Fabio D. Ledda and Renata Manconi considered endangered due to a pressing demand of chemicals from the sea (e.g. Avarol). To face the uncontrolled harvesting of new and historical target sponge species, it is necessary to drive the Mediterranean sponge exploitation toward a rational and sustainable long-term management. Sponge culture in situ is suggested as a key approach to support a successful conservation strategy of this high value biological resource.
A CULTURAL HERITAGE A non-documented use of sponges was reported in the Egyptian and Phoenician civilizations [1]. “Sponge molds” decorations on the walls of the queen‟s house at Knossos (1900-1750 B.C.) historically ascertain the first use of sponges during the Crete-Minoic culture. “Sponge painting” is a well known, still used, early practice of wall decorating by a round sponge Hippospongia communis (Lamarck, 1813), soaked with paint and dabbed on the wall randomly, to produce the desired decorative effect [2]. Sponges have been variously represented on Greek pottery (765-755 B.C.) as “status symbol” objects or in thermal scenes [3]. Hómērus (6th-8th century B.C.) [4] [5] refers to sponges as hygienic tools in both his poems Yliade (XVIII, 414) and Odyssea (I, 111; XX, 151; XXII, 22, 439, 453). Aristophanes (405 B.C., XIV, 60) [6], Plato (360 B.C., XIV) [7], Athenaeus Naucratita (170-230 B.C., I, 32) [8] and other poets or philosophers also reported different uses of sponges and their diffusion in the ancient Greece. Historical syntheses on Greek sponges were recently performed [9] [10] [11]. The Latin literature of the Classic Roman Age also reports sponge‟s natural history (Plinius, 23-79, XI, 129) [12], while other authors wrote about the sponge use mainly in agriculture and cooking (Cato, ca. 160 B.C., CLXII, 3; Columella, ca. 65 A.D., XI, 3, 43; Apicius, 230 A.D., I, 26) [13] [14] [15]. Cicero (56 B.C., CXXII) [16] used the verb “spongiare” in the sense of washing or cleaning. Each legionary during the Roman Empire was supplied with sponges to be inserted under his own helmet and armour, and also used for cosmetic and personal care, and to drink. It is for this common use by Roman soldiers that we can read: “…and straight-way one of them ran, and took a sponge, and filled it with vinegar, and put it on a reed, and gave him to drink…” (Matthaeus, 1th century B.C., XXVII, 48) [17]. Sponges imbibed of honey were also used as baby-suckers, and imbibed alternatively in warm and cold water were largely applied in hydrotherapy. Sponges are also reported as human food. Aldrovandi, at page 585 of his “De reliquis animalibus exanguibus”, reports edible “Tethyae”, represented in illustrations [18] as sponges with the typical habitus of species of the genera Tethya, Suberites and probably Chondrosia. In the 19th century, Chondrosia reniformis Nardo, 1847, known in Italy with the most common trivial name “rognone di mare” (sea kidney) [19] [20], was sold in the market of Trieste (Italy) as “fegato di mare” (sea liver) [21]. Sponges wrapped in silk with a string attached, used by ancient Jews have been considered the most effective contraceptive in the past. “Vaginal sponges”, soaked in diluted lemon juice and vinegar, and inserted prior to intercourse as a barrier method of contraception, were largely used along many centuries. In 1823, the first campaign for birth control by the neo-Malthusian Francis Place [22] was opened up in England with contraceptive advices recommending the pre-coital insertion
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of a sponge into the vagina. Since that time, the “contraceptive sponge”, become popular up to the first half of the 20th century (1920-1960). In 1983, a modern synthetic tool, containing spermicidal, the “Today Sponge” was introduced in the United States [23] [24] [25].
Figure 1. To collect naphtha from the sea surface a) is a really uncommon use of sponges documented by Stradanus in “Venationes Ferarum”; b) in the Middle Age the “soporific sponge” was largely used as anesthetic by Arabian and European surgeons; this technique was still in use at the end of the 18th century (c).
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The plate 86 of the “Venationes Ferarum, Avium, Piscium” by Jan Van der Straet (Figure 1) shows an unusual employment of sponges documenting the use of sponges to collect naphtha from the surface of the waves [26], as suggested by the original text of the copper print: “Naphtha Bituminis est liquidi genus: in mare manta Montibus e Siculus, fluidisque supernatat undis: Spongia eam excipiunt Nautae, expressamque recundunt Ollis, ut varios hominum seruetur in usus”. As for the locality where this activity was performed, some Antiquarian Print sellers (e.g. Philographikon Gallery, Rottenbuch, Germany) suggest, with caution, the northern coast of Sicily. Commercial sponges covered a wide range of uses. They were traditionally used in surgery, since the Middle Age. The technique to administer inhalation anesthesia by a “soporific sponge” is purely Arabic, and it has been commonly used in the time of the surgeon, Ibn Al Koff (1232-1286). The “soporific sponge” imbibed of hashish, papaver, and hyocymine juice, and then dried under the sun, was humidified again when called upon, and placed at the patient's nose (Figure 1) to favor the release and absorption of the anesthetic by the mucus membranes to induce deep sleep and relief of surgical pains [27][28]. In Europe, at the end of the 12th century, Nicholas of Salerno mentioned the inhalation of vapors from the “soporific sponge”. The Dominican friar Theodoric of Cervia (1205-1298) (also known as Theodoric Borgognoni of Lucca) was the pioneer of anesthesia in 1267 at Bologna. Son of a surgeon of the crusaders (Hug of Lucca), Theodoric advocates the use of sponges soaked in a narcotic in its famous “Cyrurgia seu filia principis” published in 1498, as first edition, in Venice. The European prescription was quite different from the original Arabic one. A sponge had to be boiled within a mixture in a brass vessel with a specific proportion of opium, hemlock and the juice of mandragora, ivy and unripe mulberry, until all has been reduced and soaked up, and then applied to the nostrils of the patient. To wake the patient up again, after the surgery, a sponge full of vinegar should be applied to his nose. Arnold of Villanova (Arnaldus Villanovanus, ca. 1238-1310), an alchemist, astrologer and magician, reports that variations on the spongia somnifera were quite common in the 9th to 14th centuries [29], and were still used in the 19th century (Figure 1). In synthesis, sponges represent a very ancient Mediterranean cultural and historic heritage. The use of “bath sponges” firstly common among Greeks and Romans, was widely extended in the entire circum-Mediterranean area, and spread in the entire continental Europe during the Middle Age and the Renaissance. The Venice Republic was the ruler of sponge trades directed to the Oriental Roman Empire and the Sacred Roman Empire.
HARVESTING TOOLS The first statement on sponge fishery is from the ancient Greek poet Oppian (2th century B.C.)[30] who refers to the diet and life-style of spongers, and to their terror of sharks. Aristotle (350 B.C.), in his detailed morphological and ecological descriptions of sponges, reports also a peculiar “elephant nose-like tube” allowing sponge fishermen to breath during diving (The history of animals, I, 1; V, 16; VIII, 1; IX, 14, 32, 44; On the parts of animals, IV, 5) [31] [32]. Ancient sponge fishermen were skin-divers able to reach the seabed with the help of a lead pound and to light up the sea depths by spiting out an “ointment”; while a security rope monitored by the boat crew helped divers reach the sea surface [32]. The
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presence of particular fishes was considered as an indicator of shark absence, and consequently of suitable areas for sponge fishery. The traditional systems of sponge harvesting used by Greek spongers, almost exclusively native of the Symi and Kalymnos islands, remained practically unchanged up to the end of the 19th century (Figure 2). In 1866, the “scafandro”, “skafandra” or “macchina” (diving suit), an innovative device allowing to explore notable depths and to dive for a very long time (Figure 2), was introduced and highly appreciated by sponge traders and boat captains in Greece and Turkey to perform sponge harvesting [33]. The result was a conspicuous increase of catch, but also a dramatic incidence on divers‟ health and survival. The absolute ignorance of hyperbaric dangers determined the death by embolism of dozen of fishermen. A high-school retired professor, Charles Flegel, begun and won a real “crusade” to obtain the prohibition of this diving technique in the whole Mediterranean as the result of an international agreement in 1902-1904 [34] [35].
Figure 2. Sponge fishery systems. a) skin diver from a small boat (Dodecanese, Greece); b) the “gangava”, a sponge harvesting device dragged on the sea bottom from a boat in deep water (Kalymnos Island, Greece); c) shallow water sponges collected by means of a long harpoon (kamakys) and a “glass-bottom-bucket” (Tunisian coasts, North Africa); d) hard heat sponge-diver and its crew (Kalymnos Island, Greece).
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According with the fear of the new diving suit, traditional sponge-skin-divers represented the majority of workers in sponge fishery up to the second half of the 20th century. Spongeskin-divers performances went down in history. The dive to a depth of 77m (August 1913), at the Skarpanto Island by Gheorghios Haggi Satti, a frail Greek sponge fisherman, to pick up the loose ship anchorage, with the only help of a flat stone as weighting device (15 Kg) is registered in the Captain‟s logbook of the Italian warship Regina Margherita. In 1912, Maurice Fernez created the namesake underwater breathing apparatus representing a real innovation for sponge divers [36]. The “fernez” became instantaneously very popular, and, around 1930, was practically the single method used for sponge harvesting (Figure 2). At the same time, the knowledge on hyperbaric physiology was in a rapid phase of progress and the safety of divers became acceptable. Just before the second World War, over three hours of diving was the time scheduled for a typical working day of a sponge fisherman, with a potential collection of about 100 sponge specimens 20-40 cm in diameter. After World War II, this technique evolved in the “narghilè” that is still used at present with few variations from the original project of the French inventor. Diving for sponges was not the only method of harvesting. In very ancient times, coastal Mediterranean populations collected naked skeletons directly from the beach. This is still possible today to successfully collect conspicuous quantities of bath sponges while walking on the beach in some pristine coastal localities (e.g. Sardinia, Corsica, Tunisia). Another v common method was the catch from a boat, by means of primitive tools (Figure 2) such as a “specchio” (a glass plate which forms the bottom of a wooden bucket) and a long harpoon [37]. Along the Dalmatian coast, this system is not yet dismissed, and fishermen scrutinize the shallow water seabed in the same traditional way of the past as reported by the first written document on sponge diving in Croatia (pers. com. Lucio Pesle, Spugnificio Rosenfeld, Trieste). Another harvesting technique, still in use on a few boats, but nearly disappeared, is the “gangava” a particular combination between a net and a dredge (Figure 2) allowing to exploit sponge banks at 100-200 m of depth [37].
HARVESTING EFFORT AND TRADE During the Ancient and Middle Ages, a very low harvesting effort had been exerted to exploit sponges, whereas the increasing of market demand during the 19th century expanded to excess the number of vessels and fishermen working on sponges [38]. Just to give some examples, about 850 vessels and over 2000 workers exploited the sponge banks along the coast from Tunis to Zarziz in the winter 1882 [37] [39], and the exploitation of Tripolitania and Cyrenaica coasts from 1860 to 1890 was performed by 100200 vessels per season with about 2000 workers [40]. In 1878, a single fisherman with “scafandro” was able to harvest 25,000-35,000 sponge specimens at Bengasi [40]. The latter are clear cases of over-fishing. As a consequence of over exploitation of Mediterranean sponge banks and demand increase, the industrial fishery of the sponge banks in the Caribbean Sea and Mexican Gulf started in 1841 [41] [42] [43] [44] [45] [46] [47]. In the 19th century, Paris became the center of an increasing sponge commerce; at the same time, many hundreds of vessels and thousands of fishermen and divers, from Italy, Malta, Greece, Turkey and Tunisia, harvested sponges from natural banks without any
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control, supplying the continuous increasing of a world-wide demand. If the sponges subjected to commercial exchange in the Ancient and Middle Age were probably harvested along the Dalmatian and Aegean coasts, successively new sponge banks were discovered in northern Africa, around 1840 [1]. In the same period, the importation of commercial sponges from the Caribbean begun (Florida, Bahamas, Cuba, Hispaniola, Yucatan, Belize, Honduras, Colombia) and commercial exploitation of bath sponges was also developed in Indo-Pacific areas such as the Philippines, Eastern Chinese Sea, a few Pacific Islands, Madagascar and the Red Sea [48] [49]. Several “sponge factories” (Figures 3, 4) were established from 1870 to 1890 in Greece (Vuvalis, Kalymnos), Austria (Echkel, Rosenfeld and Vidich, Trieste), France (Colombel freres and Devismas, Weill and Cypréos, Paris), Italy (Banco di Roma, Rome; Zafferoni, Milan), Sweden (Theodoridis and co., Stockholm), Great Britain (Petrides Bro.s, London), and Belgium (Labre, Liege).
Figure 3. Sponge trade. a) an Italian sponge factory (Rosenfeld, Trieste) with sponges selected for dimensions and quality; b) Mr. M.G. Weill with factory personnel (Rue des Francs-Bourgeois, Paris); c) the docks for sponge storage in the Tarpon Spring harbor (Florida); d) open air sponge drying at Batabano (Cuba); e) sponge packaging for shipping to Europe from Bahamas; f) sponges are trampled over and soaked in the sea just after harvesting to release soft skeleton from living tissues (Dalmatia, Adriatic Sea). All images are from the second half of 19th century when a huge bath sponge amount was harvested and traded from the Caribbean and Mediterranean areas.
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Figure 4. Road sponge vendors, during the 1950s, in France (a) and Italy (b).
Ascertained data on sponge trade trends date back to the second half of the 19th and the beginning of the 20th centuries. In 1893, over 2,000.000 sponge specimens were collected in Libya, and the fishing campaign produced over 30 tons of sponges from the same area in 1902. Statistic data of the Italian colonial government refer that a fishing effort of over 30 tons per year was considered the rule for more then a decade, up to 1911 [40]. It is problematic to outline at present a general overview on sponge trade because of the lack or unreliability of data [50]. Around 1930, the sponge trade reached about 100 tons per year in both Great Britain and Germany [49], while this value was 30 tons in Italy, Holland and France. In the same period, each citizen in Holland, consumed over 9 grams of sponges per year [49]; it means that a sponge 20 cm in diameter was completely consummated in 12 months needing to be replaced. As for the Greek export, a negative trend was reported from 1964 to 1978, with a decrease from about 115 to about 70 tons/year, on a period of 15 years [51]. Few data is available about general trends. A worldwide increasing of trades is stated in the period 1980-1985 (190 tons in 1980, up to 200 tons in 1985), while a decrease is reported for the following 9 years (145 tons/year in 1989) [52]. In 1991, a similar trend was outlined, but data is 20-30 % higher [53]. Starting from 1920-1930, the “sponge disease” severely affected natural sponge populations driving in the brink of extinction this millennial natural resource; time after time the epidemics became more frequent and merciless. The year 1985 dates the first Mediterranean sponge-disease scientifically investigated [54] [55] [56] [57] [58]. Around 1990, the sponge stock decline was dramatic, driving the sponge fishery to zero. In Greece, Turkey, Cyprus, Syria, Egypt and Tunisia, sponge harvesting was stopped for almost two harvesting seasons [59] [60]. In the last decade, Mediterranean sponge fishery fluctuates around 50 tons/year [50] with Tunisia, Greece, and Libya as major “sponge-producers”. In summary, only scattered and local data is available up to the 1930s. From 1927 to 1932, a mean of about 350 tons/year is reported by FAO for the Mediterranean Sea, while about 100 tons was the average catch per year in the Caribbean area; successively, the same source reports a drastic reduction: annual average of about 120 and 65 tons for Mediterranean and Caribbean, respectively, from 1977 to 1986 (http://www.fao.org/docrep/field/003/ AC286E/AC286E02.htm).
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OVERFISHING AND DISEASE Data clearly indicates a very long over-exploitation of Mediterranean commercial sponge beds (Figure 5). The beginning of the over-fishing phenomenon matches the long-term employ of “scafandro” that determined a critical status, in less than 50 years, of the majority of Mediterranean sponge banks. Unfortunately, this “suggestion” is not supported by scientific data; where and when available, the trends in the market indicates an increasing of Mediterranean sponge commerce up to a maximum of 250 tons in 1985 [52], and a successive decrease with no possibilities of harvesting because of the devastating epidemic between 1985 and 1989 [60] [61] (Figure 5). Over-fishing in the Mediterranean Sea is reported only by few long-term studies. Local fishermen regularly exploited several small sponge banks along the eastern Ionian coasts. The values of population density ranged from ca. 20 ind/100 m2 (1970) to 10 ind/100 m2 (1980) and 5 ind/100 m2 (1990) with a mean diameter ranging from 5 cm to 12 cm and 7 cm, respectively [62].
Figure 5. Sponge trade trends (tons/year). a) the decade 1980-90 coincided with a heavy sponge disease in the Mediterranean Sea. Since 1985, a dramatic decrease of commercial exchanges is reported, from ca. 250 to 150 tons/year; b) on the long-term, a seventy year-long persisting negative trend is evident (dotted lines indicate not available landing data). At present, the Mediterranean sponge catch is about 1/7 of that of 1935: 50 vs. 350 tons/year.
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The synergy of over-fishing and disease drastically reduced the sponge beds exploitation to zero. During the occurrence of the disease, sponges are firstly covered by a “thin white layer” of pathogenic micro-organisms; in a second phase, collagen and cells disappear from the mesohyl; finally, spongin fibers are also attacked by excavating bacteria with the result of an extreme fragility of the sponge skeleton [56]. The epidemic was absolutely severe and almost all specimens of Spongia officinalis (Linnaeus, 1759) apparently died in some studied populations. At the Portofino promontory, over 60% of specimens involved in the epidemic were, however, able to recover starting from a small healthy remaining fragment of sponge [63]. At the population level, sponge recovery has been reported as very slow. The population of S. officinalis apparently totally extinct from the “Stagnone” of Marsala in 1987, reappeared only five years later after a process of re-colonization, while a total recovery took about ten years to reach a density similar to that detected in 1986; that notwithstanding, after ten years of observations, the average size of re-colonizing specimens was less than a half of the value measured before the disease [64]. Other historically exploited banks were recently focused in the Aegean Sea by [65] [11], highlighting the long-term recovering of sponge populations (e.g. Crete, Dodecanese). All of this data matches those collected along the Mediterranean coast of Egypt [66] where young individuals were abundant in shallow water after the disease. At the end of the summer of 1999, an extreme sea water warming stressed many species of sponges and sea fans evidencing “local extinctions” in the Mediterranean Sea; during this catastrophic epidemic not only Spongia officinalis and Hippospongia communis resulted affected, but also other massive species, such as Cacospongia scalaris (Schmidt, 1862) and Petrosia ficiformis (Poiret, 1789) were decimated in a wide geographic area along the coasts of Liguria, Provence, Tuscany, Corsica and Sardinia [67] [68] [69] [70] [71]. Unfortunately, the so-called “sponge-disease” became a common event in the successive years and sponge fishermen and traders are constantly in apprehension for that. A recent further alarm is coming from North Africa (Egidio Bellini Spugnificio, Cogozzo di Viadana, Mantova, Italy, pers. com.) and western Sardinia (R. Manconi, pers. obs.) where bath sponges have been reported as suffering and dying, although not massively, in the autumn and winter 2006-2007. The sponge-disease is not only a Mediterranean phenomenon [72]. The first sponge mass mortality was reported for the West Indies in 1938-39 [73] [74]. As a result of this and successive feral events, the species Hippospongia gossypina (Duchassaing and Michelotti, 1864) resulted locally extinct from the Florida coasts, while other commercial sponge species are disappearing from Cuban waters [45] [75]. Along over five thousand years of manhold, only a few Mediterranean sponge species entered the human common use. Other species, particularly from the West Indies, with quite similar properties of softness and absorption, begun to be commercially exploited around 1840. At present, bath sponges represent a “niche product” supplied only to “selected consumers”. The price, especially during the last critical years, has enormously increased. In specialized shops, a good Mediterranean “Zimocca”, “Fine-Levantine” or “Horse-Sponge” has a retail price of over 30 Euro for a specimen 20-25 cm in diameter. Large specimens of Hippospongia communis can overcome 100 Euro. Detail market price of Mediterranean sponges is calculated around 500-1000 Euro/Kg dry weight according to the quality level (authors‟ pers. obs.).
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DRUGS IN THE PAST AND PRESENT Sponges were, and still are, used as ethno-medicine. New Zealand natives to aid healing, have long used Halichondria moorei (Bergquist, 1961). Recent chemical analyses of the sponge discovered that nearly 10% of the sponge weight is composed of the potent antiinflammatory drug: potassium fluorosilicate [76]. Even though the Maori did not know what the compound itself was, they quickly learned that the use of this sponge had amazing ability to reduce painful swelling on injuries. After 1606, in Europe, some sponge species, called Tethyae, were effective as digestive or against bad breath, dysentery, flatulence, and sciatica [18]. In traditional medicine and homeopathy, sponges have been commonly administered as “spongia tosta”, “spongia usta” and “carbo-spongiae” [77] [78] [49]; sponges were also used in a few cases to realize small esthetical prosthesis [79]. In any case, these biomedical uses, still practiced, never involved considerable quantities of sponges. This has been the scenario up to date. In the last years, many other species belonging to the phylum Porifera have been tested for the potential production of natural compounds of biological interest [80]. Recently it was proved that sponge compounds are the most active and useful among marine natural products [81] [82] [83] [84] [85] [86] [87] [88] [89] [90] [91] [92] [93] [94] [95] [96] [97] [98] [99]. Some of these compounds have been listed in catalogues of chemicals for many years and prices are so high that the species from which they are extracted must be inserted in an emended list of “commercial sponges”. New drugs, based on sponge extracts, are under testing or have been recently introduced in the market and it is necessary to focus on the rapid harvesting increasing of large quantities of sponges (hundreds kilos or tons). The Mediterranean Dysidea avara (Schmidt, 1862) (Dictyoceratida, Dysideidae) is in danger because of a pressing demand for industrial extraction of chemicals. Avarol, a novel hydroquinone, was extracted from this species [100] and its anti-inflammatory properties were successfully proved [101]. After that, many authors focused their interest on the biological effects of this compound [81] [82] [102] [103] [104] [105]. In particular, the antipsoriatic activity opened its use in pharmacological industry and Avarol represents one of the first sponge isolated compounds sold in the new market of natural products of marine origin. Avarol can be chemically synthesized [106], however the combination of a number of expensive substrates, a low maximum overall yield of 2 %, and more than 20 reaction steps contributed to not assess the feasibility of its chemical synthesis [97] [98]. The Mediterranean natural sponge population of Dysidea avara has been widely harvested for clinical trials [97][98]. Unfortunately, the prevision of previous authors that: “an amount of 75 tons of sponge would be needed every year to sustain the market demand” is absolutely underestimated. The suggested percentage of Avarol in ointments produced at present is 10% (S. De Rosa, pers. com.), while the percentage of Avarol in D. avara specimens is about 5% [107]. Around 2% of the European (over 700,000,000 of persons) and North American (over 500,000,000 of persons) populations are subjected to this multi-hereditary skin degenerating disease; as a consequence, around 24 millions of persons need continuous care. Just to give an example, 2.5 billion of US $ per year is the social cost of psoriasis in the United States [108].
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Depending on the body extension of psoriatic areas, 50 grams of ointment mixture can endure no more than 1 month, requiring over 1 Kg per year for each patient. Considering a scenario where about 10% of the entire market could be covered by the “Avarol-ointment”, a harvesting effort of 240 tons per year is presumably expected for D. avara, which is in absence of any data on distribution and population densities. The possibility that some other allied species could also produce Avarol does not reduce the potential environmental danger and tremendous risk for biodiversity.
BIORESOURCE, BIODIVERSITY ASSESSMENT AND HISTORICAL TREND IN TAXONOMY The Mediterranean “bath sponges” are represented by 5 species belonging to the genera Spongia (Linnaeus, 1759), and Hippospongia (Schulze, 1879) namely Spongia lamella (Schulze, 1879), Spongia mollissima (Schmidt, 1862), Spongia officinalis (Linnaeus, 1759), Spongia zimocca (Schmidt, 1862) (Hippospongia communis) (Figure 7; Tables I, II). They are characterised by an extremely plastic collagenous skeletal network of fibers, made mainly by spongin, that is commonly used as a multipurpose tool for its properties of elasticity, tenacity, softness and capacity to absorb large quantities of water. This natural proteic skeleton was mimicked and reproduced as synthetic sponge after the industrial revolution of plastic materials.
Figure 6. Spongia officinalis. a) a healthy sponge in situ showing a small whitish area of re-growth after disease; b) a seriously damaged sponge in situ with almost bared skeleton; c, d) healthy skeletal network by scanning electron microscopy (SEM); e) a deeply degraded fiber looses elasticity and strength.
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Figure 7. Bath sponges (dry skeletons) belonging to the five Mediterranean species of Spongidae. a) Spongia mollissima (fine-levantine); b) Hippospongia communis (horse-sponge); c) Spongia lamella (elephant-hear); d) Spongia zimocca (zimocca-sponge); e) Spongia officinalis (fine-dalmatian). Scale bar = 10 cm.
Sponge banks bathymetric distribution is variable between species, from the sea surface to over 200-300 m of depth. Preferential habitat is on rocky bottoms, but other relevant records are linked to detritic or muddy bottoms, seagrass beds and submerged caves. As for the geographic distribution of bath sponges in the Mediterranean, scattered data is reported on a few maps [109] [49] [110]; other, more detailed maps show sponge banks along the Dalmatian and Libyan coasts [40] [111]. Commercial sponges are a key biological resource for a sustainable development of circum-Mediterranean populations along the coasts of Dalmatia, Greece, Aegean islands, Turkey, Cyprus, Syria, Egypt, Libya and Tunisia. Sponge banks of Sicily and Sardinia are still exploited by expert sponge divers i.e. by immigrants from North-African countries (R. Manconi, pers. obs.). The northern coasts of the Western Mediterranean basin (peninsular Italy, France and Spain) are generally considered not economically exploitable, notwithstanding the ascertained presence, in the past, of conspicuous populations of bath sponge species. The global taxonomic richness of Porifera characterized by an exclusive fibrous skeleton of collagen and spongin (horny sponges) previously comprised in the order Keratosa, is stated at the higher level after the last fundamental taxonomic revision (Hooper and van Soest, 2002) [112]. Systematics and phylogenetic relationships of sponges were only recently tested using current biochemical and molecular approaches, matching in part the classical morphological classification [113] [114] [115] [116] [117] [118] [119] [120] [121]. At the global level, the group of „horny sponges‟ consists of 56 genera and 11 families ascribed in 4 orders [122] [123] [124] [125] [126] [127] [128] [129] [130] [131] [132] although the species richness is at present in a dynamic status as we can speculate from taxonomic data of the World Porifera Database [133] and the Fauna d‟Italia [134]. At the regional level in the Mediterranean Sea, data available on horny sponges indicates that taxonomic richness is notably high with 57 species, 21 genera and 9 families ascribed in 4 orders [134].
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Roberto Pronzato, Fabio D. Ledda and Renata Manconi Table I. Taxonomic richness of horny sponges (Porifera: Demospongiae) in the Mediterranean Sea. Species of bath sponges are indicated by asterisks. Protected species (Bern Convention and Barcelona declaration) are in bold Orders (4) Dendroceratida Minchin, 1900; Dictyoceratida Minchin, 1900; Halisarcida Bergquist, 1996; Verongida Bergquist, 1978 Families (9) Aplysinidae Carter, 1875; Darwinellidae Merejkowsky, 1879; Dictyodendrillidae Bergquist, 1980; Dysideidae Gray, 1867; Halisarcidae Schmidt, 1862; Ianthellidae Hyatt, 1875; Irciniidae Gray, 1867; Spongiidae Gray, 1867; Thorectidae Bergquist, 1978) Genera (21) Acanthodendrilla Bergquist, 1995; Aplysilla Schulze, 1878; Aplysina Nardo, 1834; Cacospongia Schmidt, 1862; Chelonaplysilla Laubenfels, 1948; Collospongelia Ferrer Hernandez, 1923; Coscinoderma Carter, 1883; Darwinella Müller, 1865; Dendrilla Lendenfeld, 1883; Dysidea Johnston, 1842; Euryspongia Row, 1911; Fasciospongia Burton, 1934; Halisarca Johnston, 1842; Hexadella Topsent, 1905; Hyrtios Duchassaing and Michelotti, 1864; Hippospongia Schulze, 1879; Ircinia Nardo, 1833; Pleraplysilla Topsent, 1905; Sarcotragus Schmidt, 1862; Spongia Linnaeus, 1759; Spongionella Bowerbank, 1862 Species (57) Aplysilla rosea (Barrois, 1876); Aplysina aerophoba Nardo, 1843; Aplysina cavernicola (Vacelet, 1959); Acanthodendrilla levii Uriz and Maldonado, 2000; Cacospongia mollior Schmidt, 1862; Cacospongia proficiens Pulitzer-Finali and Pronzato, 1980; Cacospongia scalaris Schmidt, 1862; Chelonaplysilla erecta (Row, 1911); Chelonaplysilla noevus (Carter, 1876); Chelonaplysilla psammophila (Topsent, 1928); Collospongelia nicaeense (Pulitzer-Finali and Pronzato, 1980); Coscinoderma sporadense Voultsiadou-Koukoura, Soest and Koukouras, 1991; Darwinella australiensis Carter, 1885; Darwinella corneostellata Carter, 1872; Darwinella dalmatica Topsent, 1905; Darwinella gardineri Topsent, 1905; Darwinella intermedia Topsent, 1893; Dendrilla acantha Vacelet, 1958; Dendrilla cirsioides Topsent, 1893; Dysidea avara (Schmidt, 1862); Dysidea fragilis (Montagu, 1818); Dysidea incrustans (Schmidt, 1862); Dysidea perfistulata Pulitzer-Finali and Pronzato, 1980; Dysidea tupha (Martens, 1824); Euryspongia raouchensis Vacelet, Bitar, Carteron, Zibrowius and Perez, 2007; Fasciospongia cavernosa (Schmidt, 1862); Fasciospongia coerulea Vacelet, 1959; Halisarca dujardini Johnston, 1842; Halisarca sputum Topsent, 1893; Hexadella detritifera Topsent, 1913; Hexadella racovitzai Topsent, 1896; Hippospongia communis* (Lamarck, 1813); Hyrtios collectrix (Schulze, 1879); Hyrtios erectus (Keller, 1889); Ircinia chevreuxi (Topsent, 1894); Ircinia dendroides (Schmidt, 1862); Ircinia oros (Schmidt, 1864); Ircinia paucifilamentosa Vacelet, 1961; Ircinia retidermata Pulitzer-Finali and Pronzato, 1980; Ircinia variabilis (Schmidt, 1862); Ircinia vestibulata (Szymanski, 1904); Pleraplysilla minchini Topsent, 1905; Pleraplysilla spinifera (Schulze, 1879); Sarcotragus fasciculatus (Schmidt, 1862); Sarcotragus foetidus Schmidt, 1862; Sarcotragus pipetta (Schmidt, 1868); Sarcotragus spinosulus Schmidt, 1862; Spongia lamella* (Schulze, 1879); Spongia mollissima* Schmidt, 1862; Spongia nitens (Schmidt, 1862); Spongia officinalis* Linnaeus, 1759; Spongia virgultosa Schmidt, 1868; Spongia zimocca* Schmidt, 1862; Spongionella depressa (Topsent, 1928); Spongionella gracilis (Vosmaer, 1883); Spongionella pulchella (Sowerby, 1806); Spongionella ramodigitata (Topsent, 1901).
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Table II. Records of endemic and rare species of Mediterranean horny sponges Endemic Mediterranean species (8) recorded in several localities Dysidea incrustans (Schmidt, 1862) Hippospongia communis (Lamarck, 1813) Ircinia oros (Schmidt, 1864) Ircinia retidermata Pulitzer-Finali and Pronzato, 1980 Sarcotragus pipetta (Schmidt, 1868) Spongia lamella (Schulze, 1879) Spongia mollissima Schmidt, 1862 Spongia zimocca Schmidt, 1862 Endemic Mediterranean species (10) recorded in a single locality Acanthodendrilla levii Uriz and Maldonado, 2000 Cacospongia proficiens Pulitzer-Finali and Pronzato, 1980 Collospongelia nicaeense (Pulitzer-Finali and Pronzato, 1980) Coscinoderma sporadense Voultsiadou-Koukoura, van Soest and Koukouras, 1991 Dysidea perfistulata Pulitzer-Finali and Pronzato, 1980 Euryspongia raouchensis Vacelet, Bitar, Carteron, Zibrowius and Perez, 2007 Fasciospongia coerulea Vacelet, 1959 Ircinia chevreuxi (Topsent, 1894) Ircinia paucifilamentosa Vacelet, 1961 Ircinia vestibulata (Szymanski, 1904) Not endemic species (3) recorded in a single Mediterranean locality Darwinella corneostellata (Carter, 1872) Darwinella gardineri Topsent, 1905 Spongionella ramodigitata (Topsent, 1901) Lessepsian migrant species (1) recorded in a single Mediterranean locality Hyrtios erectus (Keller, 1889)
The Mediterranean area represents a hot-spot of biodiversity [135] [136]. For the implementation of a correct management and conservation of the biological resources, more detailed studies on the status of its marine fauna, particularly at the level of endemic taxa and exploited species, are needed. The Mediterranean sponge fauna characterized by a high endemicity value (ca. 40 %) comprises over 600 species with only a few Lessepsian immigrant (not more than 5 %). Faunal affinities range from 20% with the North Atlantic and 6% with the South Atlantic [137] [138] [139] [140] [141] [142] [143] [144]. Only Caribbean-CentroAmerican (640 spp), Sino-Japanese (589 spp) and Indonesian (965 spp) areas harbor a sponge fauna with higher levels of species richness [145] [146]. From the historical point of view, the discovery of new taxa of horny sponges in the Mediterranean Sea showed a continuous and constant increase up to our days with the golden period between 1860 and 1930 [140]. The taxonomical approach was poorly valued in recent times [147] [148] [124] [125] [110]. First data on Mediterranean horny sponges preserved in the Linnean herbarium is reported in the 13th edition of Systema Naturae [149]. Starting from the discovery of Spongia officinalis, more or less 20 authors are involved from 1759 to 2010 in the description of new species with a mean number of 2-3 new species per decade. Out of that trend is the period in
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which a large amount of new species descriptions and faunal inventories were published by Emile Topsent (12 species from 1892 to 1938), Oscar Schmidt (16 species, most in 1862) and Franz Eilhard Schulze (Figure 8). Most Mediterranean horny sponge species (around 70% of the total) were described more than one hundred years ago; many nominal species could represent simply “names written on paper” (i.e. type specimens are lost; no further findings known except the first one). Most historical collections have been forgotten, for many decades, in several natural history museums, without further studies supporting the axiom „research on systematics does not pay in terms of financings and career‟. In a few words, the real number of sponge species of the Mediterranean Sea (not only horny) is unknown and, probably, it is highly over- or underestimated. At present, the Mediterranean horny sponge fauna is characterized by high levels of taxonomic richness with 18 endemic species and an endemicity value of 31.6% [124] [125] [137] [138] [139] [140] [142] [143] [134].
Figure 8. The majority of Mediterranean horny sponge species were described before the ending of XIX century and the last new species was described in the year 2000. a) trend in species richness; b) new species described per year.
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NEW TRENDS IN SUSTAINABLE MANAGEMENT AND INTEGRATED CONSERVATION Among Mediterranean horny sponges, attention must be paid to the endemic species of economic interest for which a correct management and a rational exploitation plan appear to be necessary for the conservation of biodiversity and the maintenance of resources availability and ecosystem homeostasis. The Mediterranean sponge fauna is protected by both local and international protocols [140] [134]. Most species under safeguard are common, widespread and well known. They are characterized by peculiar species-specific macrotraits and conspicuous body dimensions that render them easy to recognize. On the contrary, other species considered rare or restricted to a very small geographic area are not included in these directories (see Tables I, II). Sponge species are occasionally inserted in “Red Lists” or indicated as being at risk of extinction [140]. Reliable data on sponge harvesting and trade is referred at present exclusively to bath sponges. In the Mediterranean Sea, only 15 species are under observation, but no harvesting rules or restrictions are indicated except for the minimum size of traditional “bath sponges” species [150] [140] although the continue harvesting severely affect wild populations. Sponge harvesting and trade must be regulated at the global scale for sustainability and social equity both to enforce existing species protection laws and to establish legal frameworks to monitor and manage this bioresource. Although conservation plans are active in other areas of the world in order to avoid impoverishment or local extinction of bath sponge banks [46] [47], no real active protection has been performed, until our days, in the Mediterranean Sea. The trend in the field of the pharmacological approach to sponge metabolites is clearly ascending [96] and many hundreds of species are involved in a possible massive exploitation. Horny sponges represent a conspicuous source of bioactive compounds e.g. aeroplysinin1 from Aplysina aerophoba (Nardo, 1833) and A. cavernicola (Vacelet, 1959) and are among the most suitable Porifera species for sponge culture [151], as highlighted by the tradition and success in culturing bath sponges [110] [152]. The increasing awareness in the last decades on the value of marine biological resources in general, and on marine sponges in particular, as producers of novel drugs matches well the need to adopt aquaculture technologies for a sustainable exploitation of these resources. Sponge culture represents a reliable and simple technique to obtain conspicuous biomass of sponges for commercial purpose, mimicking a simple marine model systems that could benefit the development of medicine and industry. The ability of sponges to recover damaged parts of their body is well known since ancient times and Aristotle's “Historia animalium” [32] reports that “when the sponge is broken off, it grows again from the remaining stump and the place is soon as well covered as before”. Sponge-culture also is not a recent idea. The Italian scientist Filippo Cavolini [153], strongly supporting the animal nature of sponges, was the first (1785-1790) to test the experimental fragmentation and the recovery process in sponges as reported also by Marenzeller [154]. At the end of the last century, Eduard Oskar Schmidt awarded of the over-exploitation problem, tried to farm commercial sponges, testing the self-regeneration properties of sponge
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fragments. Buccich and Schmidt performed the first attempt of sponge cultivation between 1863 and 1872 [155] [156] when the Austrian Government financially supported the sponge farming experiment in front of Hwar Island (Croatia). As a result, they obtained specimens of about 30 cm in diameter in 3 years, with a mortality rate of 10 % using sponge fragments fixed onto wood boxes [157] (Figure 9). In 1907, Wilson tested bath sponge growth from cell aggregates that failed at an initial development stage [157] (Figure 9). In 1911, Dubois tested a new method of sponge culture, starting from larvae but he failed for the difficulties encountered in collecting larvae of other sponge species, more competitive than the commercial ones [158]. Although a general necessity of sponge farming was strongly suggested [159], a long period of inactivity occurred after the promising experiments performed between the end of the 19th century and the beginning of the 20th century.
Figure 9. Sponge farming. a) The first modern sponge farming method was introduced by Moore in Florida (1910); b) as for the Mediterranean area Verdenal and Vacelet (1990) obtained promising results; c) a modular plant system for spongeculture (USAMA ®) proposed by Pronzato et al. (2006).
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More recently, sponge transplantation to different conditions of light and current was tested together plants for sponge farming in front of Marseille (France) to study relationships between growth and mortality rates and environmental factors [160] [51]. These experiments played the role of promoters for a colossal revival of sponge farming in the western Mediterranean Sea by testing successfully several sponge species [161] [162] [163] [70] [164] [165] [166] [67] [167] [168] [169] [170] [171] [172] [173] [174]. Sponge farming is a simple and eco-sustainable solution to exploit the sponge performances by mimicking the natural processes of clonal proliferation by fragmentation (Figure 9). These technologies are easy and not expensive [164] [166], and this productive activity is possible in combination with all other mariculture approaches, e.g. fish farming in floating cages [165] and in tourist harbors [166] [172]. In addition, sponges are very efficient natural water cleaners and they can filter an enormous water volume, performing very high cleaning rates [175] [176] by feeding on bacteria, organic suspended particles and dissolved organic matter. Production of conspicuous amounts of sponge biomass is possible in situ for some target species [177] and clearance rates values calculated are very high both in vitro [178] [171] [179] [180] and in situ conditions [170] [172]. A combination between sponge farming and sources of organic pollution can reduce environmental damages in coastal areas [165] [140]. The experimental use of sponge farming, as bio-remediation device, lasting over twenty years was tested under different environmental conditions [181] [182] [165] [75] [172]. The technical equipment and material necessary to start a sponge farm is simple and cheap and satisfactory results are the rule. The plant structures are modular and adaptable to all sea beds and also in very shallow water. Three or four days after transplantation, sponge fragments regenerate their outer layer, after one week, pigmentation on the surface is restored and they tend to assume a rounded shape after two months. A critical phase with mortality lower than 10% occurs during the first days [67] [162] [163] [70]. Moreover, the natural reproductive cycle of sponges by sexual or asexual modes with production of swimming larvae or buds, respectively, seems to be the rule in farming conditions [166]. Sponge mariculture could avoid the risk of local population extinction due to the harvesting effort on already stressed natural sponge banks, whose density in the whole Mediterranean Sea is almost unknown. This low-cost and productive approach together with the reduction of harvesting efforts on the wild populations can promote and successfully support conservation, when resources are one of the main constraints in conservation plan management, by suitable tools to minimize the risks of disappearance for these endangered species. Sponge culture matches well the strategies for the development of marine biotechnologies. Indeed, sponge rearing represents a suitable “blue biotechnology” that allow species diversification increase for aquaculture practices thus supporting the improve of marine based economy in a sustainable way, both for rational economy and biodiversity conservation. As a benefit for the protection of these endangered species in order to continue to be considered available and renewable resources, we are supporting a circum-Mediterranean strategic plan as a network of several, scattered, small sponge culture plants along the coasts of southern Europe, northern Africa and Near East, both within and outside of harvested sponge banks as a source propagules to re-colonize exploited wild sponge populations. Extensive sponge culture could be very effective to increase re-population and conservation
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potentialities of target species promoting their both persistence and dispersal of populations by enhancing connectivity by panmixis [140] [110] and the production of asexual propagules [166] . The growing interest by chemical, pharmacological and cosmetic industries in sponge compounds is opening new fields of application for sponge farming in the framework of a coastal areas integrated development under a sustainable strategy of bioresources management. An efficient strategic plan focusing on this threatened and high value resource is a key issue at the local, regional and global level especially in areas constrained by limited resources with a good potential contribution to social benefit and to promote a sustainable productive activity by coastal populations (i.e. north Africa and Near East countries). On the other hand, the risk of a “money-business” approach of the applied research on sponges is real. At present, young scientists are in strong competition, and a hurried overevaluation of results is possible. Day-by-day, year-by-year, the number of studies on life history and, particularly, on biogeography, faunistics and taxonomy of sponges, as in general for all phyla, are dramatically decreasing, while those on “sponge biotechnology” are vice versa dramatically increasing [96]. This evidence highlights the pressing of events in the applied field. To slow-down and settle the knowledge would be a suitable approach for a general overview on the future of sponge sustainable management in a general global trend of integrating models for natural resources management.
ACKNOWLEDGMENTS Research supported by the Italian Ministero dell‟Università e della Ricerca Scientifica e Tecnologica (MIUR-PRIN „L‟endemismo nella fauna italiana: dalla conoscenza sistematica e biogeografica alla conservazione‟), EU-7FP Project BAMMBO (Biologically Active Molecules of Marine Based Origin, contract n. 265896), and Fondazione Banco di Sardegna. F.D. Ledda was supported in part by a grant from RAS (“Promozione della Ricerca scientifica e dell‟innovazione tecnologica in Sardegna”, PO/FSE/Sardegna2007/13, L.R.7/2007, CRP1_324).
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In: Endangered Species Editor: Manuel Esteban Lucas-Borja
ISBN: 978-1-62257-532-9 © 2012 Nova Science Publishers, Inc.
Chapter 6
ENDANGERED MEXICAN FISH UNDER SPECIAL PROTECTION: DIAGNOSIS OF HABITAT FRAGMENTATION, PROTECTION, AND FUTURE – A REVIEW Ricardo Dzul-Caamal, Hugo F. Olivares-Rubio, Cynthia G. Medina-Segura and Armando Vega-López Laboratorio de Toxicología Ambiental, Escuela Nacional de Ciencias Biológicas, Instituto Politécnico Nacional, Unidad Profesional Zacatenco, México D.F. México
ABSTRACT Mexico is a megadiverse country and interior water fish are a clear example of high endemism and specialization. These phenomena are the result of diverse geographic and historic factors that have brought about the spatial isolation of catchment basins. The latter are usually in areas of high climate diversity and include transitional zones between the Nearctic and Neotropical regions. The exact number of fish species endemic to Mexico is unknown since many have probably disappeared before being identified. A considerable number of them are believed to have been lost during the second half of the 20th century. In all of Mexico the microendemism of the fish species creates the enormous diversity but in the same way is the main risk factor associated with the destruction of the habitat. The official Mexican norm currently lists 204 fish species considered endangered, threatened, under special protection or probably extinct in their natural habitat, distributed among 16 orders and 33 families. The reasons behind population reductions in these species are diverse. Intrinsic factors such as low fecundity and highly specialized courting behavior can be cited, but extrinsic factors have probably been fundamental in these reductions, including habitat loss due to tributary diversion or discharges of contaminants and toxicants, construction of hydraulic infrastructure, exotic species
Corresponding author: Tel: (+52 55) 57296300x52320. Fax (+52 55) 57296300x52301, E-mail address:
[email protected] (A. Vega-López).
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Ricardo Dzul-Caamal, Hugo F. Olivares-Rubio, Cynthia G. Medina-Segura et al. introduction, and human pressure in the form of commercial and subsistence fishing. However, the problem is very complex and in the same manner the knowledge on the ichthyofauna in the areas of the country is different, but can nevertheless be subdivided by geographic regions of the country. In the North of Mexico has a more in-depth knowledge on the fishes in such a way that the number of extinct or endangered species is twofold greater with respect to the center and south of the country. The main cause of risk is the destruction of habitat since it does not reach the amount of water through international rivers to maintain biodiversity coupled with desertification and competition from exotic species. In Central Mexico species diversity was high, but human pressure and habitat degradation together with high specialization and fishing for subsistence and/or for sale have been fundamental in its disappearance. In the South of Mexico, although it is the area with slighter environmental impact a lack of knowledge about the native ichthyofauna prevails and hence, there is less of a number of species reported at risk. Firm evidence indicates that endemic fish species in Mexico have adapted to compensate for the effect of naturally occurring toxicants, but there is an absence of information on many other aspects associated with their disappearance. The present study reviews habitat, protection measures, and impact of human activities and water quality in relation to the state of conservation of these fish species.
INTRODUCTION Mexico is a megadiverse country, harboring almost 10% of the species known worldwide although it covers only 1.3% of the land surface of the planet [1]. It has a great wealth of fish: nearly 2,300 marine and freshwater species, but the survival of more than a dozen marine species and one third of its freshwater species is compromised [2]. With the aim of establishing a normative framework, the guidelines provided by the Convention on Biological Diversity, signed at Rio de Janeiro, Brazil, in 1992, have been adopted by the Mexican government in order to identify and monitor ecosystems and habitats harboring great biodiversity or a large number of endemic or endangered species, or wildlife in general. These regulations, set down in the Official Mexican Norm (NOM-059-SEMARNAT-2010) [3], include species probably extinct in the wild as well as endangered and threatened species and those under special protection. Each one of the latter concepts on the conservation status of species differs from the others, with stress being laid on the following: (i) Endangered species are those whose distribution range or population size has diminished drastically within Mexican territory, putting at risk their biological viability throughout their natural habitat as a result of factors such as drastic habitat modification or destruction, unsustainable exploitation, disease, or predation, among others. (ii) Threatened species are those that could be in danger of disappearing at short or mid term should factors that can adversely affect their viability continue to exist, producing habitat deterioration or modification, or directly reducing the size of their populations. Last of all, (iii) species under special protection are those at risk of becoming threatened due to factors adversely affecting their biological viability, there being an evident need to promote their recovery and conservation or those of populations of related species. The Official Mexican Norm lists 204 fish species considered to be endangered, threatened, under special protection, or probably extinct in their natural habitat, distributed among 16 different orders and 33 families.
Endangered Mexican Fish under Special Protection
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Of these, freshwater fish are the ones most affected, and the number of species included in each of the different risk categories is rapidly increasing. For instance, in 1961, 11 Mexican species were documented in danger of extinction and seven more as extinct or eradicated in the country. Four decades later, these figures had risen to 83 and 25 species, respectively [4]. Seeking to establish the criteria for assessment of extinction risk, the Official Mexican Norm defines a methodology of study applicable to different zoological groups, called the Method for Evaluation of Extinction Risk in Mexican Wildlife Species (known as MER in Spanish), which is based on four independent criteria [3]. The MER includes a numerical (point-based) hierarchy for each criterion. Each of these criteria is independent from the others, while their aggregate score constitutes the comprehensive assessment of extinction risk. A MER score between 10 and 11 indicates that the species is threatened, while a species with a score between 12 and 14 is considered endangered. The criteria to be considered include: (a) The extent of the geographic distribution of the species in Mexico, which takes into account four categories: Very restricted (4 MER pts.), applicable to microendemic species and mostly extralimital species with a limited distribution in Mexico (
