The Xingu Seed Network and mechanized direct seeding

Abstract

Direct seeding proved to cost less than planting seedlings (approximately US2000/ha,comparedwithUS2000/ha, compared with US5000/ha) and to be more practical, since seeds are easier to carry and to plant. To plant one hectare, approximately 60 kilos of seeds of native trees (200 000 seeds) are mixed with 100 000 seeds of annual and subperennial legumes and sand, in a mixture called muvuca. The legumes help to create a multilayer vegeta-
tion, reducing niches for invasive grasses. Their root systems can contribute to soil aeration and decompaction, enhancing water absorption. Their ability to fix nitrogen and their intense leaf fall contribute to enhancing nutrient cycling and soil fertility. However, if they grow too densely, they can shade out the tree seedlings, slowing tree growth. If this occurs, manual or chemical weeding or thinning will be necessary. Ninety-one of the native tree species planted have germinated and survived droughts of up to six months without irrigation. Tree populations of between 2500 and 32 250 trees/ha have been established on the reseeded areas. The campaign has contributed to the restoration of 2565 ha of riparian forest at 238 sites until 2014.

Full text

THE STATE
OF THE WORLD’S
FOREST GENETIC RESOURCES
THEMATIC STUDY
GENETIC
CONSIDERATIONS
IN ECOSYSTEM RESTORATION
USING NATIVE TREE SPECIES
THE STATE OF THE WORLD’S FOREST GENETIC RESOURCES – THEMATIC STUDY
GENETIC CONSIDERATIONS IN ECOSYSTEM RESTORATION USING NATIVE TREE SPECIES
There is renewed interest in the use of native tree species in ecosystem restoration
for their biodiversity benefits. Growing native tree species in production systems
(e.g. plantation forests and subsistence agriculture) can also ensure landscape
functionality and support for human livelihoods.
Achieving full benefits, however, requires consideration of genetic aspects that are
often neglected, such as suitability of germplasm to the site, quality and quantity of
the genetic pool used and regeneration potential. Understanding the extent and
nature of gene flow across fragmented agro-ecosystems is also crucial to successful
ecosystem restoration.
This study, prepared within the ambit of
The State of the World’s Forest Genetic Resources
,
reviews the role of genetic considerations in a wide range of ecosystem restoration
activities involving trees. It evaluates how different approaches take, or could take,
genetic aspects into account, thereby leading to the identification and selection of
the most appropriate methods.
The publication includes a review and syntheses of experience and results; an
analysis of successes and failures in various systems; and definitions of best
practices including genetic aspects. It also identifies knowledge gaps and needs for
further research and development efforts. Its findings, drawn from a range of
approaches, help to clarify the role of genetic diversity and will contribute to future
developments.
I3938E/1/07.14
ISBN 978-92-5-108469-4
9789251 0 8 4694

i
THE STATE OF THE WORLD’S
FOREST GENETIC RESOURCES –
THEMATIC STUDY
FOOD AND AGRICULTURE ORGANIZATION OF THE UNITED NATIONS
Rome, 2014
GENETIC CONSIDERATIONS
IN ECOSYSTEM RESTORATION
USING NATIVE TREE SPECIES
Editors
Michele Bozzano,1 Riina Jalonen,1 Evert Thomas,1 David Boshier,1,2
Leonardo Gallo,1,3 Stephen Cavers,4 Sándor Bordács,5 Paul Smith6 and
Judy Loo1
1 Bioversity International, Italy
2 Department of Plant Sciences, University of Oxford, United Kingdom
3 Unidad de Genética Ecológica y Mejoramiento Forestal, INTA Bariloche, Argentina
4 Centre for Ecology and Hydrology, Natural Environment Research Council, United
Kingdom
5 Central Agricultural Office, Department of Forest and Biomass Reproductive
Material, Hungary
6 Seed Conservation Department, Royal Botanic Gardens, Kew, United Kingdom

ii
Recommended citation:
Bozzano, M., Jalonen, R., Thomas, E., Boshier, D., Gallo, L., Cavers, S., Bordács, S., Smith, P. & Loo, J., eds.
2014. Genetic considerations in ecosystem restoration using native tree species. State of the World’s
Forest Genetic Resources – Thematic Study. Rome, FAO and Bioversity International.
Photo credits:
p. 47 A. Borovics
p. 69 Leonardo Gallo, Paula Marchelli
pp. 139-140 Nik Muhamad Majid and team members
p. 154 Mauro E. González
p. 158 Philip Ashmole
p. 162 Dannyel de Sá, Cassiano C. Marmet, Luciana Akemi Deluci
p. 163 Luciano Langmantel Eichholz (top photos), Osvaldo Luis de Sousa, Elin Rømo Grande
p. 170 Wilmer Toirac Arguelle
p. 171 Orlidia Hechavarria Kindelan
p. 197 Lewis Environmental Services Inc.
pp. 217-218, 220 Luis Gonzalo Moscoso Higuita
pp. 231-232 Fulvio Ducci
p. 234 Sándor Bordács, István Bach
p. 238 Jesús Vargas-Hernández
p. 239 Alfonso Aguirre
The designations employed and the presentation of material in this information
product do not imply the expression of any opinion whatsoever on the part of the
Food and Agriculture Organization of the United Nations (FAO) or of Bioversity
International concerning the legal status of any country, territory, city or area or of its
authorities, or concerning the delimitation of its frontiers or boundaries. The mention
of specific companies or products of manufacturers, whether or not these have been
patented, does not imply that these are or have been endorsed or recommended by
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The views expressed herein are those of the authors and do not necessarily represent
those of FAO or Bioversity International.
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iii
Foreword
One of the major and growing environmental challenges of the 21st century will be the
rehabilitation and restoration of forests and degraded lands. Notwithstanding the large-
scale restoration projects initiated in Africa and Asia as of the 1970s, the current level
of interest in forest and landscape restoration is more recent. With the adoption of the
strategic plan of the United Nations Convention on Biological Diversity for 2011-2020, a
strong new impetus has been given not only to halt degradation, but to reverse it. The
plan states that, by 2020, 15 percent of all degraded lands should be restored. This target
is consistent with the Bonn Challenge, which calls for restoring 150 million hectares of
degraded land by 2020.
Forests play a crucial part in resilient landscapes at multiple scales. Restoring forest
ecosystems is therefore a key strategy not only for tackling climate change, biodiversity
loss and desertification, but can also yield products and services that support local people’s
livelihoods.
Restoration is not only about planting trees. Its success requires careful planning, as
painfully demonstrated by numerous past restoration projects that have not attained
expected goals. Restoration practices must be based on scientific knowledge, particularly so
in these times of progressive climate change. The trees we plant today and other associated
measures for restoration and rehabilitation of degraded ecosystems must be able to
survive abiotic and biotic pressures, including social ones, in order to be self-sustaining
and generate the products and services vital to supporting the world’s population and
environment for the years to come.
Biodiversity International coordinated this thematic study as an input to FAO’s landmark
report on The State of the World’s Forest Genetic Resources. The report was requested by the
Commission on Genetic Resources for Food and Agriculture, which guided its preparation,
and agreed, in response to its findings, on strategic priorities which the FAO Conference
adopted in June 2013 as the Global Plan of Action for the Conservation, Sustainable Use
and Development of Forest Genetic Resources.
The publication of this study is an important step in the implementation of the Global
Plan of Action. It provides fundamental information for the achievement of knowledge-
based ecosystem restoration using native tree species. It draws attention to the importance
of embedding genetic considerations in restoration activities, an aspect which is often
overlooked both by restoration scientists and practitioners, but is nonetheless crucial to
rebuilding resilient landscapes and ecosystems. We trust that it will contribute to informing
future restoration efforts and help to ensure their success.
Eduardo Rojas-Briales Stephan Weise
Assistant Director-General, Forestry Department Deputy Director General – Research
Food and Agriculture Organization of the United Nations Bioversity International

iv
Acknowledgements
We would like to express our gratitude to the scientists who contributed to the writing
of the scientific overviews presented in Part 2 of this thematic study. We would also like
to thank all of the practitioners who shared the experiences collected in Part 3, and who
completed the survey, which allowed us to undertake the analysis (Part 4) and to derive the
conclusions and recommendations (Part 5) of this study
The text was edited by Paul J.H. Neate, who was very helpful in standardizing and
simplifying the language. Gérard Prosper carried out the layout. We are grateful for their
professional work.
This thematic study was prepared thanks to funding from the CGIAR Research Program
on Forests, Trees and Agroforestry.

v
Contents
Foreword iii
Acknowledgements iv
Part 1 Overview 1
Chapter 1 Introduction 3
Evert Thomas, Riina Jalonen, Leonardo Gallo, David Boshier and Judy Loo
1.1. Objectives and organization of the study 8
Insight 1 Examples illustrating the importance of genetic considerations
in ecosystem restoration 13
David Boshier, Evert Thomas, Riina Jalonen, Leonardo Gallo and Judy Loo
Insight 2 The Great Green Wall for the Sahara and the Sahel Initiative:
building resilient landscapes in African drylands 15
Nora Berrahmouni, François Tapsoba and Charles Jacques Berte
Insight 3 Invasive species and the inappropriate use of exotics 19
Philip Ivey
Part 2 Theoretical and practical issues in ecosystem restoration 23
Chapter 2 Seed provenance for restoration and management: conserving
evolutionary potential and utility 27
Linda Broadhurst and David Boshier
2.1. Local versus non-local seed 28
2.2. Basic concepts and theory 28
2.3. Historical perspective of local adaptation 29
2.4. The scale of local adaptation in trees: how local should
a seed source be? 29
2.5. Are non-local seed sources ever appropriate? 30
2.6. Local seed sources may not produce restoration-quality seed 31
2.7. Adaptation and climate change 32
2.8. Benefits of using larger but more distant seed sources 33
2.9. Conclusions 33
Chapter 3 Continuity of local genetic diversity as an alternative
to importing foreign provenances 39
Kristine Vander Mijnsbrugge
3.1. Why should autochthonous diversity be protected? 39
3.2. Inventory of autochthonous woody plants 40
3.3. Producing autochthonous planting stock 40
3.4. Seed orchards 41

vi
3.5. Promotion of use 43
3.6. Discussion 45
Insight 4 Historical genetic contamination in pedunculate oak
(Quercus robur L.) may favour adaptation 47
Sandor Bordacs
Insight 5 The development of forest tree seed zones in
the Pacific Northwest of the United States 49
Brad St Clair
Chapter 4 Fragmentation, landscape functionalities and connectivity 53
Tonya Lander and David Boshier
4.1. Genetic problems related to fragmentation 53
4.2. Management of fragmented landscapes 55
4.3. The use of native species in ensuring functionality in fragmented
landscapes 57
4.4. Conclusions: policy and practice 59
Chapter 5 Gene flow in the restoration of forest ecosystems 67
Leonardo Gallo and Paula Marchelli
5.1. Genetic effects at different scales 68
5.2. Considerations in restoration and management 68
Chapter 6 The role of hybridization in the restoration of forest
ecosystems 75
Leonardo Gallo
6.1. The impact of restoration 75
6.2. Promoting hybridization 76
6.3. Avoiding hybridization 76
6.4. Seed sources and seed-zone transfer 76
Chapter 7 Collection of propagation material in the absence of genetic
knowledge 79
Gösta Eriksson
7.1. Evolutionary factors 79
7.2. Methods for sampling diversity 80
7.3. Genetic variation 80
7.4. Avoidance of genetic drift 83
7.5. Conclusion 84
Chapter 8 Evaluation of different tree propagation methods in ecological
restoration in the neotropics 85
R.A. Zahawi and K.D. Holl
8.1. Establishing tree seedlings from seed in nurseries 85
8.2. Establishment by vegetative propagation 88
8.3. Direct seeding 90
8.4. Choosing an appropriate restoration strategy 91

vii
Chapter 9 Seed availability for restoration 97
David J. Merritt and Kingsley W. Dixon
9.1. Landscape-scale restoration requires large quantities of seed 97
9.2. Seeding rates necessary to delivery restoration outcomes 98
9.3. Constraints to seed supply for landscape-scale restoration 99
9.4. Approaches to improving seed availability for restoration 100
9.5. Conclusion 102
Insight 6 Seed availability: a case study 105
Paul P. Smith
Insight 7 The role of seed banks in habitat restoration 106
Paul P. Smith
Chapter 10 Traditional ecological knowledge, traditional resource
management and silviculture in ecocultural restoration of
temperate forests 109
Dennis Martinez
Chapter 11 Designing landscape mosaics involving plantations of native
timber trees 121
David Lamb
11.1. How much reforestation? 121
11.2. What kind of reforestation? 122
11.3. Where to undertake reforestation? 122
11.4. How to plan and implement restoration on a landscape scale? 123
11.5. Will forest landscape restoration succeed in conserving
all biodiversity? 124
11.6. Conclusion 124
Insight 8 Identifying and agreeing on reforestation options
among stakeholders in Doi Suthep-Pui National Park,
northern Thailand 126
David Lamb
Part 3 Methods 129
Chapter 12 Ecological restoration approaches 133
12.1. Miyawaki method 133
Akira Miyawaki
12.1.1. Tropical rainforest rehabilitation project in Malaysia using the
Miyawaki Method 137
Nik Muhamad Majid
12.1.2. Adapting the Miyawaki method in Mediterranean forest
reforestation practices 140
Bartolomeo Schirone and Federico Vessella

viii
12.2. Framework species method 144
Riina Jalonen and Stephen Elliott
12.3. Assisted natural regeneration 148
Evert Thomas
12.3.1. Assisted natural regeneration in China 149
Jiang Sannai
12.4. Post-fire passive restoration of Andean Araucaria–Nothofagus
forests 151
Mauro E. González
12.5. Carrifran Wildwood: using palaeoecological knowledge
for restoration of original vegetation 157
Philip Ashmole
12.6. The Xingu Seed Network and mechanized direct seeding 161
Eduardo Malta Campos Filho, Rodrigo G. P. Junqueira, Osvaldo L.deSousa,
Luciano L. Eichholz, Cassiano C. Marmet, José Nicola M.N.da Costa, Bruna
D. Ferreira, Heber Q. Alves and André J. A. Villas-Bôas
Chapter 13 Approaches including production objectives 165
13.1. Analogue forestry as an approach for restoration and ecosystem
production 165
Carlos Navarro and Orlidia Hechavarria Kindelan
13.1.1. Restoring forest for food and vanilla production under
Erythrinaand Gliricidia trees in Costa Rica using the analogue
forestry method 168
Carlos Navarro
13.1.2. Restoration of ecosystems on saline soils in Eastern Cuba usingthe
analogue forestry method 169
Orlidia Hechavarria Kindelan
13.2. Post-establishment enrichment of restoration plots with timber
andnon-timber species 173
David Lamb
13.3. Enrichment planting using native species (Dipterocarpaceae)
with local farmers in rubber smallholdings in Sumatra, Indonesia 178
Hesti L. Tata, Ratna Akiefnawati and Meine van Noordwijk
13.4. “Rainforestation”: a paradigm shift in forest restoration 184
Paciencia P. Milan
13.5. The permanent polycyclic plantations: narrowing the gap
between tree farming and forest 188
Enrico Buresti Lattes, Paolo Mori and Serena Ravagni

ix
Chapter 14 Habitat-specific approaches 195
14.1. Mangrove forest restoration and the preservation of mangrove
biodiversity 195
Roy R. Lewis III
14.2. Forest restoration in degraded tropical peat swamp forests 200
Laura L.B. Graham and Susan E. Page
14.3. Support to food security, poverty alleviation and soil-degradation
control in the Sahelian countries through land restoration and
agroforestry 204
David Odee and Meshack Muga
14.4. The use of native species in restoring arid land and biodiversity
in China 207
Lu Qi and Wang Huoran
14.5. Using native shrubs to convert desert to grassland in the northeast
ofthe Tibetan Plateau 212
Yang Hongxiao and Lu Qi
14.6. Reforestation of highly degraded sites in Colombia 214
Luis Gonzalo Moscoso Higuita
Chapter 15 Species restoration approaches 225
15.1. Species restoration through dynamic ex situ conservation:
Abiesnebrodensis as a model 225
Fulvio Ducci
15.2. Restoration and afforestation with Populus nigra in Hungary 233
Sándor Bordács and István Bach
15.3. Restoration of threatened Pinus radiata on Mexico’s Guadalupe
Island 236
J. Jesús Vargas-Hernández, Deborah L. Rogers and Valerie Hipkins
15.4. A genetic assessment of ecological restoration success in
Banksiaattenuata 240
Alison Ritchie
Part 4 Analysis 243
Chapter 16 Analysis of genetic considerations in restoration methods 245
Riina Jalonen, Evert Thomas, Stephen Cavers, Michele Bozzano, David Boshier,
Sándor Bordács, Leonardo Gallo, Paul Smith and Judy Loo
16.1. Appropriate sources of forest reproductive material 245
16.1.1. Needs for research, policy and action 249
16.2. Species selection and availability 249
16.2.1. Needs for research, policy and action 252
16.3. Choice of restoration and propagation methods 253
16.3.1. Needs for research, policy and action 255

x
16.4. Restoring species associations 255
16.4.1. Needs for research, policy and action 257
16.5. Integrating restoration initiatives in human landscapemosaics 258
16.5.1. Needs for research, policy and action 259
16.6. Climate change 260
16.6.1. Needs for research, policy and action 263
16.7. Measuring success 264
16.7.1. Needs for research, policy and action 267
Part 5 Conclusions and recommendations 275
Chapter 17 Conclusions 277
Evert Thomas, Riina Jalonen, Judy Loo, Stephen Cavers, Leonardo Gallo, David
Boshier, Paul Smith, Sándor Bordács and Michele Bozzano
17.1. Recommendations arising from the thematic study 280
17.1.1. Recommendations for research 280
17.1.2. Recommendations for restoration practice 280
17.1.3. Recommendations for policy 281

Part 1
OVERVIEW

3
GENET IC CONS IDERATI ONS IN ECO SYSTEM RE STORATI ON USING N ATIVE TRE E SPECIE S
FAO (2010) estimates that 13 million hectares of
natural forests are lost each year worldwide. This
has been accompanied by an increase in the area
reforested and of forested ecosystems restored.
Between 2000 and 2010, almost 5 million hectares
of trees were planted annually, an area equiva-
lent to that of Costa Rica (FAO, 2010). It is esti-
mated that 76 percent of this area was planted
mainly for productive purposes and 24 percent
for protective purposes, although planted forests
in both categories may serve multiple purposes
(FAO, 2006). Presumably, many trees were also
planted in other types of landscape and pro-
duction systems that were not included in these
statistics, such as farmland, and for which little
information is available on a global scale. The
area of planted forests is expected to continue to
increase, reaching 300 million hectares by 2020
(FAO, 2010). Examples of large-scale reforesta-
tion and forest restoration initiatives are listed in
Table 1.1.
The global interest in planting trees holds
significant promise for restoring degraded eco-
systems, mitigating effects of environmental
changes, conserving biodiversity, and yielding
products and services that support local people’s
livelihoods. Globally, it is estimated that 2billion
hectares of land could benefit from restoration;
this is an area larger than South America (WRI,
2011; Laestadius et al., 2012). The ability of for-
est ecosystem restoration to mitigate the impacts
of numerous environmental problems, and to
slow and eventually reverse their negative ef-
fects, is widely recognized in international agree-
ments, including the United Nations Framework
Convention on Climate Change, the Convention
on Biological Diversity, the United Nations
Convention to Combat Desertification, the Aichi
Biodiversity Targets1 and the European Union
Biodiversity Targets for 2020.2 In particular, resto-
ration and reforestation hold vast potential not
only for mitigating the impacts of climate change,
through sequestration of atmospheric carbon di-
oxide in plant biomass (Canadell and Rapauch,
2008; Alexander et al., 2011a), but also for halt-
ing biodiversity loss and countering the encroach-
ment of the arid frontier (see Insight2).
In spite of serious concerns that restoration may
become a new excuse for continued agribusiness
exploitation and expanded industrial plantations
of exotic tree species that are not likely to en-
hance biodiversity and ecosystem services or ben-
efit local communities (Alexander et al., 2011a),
the growing global interest in reforestation and
restoration is accompanied by an increasing in-
terest in using native plant material (Rogers and
Montalvo, 2004; Aronson et al., 2011; Montagini
and Finney, 2011; Newton and Tejedor, 2011;
Lamb, 2012). However, an important concern in
the shift to native species is the selection of ap-
propriate genetic planting stocks for use in resto-
ration activities (Rogers and Montalvo, 2004).
1 http://www.cbd.int/sp/targets/
2 http://www.cbd.int/nbsap/about/targets/eu
Chapter 1
Introduction
Evert Thomas,1 Riina Jalonen,1 Leonardo Gallo,1,2 David Boshier1,3 and Judy Loo1
1 Bioversity International, Italy
2 Unidad de Genética Ecológica y Mejoramiento Forestal, INTA Bariloche, Argentina
3 Department of Plant Sciences, University of Oxford, United Kingdom

THE STATE OF THE WORLD’S FOREST GENETIC RESOURCES – THEMATIC STUDY
PART 1
4
In this thematic study we discuss the use of
native species and genetic considerations in a
selection of current approaches to ecosystem res-
toration, and identify the most important bottle-
necks that currently restrict the generalized use
of native species, and which may put at risk the
long-term success of restoration efforts. Our main
message is that increasing the use of native spe-
cies in restoration activities provides real environ-
mental and livelihood benefits, but also involves
clear risks, mainly related to the selection of the
appropriate genetic source for the target plant
species.
First and foremost, increasing the use of na-
tive species in restoration activities contributes
to conservation of the species themselves and
TABLE 1.1.
Examples of large-scale tree planting and forest landscape restoration initiatives (as of March 2012)
Initiative
(year of initiation) Scale Country or region Leading or coordinating
institution
Green Belt Movement (1977) 45 million trees planted Originating in Kenya, now a
worldwide movement
Established by Professor Wangari
Maathai
Green Wall of China (1978) Planned to be 4500 km long and cover
35 million ha, of which it is estimated
that two-thirds have been achieved
so far
China, bordering the Gobi
desert
Government of China
Great Green Wall
(2005)
Planned to be a tree belt 15 km wide
and 7775 km long, with an area of 11.7
million ha
Sahel across Africa, with
11 countries, from Senegal
to Djibouti, participating
African Union
Billion Tree Campaign (2006) 12 billion trees planted Global United Nations Environment
Programme, Plant for the Planet
Foundation
The Atlantic Forest Restoration
Pact (2009)
Aims to restore 15 million ha of
degraded lands in the Brazilian Atlantic
Forest biome by 2050, and to sustainably
manage the remaining forest fragments
Brazilian Atlantic Forest
biome
Joint effort of non-governmental
organizations, the private sector,
government and research
institutions
The Green Mission
(2010)
Plans to afforest or restore 5 million ha
of degraded and cleared forests, and
improve the quality of another 5 million
ha over the next 10 years
India Ministry of Environment and
Forests
Aichi Nagoya Target 15
(2010)
Restoration of at least 15% of degraded
ecosystems by 2020, as part of the
target to enhance ecosystem resilience
and the contribution of biodiversity to
carbon stocks through conservation and
restoration
Global Parties to the Convention on
Biological Diversity
Rwanda’s Forest Landscape
Restoration Initiative (2011)
Plans to restore forest nationwide “from
border to border”
Rwanda The Government of Rwanda
in collaboration with the
International Union for
Conservation of Nature (IUCN), the
Secretariat of the United Nations
Forum on Forests and the private
sector
The Bonn Challenge (2011) Targets to restore 150 million ha of
deforested and degraded lands
Global Announced at the Bonn Challenge
Ministerial Roundtable in
September 2011; supported
and promoted by IUCN, World
Resources Institute and the Global
Partnership on Forest Landscape
Restoration, among others

5
GENET IC CONS IDERATI ONS IN ECO SYSTEM RE STORATI ON USING N ATIVE TRE E SPECIE S
their genetic diversity. Second, if planting ma-
terial represents not only a native species but
originates from seed sources local to the plant-
ing site, it will have evolved together with oth-
er native flora and fauna of the area. It should
therefore be well adapted to cope with the local
environment and should support native biodiver-
sity and ecosystem resilience to a greater extent
than would introduced (exotic) planting material
(Tang et al., 2007). Third, native species may be
less likely either to become invasive or to suc-
cumb to introduced or native pests than exotic
species (Ramanagouda et al., 2010; Hulme, 2012).
Finally, native species may correspond better to
the preferences of local people, and chances are
also higher that local people hold ethnobotani-
cal and ethno-ecological knowledge of native
species, which may facilitate their successful use
in restoration projects (Shono, Cadaweng and
Durst, 2007; Chazdon, 2008; Douterlungne et al.,
2010). In turn, promoting native species that pro-
duce non-timber forest products can contribute
to the conservation of related traditional knowl-
edge as well as the cultures that maintain it.
Use of exotic species in reforestation and forest
restoration can result in negative impacts for con-
servation and the environment (Richardson, 1998;
Pimentel, Zuniga and Morrison, 2005; Stinson
et al., 2006; Tang et al., 2007; see also Insight 3:
Invasive species and the inappropriate use of ex-
otics). However, it must be recognized that the
exotic versus native species debate is not free of
controversy. There may be situations in which the
benefits generated by exotics largely outweigh
the disadvantages, not only in socioeconomic
terms but also in ecological terms (D’Antonio and
Meyerson, 2002; Alexander et al., 2011a). In addi-
tion, it would be unrealistic to think that exotics
can be completely eliminated from the environ-
ments in which they have been introduced and
in some cases have become naturalized. Better
understanding of local people’s preferences can
help promote the use of those exotics already
introduced, with clear benefits for restoration
projects. However, species with known invasive
potential should be avoided.
It is not always easy to establish with certainty
whether a species is native to a particular area
or has been introduced by humans, possibly long
ago (e.g. Vendramin et al., 2008). Some exotic tree
species – most notably Eucalyptus and Pinus spp. –
have been deliberately introduced to various parts
of the world for their perceived greater utility or
production capacity, and because know ledge
about their propagation is generally greater than
that about native alternatives. The global spread
of homogeneous planted forests, centred on eu-
calypts, pines and poplars, was largely driven by
industry that had developed in areas where these
species occurred naturally and had tailored its pro-
duction lines to the wood properties of these spe-
cies. In addition, the distribution of species (and
provenances) by humans is often an outcome of
unplanned events (Finkeldey, 2005).
It is clear that in the short term it will not be
possible to replace the predominant use of exotics
with use of native species for restoration and re-
forestation. Currently, most of the planted forests
in the tropics still comprise exotic tree species se-
lected mainly for their production functions. The
proportion of exotic species in afforestation or re-
forestation initiatives between 2003 and 2007 was
reported to be 82 percent in western and central
Africa, 99 percent in eastern and southern Africa,
28 percent in East Asia, 94 percent in South and
Southeast Asia and 98 percent in South America
(calculated from FAO, 2010: 92). While there are
probably hundreds of native species with growth
performance and wood quality at least compara-
ble to that of the commonly used plantation spe-
cies, lack of knowledge about the biology, propa-
gation and management of such native species
is currently among the main constraints for their
wider use (Newton, 2011; Lamb, 2012), along with
the difficulties of trying to alter industrial systems
tailored to particular production species. The time
seems ripe now for large-scale investments to
overcome these limitations.
Despite the expected benefits of using na-
tive species, increasing the scale of restoration
activities will be associated with elevated risks
of failure if some basic guidelines are not fol-

THE STATE OF THE WORLD’S FOREST GENETIC RESOURCES – THEMATIC STUDY
PART 1
6
lowed. For example, only two out of 98 publicly
funded reforestation projects in Brazil were con-
sidered successful during an evaluation in 2000
(Wuethrich, 2007). Reforestation and restoration
efforts may fail for a variety of reasons, from
wrong species for wrong sites to inappropriate
silvicultural approaches and techniques (Rogers
and Montalvo, 2004; Le et al., 2012). In general,
little information is available about the global
success of tree-planting efforts, especially in areas
where ecosystems may be severely degraded or
initial growing conditions are particularly harsh.
People are often hesitant to share information on
failures in spite of the help it could provide to im-
proving current practices, and global efforts to re-
cord reforestation and forest restoration activities
started only recently (FAO, 2010). However, the
annual average area reported for afforestation
and reforestation activities globally in 2003–07
was more than twice the annual average increase
in the area of planted forests over the ten-year
period 2000–2010 (FAO, 2010). Low success rates
in establishment and survival of seedlings can be
assumed to contribute to the difference.
Although the reasons for frequent failures in
reforestation and restoration activities are not of-
ten known, it is probable that many failures are
related to poor matching of planting material to
the target site, or too narrow a genetic base for
the planting stock (Rogers and Montalvo, 2004).
Indeed, to attain a functional and resilient eco-
system, it is crucial that the genetically adapted
planting material used for establishing a plant
community represents a certain minimum level of
intraspecific diversity to ensure that its progeny
will in turn be viable and able to produce viable
offspring. Aside from the initial quality and ge-
netic diversity of germplasm, and its suitability for
the planting site, the extent of gene flow across
landscapes over subsequent generations is also of
central importance for the successful long-term
restoration of ecosystems and tree populations.
This ensemble of genetic qualities is necessary
not only to provide the desired forest functions,
products and services, but also to enable restored
populations to reproduce and survive on the site.
Genetic diversity has generally been found
to be positively related not only with the fit-
ness of individual plant populations (Reed and
Frankham, 2003; Rogers and Montalvo, 2004), but
also with the stability and resilience of ecosystems
(Gregorius, 1996; Elmqvist et al., 2003; Müller-
Starck, Ziehe and Schubert, 2005; Thompson et
al., 2010; Sgro, Lowe and Hoffmann, 2011). Tree
communities need particularly adaptive genetic
variation to succeed over time on the restored site;
such variation promotes survival and good growth
while at the same time enhancing resilience and
resistance to biotic and abiotic stresses such as en-
vironmental variations (Pautasso, 2009; Dawson
et al., 2011; Schueler et al., 2012) or pests and
pathogens (Schweitzer et al., 2005; Cardinale et
al., 2012). In the long term, adaptive genetic diver-
sity will promote successful reproduction, reduce
the risk of inbreeding and genetic impoverishment
that can result from genetic drift, and increase a
population’s ability to adapt to future site condi-
tions.
Currently little is known about the genetic
diversity of most native species, particularly the
thousands of tropical tree species that could play
an important role in restoring degraded tropical
ecosystems and their functions. Where guidelines
exist, for example on the collection of germplasm,
they appear to be largely unknown or overlooked
by restoration practitioners. Moreover, despite
the high expectations for restored forests to miti-
gate climate change, ensuring the capability of
tree populations to adapt to changing environ-
ment as a precondition for their mitigation func-
tion has received hardly any attention. The fact
that the negative effects of genetic homogeneity
are not necessarily immediately evident but accu-
mulate over time means that resulting problems
are difficult to perceive (Rogers and Montalvo,
2004) and address. Furthermore, by the time the
effects are obvious they may already have af-
fected large areas. For example, low genetic di-
versity in planting material, stemming from col-
lecting seed from single isolated trees, can lead to
increased homozygosity, particularly in the next
generation, and may result in the expression of

7
GENET IC CONS IDERATI ONS IN ECO SYSTEM RE STORATI ON USING N ATIVE TRE E SPECIE S
A degraded ecosystem “exhibits loss of biodiversity
and a simplification or disruption in ecosystem
structure, function and composition caused by
activities or disturbances that are too frequent or
severe to allow for natural regeneration or recovery”
(Alexander
et al.
, 2011b).
Ecological restoration is “the process of assisting
the recovery of an ecosystem that has been degraded,
damaged, or destroyed” (SER, 2004). Alexander
et al.
(2011b) define ecological restoration as “an intentional
activity that initiates or facilitates the recovery of
ecosystems by re-establishing a beneficial trajectory
of maturation that persists over time. The science
and practice of ecological restoration is focused
largely on reinstating autogenic ecological processes
by which species populations can self-organize into
functional and resilient communities that adapt to
changing conditions while at the same time delivering
vital ecosystem services. In addition to reinstating
ecosystem function, ecological restoration also fosters
the re-establishment of a healthy relationship between
humans and their natural surroundings by reinforcing
the inextricable link between nature and culture and
emphasizing the important benefits that ecosystems
provide to human communities.”
Forest restoration aims to “restore the forest to
its state before degradation (same function, structure
and composition)” (ITTO, 2002).
Forest landscape restoration is “a planned
process that aims to regain ecological integrity and
enhance human wellbeing in deforested or degraded
forest landscapes” (WWF and IUCN, 2001).
Rehabilitation is “a process to re-establish the
productivity of some, but not necessarily all, of the
plant and animal species thought to be originally
present at a site. For ecological or economic reasons
the new forest might also include species not
originally present at the site. The protective function
and many of the ecological services of the original
forest may be re-established” (Gilmour, San and Xiong
Tsechalicha, 2000).
Reforestation is “the re-establishment of forest
through planting and/or deliberate seeding on land
classified as forest, for instance after a fire, storm or
following clearfelling” (FAO, 2010).
Afforestation is “the act of establishing forests
through planting and/or deliberate seeding on land
that is not classified as forest” (FAO, 2010).
Planted forests are forests “composed of
trees established through planting and/or through
deliberate seeding of native or introduced species”
(FAO, 2010).
Resilience is “the ability of an ecosystem to
recover from, or to resist stresses (e.g. drought, flood,
fire or disease)” (Walker and Salt, 2006).
A native species (also indigenous species) is a
species which is part of the original flora of an area
(IBPGR, now Bioversity International).
An exotic species (also alien or introduced
species) is “a species which is not native to the region
in which it occurs” (FAO, 2002).
Naturalized species are “intentionally or
unintentionally introduced species that have
adapted to and reproduce successfully in their new
environments” (FAO, 2002).
A provenance refers to “the original geographic
source of seed, pollen or propagules” (FAO, 2002).
References
Alexander, S., Aronson, J., Clewell, A., Keenleyside, K.,
Higgs, E., Martinez, D., Murcia, C. & Nelson, C. 2011b.
Re-establishing an ecologically healthy relationship
between nature and culture: the mission and vision of
the Society for Ecological Restoration.
In
Secretariat of
the Convention on Biological Diversity.
Contribution
of ecosystem restoration to the objectives of the CBD
and a healthy planet for all people. Abstracts of posters
presented at the 15th Meeting of the Subsidiary Body
on Scientific, Technical and Technological Advice of the
Convention on Biological Diversity, 7–11 November 2011,
Montreal, Canada.
Technical Series No. 62, pp. 11–14.
Montreal, Canada, SCBD.
FAO (Food and Agriculture Organization of the United
Nations). 2002.
Glossary on forest genetic resources
(English version)
. Forest Genetic Resources Working
Papers, Working Paper FGR/39E. Rome.
FAO (Food and Agriculture Organization of the United
Nations). 2010.
Global forest resources assessment. Main
report.
FAO Forestry Paper 163. Rome.
Box 1.1.
Key concepts in ecosystem restoration

THE STATE OF THE WORLD’S FOREST GENETIC RESOURCES – THEMATIC STUDY
PART 1
8
deleterious recessive alleles, which in turn de-
creases individual fitness (i.e. inbreeding depres-
sion) (White, Adams and Neale, 2007). Inbreeding
can have impacts at any stage of development, for
example through reduced embryo viability, seed-
ling survival, tree vigour or seed production (see
Insight 1: Examples illustrating the importance of
genetic considerations in ecosystem restoration).
Restoration, rehabilitation and reforesta-
tion are all terms commonly used to refer to re-
establishing forest vegetation on deforested are-
as. In this study we use the term “ecosystem resto-
ration.” This largely coincides with “ecological res-
toration,” defined as “the process of assisting the
recovery of an ecosystem that has been degraded,
damaged, or destroyed” (SER, 2004), but also aims
to accommodate rehabilitation and reforesta-
tion activities that do not necessarily comply with
some more conservative definitions of restora-
tion (Lamb, 2012). These and other terms related
to ecosystem restoration are defined in Box1.1.
We acknowledge that restoration is not the most
appropriate term for characterizing some of the
activities described in this and the following chap-
ters because it suggests the aim of re-establishing
a pre-existing ecosystem. In some cases it is almost
impossible to define a previous state to which an
ecosystem can be restored (Hilderbrand, Watts and
Randle, 2005). It may also be impossible to return
ecosystems to historical states because of radical
changes that have already taken place (e.g. severe
aridification, soil degradation or socioeconomic
changes) (Buizer, Kurz and Ruthrof, 2012), or the
objective of a restoration activity may simply be
less ambitious with respect to the plant commu-
nity it aims to establish (Lamb, 2012). In spite of
these shortcomings, we have chosen to use “eco-
system restoration” throughout this study for the
sake of uniformity.
While the systems and approaches discussed in
this study cover a range of objectives and species
assemblages, sometimes including exotic species,
they all emphasize the use of indigenous tree spe-
cies and diversity for their intrinsic relationships
with indigenous flora and fauna and local know-
ledge and cultures.
1.1. Objectives and organization
of the study
The objective of this thematic study is to review
and analyse current practices in ecosystem resto-
ration, with a particular focus on the use of native
tree species and genetic considerations related to
the selection of appropriate planting material.
Based on this analysis we put forward a number
of practical recommendations, including genetic
considerations in ecosystem restoration, that are
intended to help practitioners to avoid genetic
problems and enhance both the short- and long-
term success of future restoration activities. Our
target audience includes researchers, restoration
practitioners and policy-makers.
Gilmour, D.A., San, N.V. & Xiong Tsechalicha. 2000.
Rehabilitation of degraded forest ecosystems in Cambodia,
Lao PDR, Thailand and Vietnam: an overview.
Pathumthani,
Thailand, IUCN, The World Conservation Union, Asia
Regional Office.
ITTO (International Tropical Timber Organization). 2002.
ITTO guidelines for the restoration, management and
rehabilitation of degraded and secondary tropical forests.
ITTO
Policy Development Series No. 13. Yokohama, Japan, ITTO.
SER (Society for Ecological Restoration). 2004.
SER
international primer on ecological restoration
. SER,
Washington, DC (available at: http://www.ser.org/
resources/resources-detail-view/ser-international-primer-
on-ecological-restoration).
Walker, B. & Salt, D. 2006.
Resilience thinking: sustaining
ecosystems and people in a changing world
. Washington,
DC, Island Press.
WWF & IUCN. 2000.
Forests reborn. A workshop on forest
restoration.
WWF/IUCN International Workshop on Forest
Restoration: 3–5 July 2000, Segovia, Spain (available
at: http://cmsdata.iucn.org/downloads/flr_segovia.pdf).
Accessed 21 January 2013.
Box 1.1. (continued)
Key concepts in ecosystem restoration (continued)

9
GENET IC CONS IDERATI ONS IN ECO SYSTEM RE STORATI ON USING N ATIVE TRE E SPECIE S
This report is organized in five main parts,
including this introduction. In the second part,
experienced scientists briefly present theoretical
and practical issues relevant to ecosystem resto-
ration, with particular emphasis on genetic as-
pects. This more theoretical series of contributions
serves as a basis for the analysis of the restora-
tion methods and approaches and underpins the
recommendations. The third part is an overview
of various methods and approaches that are cur-
rently used in ecosystem restoration and are based
– at least partially – on the use of native species.
The authors contributing to the presentation of
these methods and approaches were requested to
reply to a set of questions aimed at facilitating an
analysis of the methods they used and their genet-
ic implications; the questionnaire is available on
the Bioversity website.3 The fourth part presents
an analysis of the use of genetic considerations in
current restoration methods, as well a number of
action and research recommendations, building
on the previous chapters of theoretical and gen-
eral considerations, presentation of the methods
and approaches, and the responses to the survey.
The fifth and final part summarizes the main con-
clusions of this thematic study.
References
Alexander, S., Nelson, C.R., Aronson, J., Lamb, D.,
Cliquet, A., Erwin, K.L., Finlayson, C.M., de
Groot, R.S., Harris, J.A., Higgs, E.S., Hobbs, R.J.,
Robin Lewis, R.R., Martinez, D. & Murcia, C.
2011a. Opportunities and challenges for eco-
logical restoration within REDD+. Restor. Ecol., 19:
683–689.
Alexander, S., Aronson, J., Clewell, A., Keenleyside,
K., Higgs, E., Martinez, D., Murcia, C. & Nelson,
C. 2011b. Re-establishing an ecologically healthy
relationship between nature and culture: the mission
and vision of the Society for Ecological Restoration.
In Secretariat of the Convention on Biological
Diversity. Contribution of ecosystem restoration
to the objectives of the CBD and a healthy planet
for all people. Abstracts of posters presented at
the 15th Meeting of the Subsidiary Body on
Scientific, Technical and Technological Advice of the
Convention on Biological Diversity, 7–11 November
2011, Montreal, Canada. Technical Series No. 62,
pp. 11–14. Montreal, Canada, SCBD.
Aronson, J., Brancalion, P.H.S., Durigan, G.,
Rodrigues, R.R., Engel, V.L., Tabarelli, M.,
Torezan, J.M.D., Gandolfi, S., de Melo, A.C.G.,
Kageyama, P.Y., Marques, M.C.M., Nave, A.G.,
Martins, S.V., Gandara, F.B., Reis, A., Barbosa,
L.M. & Scarano, F.R. 2011. What role should gov-
ernment regulation play in ecological restoration?
Ongoing debate in São Paulo State, Brazil. Restor.
Ecol., 19: 690–695.
3 See https://www.bioversityinternational.org/fileadmin/user_
upload/SoW_FGR_RestorationSurvey.pdf.
While the emphasis here has been on sourcing seed
for restoration, it is important to recognise that
many species have intimate associations with a
range of organisms and that these too may require
restoration. The “If you build it, they will come”
paradigm does not always apply and ill-considered
placement of restoration projects can lead to poor
utilization by the very organisms they are expected
to attract to recreate interactions and processes at
the population and community level. In addition,
there can be considerable benefits for simultaneously
restoring plants and associated organisms. For
example, the survival and growth of acacias is
significantly improved if seed is simultaneously
planted with nitrogen-fixing bacterial symbionts, with
excess nitrogen benefiting other co-planted species,
resulting in a better and more rapid restoration
outcome (Thrall
et al.,
2005).
Reference
Thrall, P.H., Millsom, D.A., Jeavons, A.C., Waayers, M.,
Harvey, G.R., Bagnall, D.J. & Brockwell, J. 2005. Seed
inoculation with effective root-nodule bacteria enhances
revegetation success.
J. Appl. Ecol.,
42: 740–751.
Box 1.2.
It’s not just about restoring plants

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Buizer, M., Kurz, T. & Ruthrof, K. 2012. Understanding
restoration volunteering in a context of environmen-
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Canadell, J.G. & Raupach, M.R. 2008. Managing
forests for climate change mitigation. Science, 320:
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Cardinale, B.J., Duffy, J.E., Gonzalez, A., Hooper,
D.U., Perrings, C., Venail, P., Narwani, A., Mace,
G.M., Tilman, D., Wardle, D.A., Kinzig, A.P.,
Daily, G.C., Loreau, M., Grace, J.B., Larigauderie,
A., Srivastava, D.S. & Naeem, S. 2012. Biodiversity
loss and its impact on humanity. Nature, 486:
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Chazdon, R.L. 2008. Beyond deforestation: restoring
forests and ecosystem services on degraded lands.
Science, 320: 1458–1460.
D’Antonio, C. & Meyerson, L.A. 2002. Exotic plant
species as problems and solutions in ecological resto-
ration: a synthesis. Restor. Ecol., 10: 703–713.
Dawson. I., Lengkeek, A., Weber, J. & Jamnadass,
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Douterlungne, D., Levy-Tacher, S.I., Golicher, D.J. &
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Elmqvist, T., Folke, C., Nyström, M., Peterson, G.,
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Gilmour, D.A., San, N.V. & Xiong Tsechalicha. 2000.
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Thrall, P.H., Millsom, D.A., Jeavons, A.C., Waayers,
M., Harvey, G.R., Bagnall, D.J. & Brockwell, J.
2005. Seed inoculation with effective root-nodule
bacteria enhances revegetation success. J. Appl.
Ecol., 42: 740–751.
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flr_segovia.pdf). Accessed 21 January 2013

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GENET IC CONS IDERATI ONS IN ECO SYSTEM RE STORATI ON USING N ATIVE TRE E SPECIE S
Poor genetic matching of planting material to
the target site may result in reduced viability
of restoration projects
The widespread and severe dieback in three pon-
derosa pine plantations planted south of Pagosa
Springs, Colorado, United States, in the late
1960s to mid-1970s has been related to the use
of inappropriate genetic seed source. A pathogen
(Cenangium ferruginosum) has been identified
in the plantations, but observations are consist-
ent with this being a secondary impact and not
the primary cause of failure (Worral, 2000; Rogers
and Montalvo, 2004).
Use of provenance trials to guide genetic
matching
The natural range of black walnut (Juglans
nigra L.) extends from the eastern United States
west to Kansas, South Dakota and eastern Texas.
A subset of 15 to 25 sources from 66 sampled
provenances was planted in each of seven geo-
graphically disparate common-garden field trials.
After 22 years, survival was much higher for local
trees (71 percent) than for the other provenances
(zero survival at some sites) (Bresnan et al., 1994;
Rogers and Montalvo, 2004). This allowed the
authors to make informed decisions about where
best to use what germplasm.
Selfing (self-pollination) can considerably
affect survival and size of offspring
In a study in which offspring of Pseudotsuga
menziesii selfed and outcrossed crosses were
compared 33 years after establishment of seed-
lings, the average survival of selfed offspring was
only 39 percent that of the outcrossed individuals.
Moreover, the average diameter at breast height
(DBH) of the surviving selfed trees was 59 percent
that of the surviving outcrossed siblings (White,
Adams and Neale, 2007).
Low levels of genetic diversity can
compromise successful mating between plant
individuals
Attempts to restore the endangered daisy Ruti-
dosis leptorrhynchoides were constrained by the
limi ted reproductive potential of small popu-
lations (fewer than 200 plants) where the low
number of self-incompatibility alleles prevented
successful mating between many of the remnant
plants (Young et al., 2000). Among trees, several
Prunus species are known to have self-incompat-
ibility alleles, so the same considerations could
apply.
Insight 1
Examples illustrating the importance
of genetic considerations in
ecosystem restoration
David Boshier,1,3 Evert Thomas,1 Riina Jalonen,1 Leonardo Gallo1,2 and Judy Loo1
1 Bioversity International, Italy
2 Unidad de Genética Ecológica y Mejoramiento Forestal, INTA Bariloche, Argentina
3 Department of Plant Sciences, University of Oxford, United Kingdom

THE STATE OF THE WORLD’S FOREST GENETIC RESOURCES – THEMATIC STUDY
PART 1
14
Negative consequences of low genetic
diversity of the source material usually
accumulate in the subsequent generations
Acacia mangium was first introduced to Sabah
(Malaysia) from Australia in 1967 in two small
stands of 34 and approximately 300 trees of the
“maternal half-sib family.” This material formed
the basis for more than 15 000 hectares of planta-
tions. A simple nursery trial comparing seedlings
from the first to third generation showed a re-
duced height growth in seedlings harvested from
the second and third generation, as compared
with the first generation (20.7 cm and 18.1 cm,
compared with 32.5 cm) (Sim, 1984).
Selection for favourable characteristics
can considerably improve the quality of
individuals where specific objectives have
been set for the planted forests
Tree improvement programmes have been suc-
cessful in dramatically increasing growth and
quality in commercially valuable and widely
planted species. For example, a study compared
the performance of Acacia auriculiformis trees
grown from seedlots obtained from: (1) a seed-
ling seed orchard (SSO), (2) a seed production
area (SPA), (3) a natural-provenance site (NPS)
and (4) a commercial seedlot from the same
provenance (CS) from Viet Nam. Four-year old
trees grown from the SSO and SPA seedlots scored
significantly higher than trees from the NPS for
a number of traits including height, DBH, conical
stem volume, stem straightness and axis persis-
tence. In contrast, trees grown from commercial
seedlots scored consistently lower for these traits
(Hai et al., 2008). Inbreeding may have contri-
buted to the poor growth and quality of trees
originating from the commercial seedlots.
References
Bresnan, D.R., Rink, G., Diesel, K.E. & Geyer, W.A.
1994. Black walnut provenance performance in
seven 22-year-old plantations. Silvae Genet., 43:
246–252.
Hai, P.H., Harwood, C., Kha, L.D., Pinyopusarerk, K. &
Thinh, H. 2008. Genetic gain from breeding Acacia
auriculiformis in Vietnam. J. Trop. Forest Sci., 20:
313–327.
Rogers, D.L. & Montalvo, A.M. 2004. Genetically
appropriate choices for plant materials to main-
tain biological diversity. Report to the USDA Forest
Service, Rocky Mountain Region, Lakewood, CO,
USA. University of California (available at: http://
www.fs.fed.us/r2/publications/botany/plantgenetics.
pdf). Accessed 21 January 2013.
Sim, B.L. 1984. The genetic base of Acacia mangium
Willd. in Sabah. In R.D. Barnes, & G.L. Gibson, eds.
Provenance and genetic improvement strategies in
tropical forest trees, pp. 597-603. Mutare, Zimbabwe,
April, 1984. Oxford, UK, Commonwealth Forestry
Institute; Harare, Zimbabwe, Forest Research Centre.
Young, A., Miller, C., Gregory, E. & Langston, A.
2000. Sporophytic self-incompatibility in diploid
and tetraploid races of Rutidosis leptorrhynchoides
(Asteraceae). Aust. J. Bot., 48: 667–672.
White, T.W., Adams, W.T. & Neale, D.B. 2007. Forest
genetics. Wallingford, UK, CABI Publishing.
Worrall, J. 2000. Dieback of ponderosa pine in planta-
tions established ca. 1970. Internal Forest Service
Report. Gunnison, CO, USA, USDA Forest Service,
Gunnison Service Center.

15
GENET IC CONS IDERATI ONS IN ECO SYSTEM RE STORATI ON USING N ATIVE TRE E SPECIE S
Desertification,4 land degradation and drought,
combined with climate change, have a strong
negative impact on the food security and liveli-
hoods of local communities in Africa’s drylands,
home to some of the world’s poorest populations.
The Great Green Wall for the Sahara and the
Sahel Initiative (GGWSSI) was launched by African
heads of state and government “to improve the
resilience of human and natural systems in the
Sahel–Saharan zone to Climate Change through
a sound ecosystems’ management, sustainable
development of land resources, protection of ru-
ral heritage and improvement of the living con-
ditions and livelihoods of populations living in
these areas.” This African Union initiative, based
on a proposal of former President of Nigeria, H.E.
Olusegun Obasanjo, involves over 20 countries
bordering the Sahara.
The Food and Agriculture Organization of the
United Nations (FAO), the European Union and
the Global Mechanism of the UNCCD are support-
ing the African Union Commission and 13 partner
countries (Algeria, Burkina Faso, Chad, Djibouti,
Egypt, Ethiopia, the Gambia, Mauritania, Mali,
Niger, Nigeria, Senegal and the Sudan) in their
efforts to implement the GGWSSI. This support
4 Desertification refers to land degradation in arid, semi-arid and
subhumid areas resulting from factors such as human pressure on
fragile ecosystems, deforestation and climate change.
involves: (i) the development and validation of
a harmonized regional strategy for effective im-
plementation and resource mobilization of the
GGWSSI; (ii) the preparation of detailed imple-
mentation plans and project portfolios in the 13
countries, identifying priorities and intervention
areas and at least three cross-border projects;
(iii) the development of a partnership and re-
source mobilization platform and a learn ing and
networking platform for enhancing knowledge
sharing, technology transfer and promotion
of best practices across GGWSSI countries and
among partners; (iv) the preparation of a capaci-
ty-building strategy and programme; and (v)the
preparation of a communication strategy and ac-
tion plan for engaging key target audiences and
stakeholders in supporting implementation of
the GGWSSI.
Among the priority interventions identified
with in the GGWSSI action plans developed to date
is the restoration of forest landscapes and de-
graded lands in the GGWSSI priority intervention
areas. Achieving this will depend on developing
the capacity of the partners in the following areas:
• use of native species adapted to the local
environmental, socioeconomic and cultural
conditions;
• selection, production and use of a wide
range of site-adapted planting material
(genotypes) from native tree, shrub and
Insight 2
The Great Green Wall for the Sahara
and the Sahel Initiative: building
resilient landscapes in African drylands
Nora Berrahmouni, François Tapsoba and Charles Jacques Berte
Forest Assessment, Management and Conservation Division, Forestry Department,
Food and Agriculture Organization of the United Nations, Rome, Italy

THE STATE OF THE WORLD’S FOREST GENETIC RESOURCES – THEMATIC STUDY
PART 1
16
grass species, including production sufficient
quantities of seeds and seedlings of
adequate quality;
• application of the principles of forest
landscape restoration planning to
restore ecological integrity and enhance
human well-being in the degraded forest
landscapes and lands;
• promoting effective stakeholder
participation and governance to ensure
effective planning, design, implementation
and sharing of benefits from afforestation
and restoration;
• promotion of sustainable management of
forests and rangelands to assist and enhance
natural regeneration;
• promotion of multipurpose agrosilvipastoral
systems and economically valuable native
plant species to improve rural livelihoods;
• combined use of traditional knowledge
and innovative forestation and restoration
techniques, with particular focus on soil and
This project, developed and implemented between
2003 and 2010, was funded by Italian Cooperation.
It aimed at strengthening the capacity of six pilot
countries (Burkina Faso, Chad, Kenya, Niger, Senegal
and the Sudan) to address food security and
desertification problems through the improvement
and restoration of the acacia-based agrosilvipastoral
systems, and at sustainably developing the resins
and gums sectors. The project benefited local
communities engaged in harvesting and processing
gums and resins. The project tested a microcatchment
water-harvesting system (the Vallerani system)1 and
restored a total of 13 240 hectares. Local people
were empowered though an intensive programme of
capacity building on the use and application of the
Vallerani system, nursery establishment and plant
production, agricultural production, and harvesting
and processing of gums and resins. Native tree
species, including
Acacia senegal, Acacia seyal, Acacia
nilotica, Acacia mellifera, Bauhinia rufescens and
Ziziphus mauritiana
, were established by planting
seedlings and by direct sowing. Herbaceous plants,
such as
Cassia tora, Andropogon gayanus
and
Cymbopogon
sp., were established by direct sowing.
The project also focused on strengthening the
Network for Natural Gums and Resins in Africa, which
involves 15 member countries, through resource
assessment, training programmes and information
sharing.
The project published a working paper, “Guidelines
on sustainable forest management in drylands in sub-
Saharan Africa,” in both English and French.
A regional meeting held in Addis Ababa,
Ethiopia, on 3–4 March 2009 identified the need
for a strategy to develop the outcomes of the pilot
project into a programme large enough to address
the magnitude of food insecurity, poverty, land
degradation and desertification in the region, and
to mitigate and adapt to climate change. The future
programme must first focus on improving livelihoods
through broadening the sources of income for local
populations, while restoring degraded lands and
increasing the productivity of agriculture, range and
forest systems. These are cross-sectoral activities and
the programme must adopt an integrated approach.
The programme will have to be of sufficient scale to
be seen as a major actor in regional initiatives, such
as the GGWSSI. Such a programme would contribute
to combating desertification, to the success of the
GGWSSI and, above all, to improving the well-being
of the whole population in the region.
Source: For further information, see http://www.fao.org/forestry/
aridzone/62998/en/.
1 See http://www.lk.iwmi.org/africa/West/projects/
Adoption%20Technology/RainWaterHarvesting/26-
ValleranisSystem.htm.
Box I2-1.
Acacia operation project – Support to food security, poverty alleviation and soil
degradation control in the gums and resins producer countries

17
GENET IC CONS IDERATI ONS IN ECO SYSTEM RE STORATI ON USING N ATIVE TRE E SPECIE S
water conservation and management;
• promoting awareness of the contribution
of forestation and drylands restoration
to climate change adaptation and
mitigation within the framework of carbon
market schemes (e.g. Clean Development
Mechanism, Reduced Emissions from
Deforestation and Forest Degradation
(REDD) and REDD+) and adaptation schemes;
• sustainable financing and investments
(e.g. through payments for environmental
services) and related policy issues;
• monitoring and evaluation of the
performance of restoration initiatives,
and the assessment of their long-
term sustainability and economic and
environmental impacts;
• considering restoration along the whole
market chain value, from the seed to the
final product.
To support the effective planning and implemen-
tation of restoration work in the priority GGWSSI
areas, FAO launched a process for developing
guidelines on dryland restoration based on a
compilation of lessons learned from past and
current forestation and restoration projects and
programmes. As a first step, the Turkish Ministry
of Forestry and Water Affairs, FAO, the Turkish
International Cooperation and Coordination
Agency (TIKA) and the German Agency for In-
ternational Cooperation (GIZ) convened an in-
ternational workshop in Konya, Turkey, in May
2012. This workshop, entitled “Building forest
landscapes resilient to global changes in drylands
This project was implemented between 2000 and
2007 by FAO and the Ministry of Environment
and Sustainable Development of Mauritania, with
financing from the Walloon Region of Belgium. The
project objective was to foster conservation and
development of agrosilvipastoral systems around
Nouakchott, while at the same time combating
encroachment of sand on the green belt around
the city. The project engaged the local community
and national authorities in planning and delivering
activities and in selecting appropriate local plant and
tree species. A total of 400 000 plants were grown in
nurseries and used to fix 857 hectares of threatened
land (inland and costal dunes).
The project employed both mechanical and
biological fixation methods. Partners and beneficiaries
were trained on field techniques and management
of tree nurseries through a participatory approach
involving the local community and the support and
supervision of technical experts from the project.
The project gave priority to the production and
use of indigenous woody and grassy species. For
example,
Aristida pungens
was planted on very
mobile strip dunes in accumulation zones. Deflation
zones were planted with
Leptadenia pyrotechnica,
Aristida pungens
and
Panicum turgidum
, while
other slow-growing woody species, such as
Acacia
raddiana
and
A. senegal
,
were planted in more stable
intermediate zones. Local grassy species were sown
using broadcast seed, while
Colocynthus vulgaris
, a
cucurbit, was sown in pouches. Establishment rate
depended on rainfall. Plantings on coastal dunes
concentrated on halophytic species, including
Nitraria
retusa, Tamarix aphylla
and
T. senegalensis
.
The techniques used and the lessons learned are
presented in detail in an FAO forestry paper published
in 2010, which is available in English, French and
Arabic. The best practices identified are now being
replicated in other regions of Mauritania and will
be promoted for adaptation and implementation in
Mauritania and other countries of the GGWSSI.
Source: FAO (Food and Agriculture Organization of the United
Nations). 2010.
Fighting sand encroachment: lessons from
Mauritania
, by C.J. Berte, with the collaboration of M. Ould
Mohamed & M. Ould Saleck. FAO Forestry Paper 158. Rome.
Box I2-2.
Support to the rehabilitation and extension of the Nouakchott green belt,
Mauritania

THE STATE OF THE WORLD’S FOREST GENETIC RESOURCES – THEMATIC STUDY
PART 1
18
– Analysis, evaluation and documentation of les-
sons learned from afforestation and forest resto-
ration,” aimed at:
• gathering lessons learned from past and
ongoing forest restoration efforts in the
countries involved in the GGWSSI;
• identifying key elements determining
the success or failure of forest restoration
projects and;
• discussing the comprehensive Forest
Restoration Monitoring Tool, recently
developed by FAO to guide planning,
implementation and evaluation of field
projects and programmes.
A number of successful forestation and forest
restoration projects exist in the GGWSSI countries
and these can be quickly upscaled to support the
effective implementation of the initiative. These
include the two projects implemented by FAO
and its partners: the Acacia operation project –
Support to food security, poverty alleviation and
soil degradation control in the gums and resins
producer countries (Box I2-1), implemented in six
sub-Saharan African countries; and the Support
to the rehabilitation and extension of the
Nouakchott green belt, funded by the Walloon
region (Belgium), and implemented in Mauritania
(Box I2-2).
For more information on the Great Green Wall
for the Sahara and Sahel Initiative, please visit
www.fao.org/partnerships/great-green-wall

19
GENET IC CONS IDERATI ONS IN ECO SYSTEM RE STORATI ON USING N ATIVE TRE E SPECIE S
Sometimes the choice of plant used in restoration
can have unexpected and dramatic consequences
both at the site of restoration and beyond. This
Insight highlights some examples in which plants
introduced from elsewhere in the world to help
restore disturbed environments resulted in inva-
sion and great environmental damage.
Exotic or non-native trees, shrubs, creepers,
succulents and grasses have all been used to reha-
bilitate sites after human or natural perturbation
has removed indigenous vegetation cover. Many
introductions of exotic plants happened late in
the nineteenth century or early in the twentieth
century, when understanding of the likely impacts
of these species was limited and not considered.
• Pueraria montana (kudzu), indigenous
to China, eastern India and Japan, was
introduced in the United States of America
as a forage and ornamental plant, but was
also extensively used in soil stabilization and
erosion control.5 It is estimated that about
120 000 hectares had been planted with
kudzu by 1946, and the species has since
spread beyond the planted range. By 2004
it was reported to be present and invasive
in 22 states of the southeastern United
States, where it causes extensive damage
by smothering indigenous vegetation. It is
not surprising that this species has a local
common name of “vine that ate the South.”
• Acacia cyclops and Acacia saligna were
5 http://www.na.fs.fed.us/fhp/invasive_plants.
both introduced in the 1830s into South
Africa from Australia to stabilize dunes and
protect roads from sand storms (Carruthers
et al., 2011) but they became invasive
species in the Western Cape of South Africa.
Successful implementation of biological
control measures to reduce seed production
of these species will reduce the long-term
threats they pose.
• Ailanthus altissima (Tree of Heaven), native
to China and northern Viet Nam, has
been used for a wide variety of purposes,
including erosion control, afforestation,
shelterbelts and to line promenades
in Europe and elsewhere in the world.
Consequently, the species has established
and become invasive in suitable, lower-
altitude environments across all of Europe.
For a comprehensive review, see Kowerik
and Säumel (2007).
• Carpobrotus edulis is known by the
descriptive local common name of
“highway iceplant” in California. The
common name refers to the species’
extensive use as a landscape plant to secure
disturbed environments along roads.
Since its introduction it has spread into
natural environments where it threatens
natural vegetation in several different
environments, from dune systems to
scrublands. Carpobrotus edulis is also a
significant problem in Mediterranean
countries, particularly Portugal.
Insight 3
Invasive species and
the inappropriate use of exotics
Philip Ivey
South African National Biodiversity Institute and Working for Water Programme

THE STATE OF THE WORLD’S FOREST GENETIC RESOURCES – THEMATIC STUDY
PART 1
20
It is important that we learn from the mistakes
of the past and if possible do not repeat them.
There are several international protocols in place
to encourage better practices to reduce the likeli-
hood of invasions. Article 8(h) of the Convention
on Biological Diversity (CBD) calls on parties to
“prevent the introduction of, control or eradi-
cate those alien species which threaten ecosys-
tems, habitats or species.” The Aichi Biodiversity
Targets6 agreed under the CBD similarly address
invasive species: “By 2020, invasive alien species
and pathways are identified and prioritized, pri-
ority species are controlled or eradicated and
measures are in place to manage pathways to
prevent their introduction and establishment”
(Aichi Target 9).
The International Standards for Phytosanitary
Measures, prepared by the Secretariat of the
International Plant Protection Convention (IPPC)
deals with “environmental risks,” including “in-
vasive plants.” The IPPC encourages each of its
regions to set regional standards. In response
to this, the European and Mediterranean Plant
Protection Organization (EPPO) has set standards
to provide support to members dealing with both
quarantine pests and more recently invasive alien
species, and members are encouraged to manage
these through national phytosanitary regulations.
Hulme (2007) estimates that 80 percent of in-
vasive alien plants in Europe were voluntarily in-
troduced for ornamental purposes. In an effort
to curb the influx of new invasive plant species
to Europe, the EPPO, in collaboration with the
Council of Europe, developed a Code of con-
duct on horticulture and invasive alien plants
(Heywood and Brunel, 2011) aimed at the horti-
cultural industry. To an extent the European code
of conduct has been based around the St Louis
Declaration of 2002,7 which calls on horticultural-
ists and the nursery industry to ensure that unin-
tended harm (risk of invasion) is kept to a mini-
mum when new plant species are considered for
introduction.
6 http://www.cbd.int/sp/targets/
7 http://www.fleppc.org/FNGA/St.Louis.htm
One of the key indicators used to assess wheth-
er a species is likely to be invasive in a particular
environment is whether it has been invasive else-
where in the world. There are numerous reference
lists of invasive and weedy plant species, including
Randall (2002), the Invasive species compendium8
and the DAISIE (Delivering Alien Invasive Species
Inventories for Europe) database.9
In order to achieve the targets set by the CBD
and to reduce the likelihood of new invasive
species being used by the horticultural industry
for landscape rehabilitation, it is important that
governments control imports of new plant spe-
cies. Horticultural interests also should regulate
their own businesses by adhering to the volun-
tary protocols to control invasive species. With
adequate control and self-regulation, the errors
of the past need not be repeated by environmen-
tal managers of today. With better knowledge of
the risks posed by certain species, the goodwill of
all stakeholders and much hard work, there is no
reason why further potentially invasive species
should be introduced for the purposes of environ-
mental rehabilitation.
8 http://www.cabi.org/isc
9 http://www.europe-aliens.org/

21
GENET IC CONS IDERATI ONS IN ECO SYSTEM RE STORATI ON USING N ATIVE TRE E SPECIE S
References
Carruthers, J., Robin, L., Hattingh, J.P., Kull, C.A.,
Rangan, H. & van Wilgen, B.W. 2011. A native at
home and abroad: the history, politics, ethics and
aesthetics of acacias. Divers. Distrib., 17: 810–821.
Heywood, V. & Brunel, S. 2011. Code of conduct on
horticulture and invasive alien plants. Convention
on the Conservation of European Wildlife and
Natural Habitats (Bern Convention). Nature and
environment, no. 162. Strasbourg, France, Council
of Europe Publishing (available at: http://www.coe.
int/t/dg4/cultureheritage/nature/bern/ias/Documents/
Publication_Code_en.pdf). Accessed 22 January
2013.
Hulme, P.E. 2007. Biological invasions in Europe: drivers,
pressures, states, impacts and responses. In R.E.
Hester & R.M. Harrison, eds. Biodiversity under
threat, pp. 55-79. Issues in Environmental Science
and Technology 25. Cambridge, UK, Royal Society of
Chemistry.
Kowarik, I. & Säumel, I. 2007. Biological flora of
Central Europe: Ailanthus altissima (Mill.) Swingle.
Perspect. Plant Ecol. Evol. Syst., 8: 207–237.
Randal, R.P. 2002. A global compendium of weeds.
Meredith, Victoria, Australia, R.G. and F.J.
Richardson.

Part 2
THEORETICAL AND
PRACTICAL ISSUES
IN ECOSYSTEM
RESTORATION

GENET IC CONS IDERATI ONS IN ECO SYSTEM RE STORATI ON USING N ATIVE TRE E SPECIE S
25
Part 2 presents issues that should be considered
in all restoration efforts, irrespective of the local
context and the specific methods used. Building
on theoretical understanding of genetic pro-
cesses, the authors discuss how selection, ge-
netic drift and gene flow can affect outcomes
of restoration efforts. Local forest remnants are
widely considered to be ideal sources of propa-
gation material because they are assumed to be
well adapted to local conditions as a result of mil-
lennia of natural selection. However, it is often
overlooked that the remnant forests may be too
small to sustain viable populations, and may suf-
fer from genetic drift that results in random loss
of diversity (Chapter 2, Insight 4: Historical ge-
netic contamination in pedunculate oak (Quercus
robur L.) may favour adaptation and Chapter 4).
Gene flow through pollen and seed dispersal can
counteract negative implications of small popu-
lations (Chapter 5). Transferring genetic material
over longer distances may, however, threaten
indigenous genetic diversity and result in a loss
of local adaptations (Chapter 6 and Chapter 3).
However, such long distance transfers may be
beneficial in certain circumstances (see Insight 4:
Historical genetic contamination in pedunculate
oak (Quercus robur L.) may favour adaptation). In
most cases, little is known about the extent and
distribution of genetic diversity of tree species
used in restoration. Rules of thumb may exist for
collecting and transferring propagation material
in such cases, although those remain little studied
in practice (Chapter 7).
The introduction to the theoretical concepts
is followed by presentation of examples of their
practical application and constraints faced in res-
toration efforts. Various types of propagation
materials are discussed and guidance is provid-
ed on choosing suitable types for local contexts
(Chapter8). Considering the current proliferation
of restoration efforts and the simultaneous deg-
radation of natural tree populations of many spe-
cies, little attention is usually given to the sustain-
able sourcing of massive amounts of propagation
material (see Chapter8 and Insight6: Seed avail-
ability: a case study). Seed banks are effective and
often-overlooked sources of material for those
species that can easily be stored as seed (Insight7:
The role of seed banks in habitat restoration).
Traditional ecological knowledge held by local
and indigenous communities can be a valuable
source of information on suitable tree propa-
gation and management practices, not least be-
cause it has played an important role in shaping
tree diversity for hundreds or thousands of years
in many areas (Chapter10). Finally, restoration ef-
forts should not be planned in isolation but must
carefully consider the local landscape context,
recognizing and appreciating the needs and pri-
orities of the various interest groups (Chapter11).

GENET IC CONS IDERATI ONS IN ECO SYSTEM RE STORATI ON USING N ATIVE TRE E SPECIE S
27
Diverse biological, cultural, environmental and
socioeconomic conditions across the world de-
mand diverse approaches to forest or habitat
restoration and sustainable farming. Trees are a
vital component of many farming systems, while
a range of agroforestry systems have the poten-
tial to conserve native species as well as to diver-
sify and improve the production and income of
resource-poor farmers. Although native species
are usually favoured in tree planting for for-
est or habitat restoration or by local people on
farms, often only a limited range of management
options and tree species (often exotics) are pro-
moted.
Tree planting depends on a ready supply of
germplasm (seeds or vegetative material) of the
chosen species, which in turn requires consid-
eration of what is the best or most appropriate
source of seed. Inevitably the choice of seed source
should be influenced by the objective of planting
(e.g. for restoration or production, future adapt-
ability or past adaptation) and the risks associated
with particular seed sources (e.g. loss of adapta-
tion, outbreeding depression, loss of diversity, ge-
netic bottlenecks or contamination of native gene
pools). Choice of seed source, both in terms of its
location and its composition, can have important
consequences for the immediate success and for
the long-term viability of plantings.
Many tree species are outbreeding and gener-
ally carry a heavy genetic load of deleterious reces-
sive alleles. This means that inbreeding, in particu-
lar selfing, can have negative impacts, including
reduced seed set and survival resulting in poorer
regeneration, progeny with slower growth rates
and lower productivity, limited environmental tol-
erance and increased susceptibility to pests or dis-
eases. Consequently, the use of genetically diverse
germplasm is vital if plantings are to be productive,
viable and resilient. Intraspecific genetic diversity
may, however, be limited by several factors relat-
ed to the sourcing of seed. For example, farmers,
nursery managers and commercial collectors may
collect seed from only a few trees as this requires
less effort than collecting from many trees; how-
ever, this captures only a small amount of the vari-
ability present. In addition, variability in fertility
between trees can contribute to a rapid accumula-
tion of relatedness and inbreeding in subsequent
generations. Genetic issues can also be of particular
concern for nursery material, where inbred mate-
rial may survive benign nursery conditions but be
genetically compromised for survival and growth
when planted out in the wider environment.
Chapter 2
Seed provenance for restoration
and management: conserving
evolutionary potential and utility
Linda Broadhurst1 and David Boshier2,3
1 CSIRO Plant Industry, Australia
2 Department of Plant Sciences, University of Oxford, United Kingdom
3 Bioversity International, Italy

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2.1. Local versus non-local seed
Many guidelines for sourcing seed to restore
plant populations and communities advocate the
use of local seed under the premise that this will
be better adapted to local conditions and deliver
superior outcomes through improved survival
and growth (Broadhurst et al., 2008 and refer-
ences therein). Apart from the possibility of non-
local seed being maladapted to local conditions,
using seed collected close to a restoration site
is also predicted to prevent negative outcomes,
such as intraspecific hybridization (potentially)
resulting in outbreeding depression, superior in-
troduced genotypes becoming invasive and im-
pacts on associated organisms such as bud burst
occurring prior to herbivore emergence, and to
help maintain a range of biotic interactions with
pollinators and pathogens (Linhart and Grant,
1996; Jones, Hayes and Sackville Hamilton, 2001;
Cunningham et al., 2005; Vander Mijnsbrugge,
Bischoff and Smith, 2010). Although the impor-
tance of local provenance in habitat conserva-
tion and restoration remains contentious (e.g.
Sackville Hamilton, 2001; Wilkinson, 2001), the
concept is easy to understand and the message
is therefore attractive and easy to “sell” (see, for
example, www.floralocale.org). Hence the Gen-
eral guidelines for the sustainable management
of forests in Europe (MCPFE, 1993) state that
“native species and local provenances should be
preferred where appropriate.” Forest certifica-
tion and timber labelling standards also require
action to conserve genetic diversity and to use lo-
cal provenances (e.g. PEFC, 2010; UKWAS, 2007).
Grants for tree planting often require the use of
local material, although this may depend on the
purpose of planting (e.g. Forestry Commission,
2003).
Despite such requirements to source seed lo-
cally, many guidelines provide little direction
as to how this should be evaluated (Broadhurst
et al., 2008), with practitioners often interpret-
ing guidelines in a spatial context at a range of
scales (e.g. as small as a particular farm or wood
to as large as a country). To fully evaluate the
superiority of local seed requires complex experi-
ments and long-term monitoring that go beyond
early effects on germination and growth, and
that are beyond the scope of most restoration
projects. Consequently, there is often little empir-
ical evidence for deciding how local a seed source
should be. Should seed come from the same
wood, the same watershed, the same county or
the same country? Is geographical or ecologi-
cal distance more important (e.g. Montalvo and
Ellstrand, 2000)? With limited information about
the extent and scale of adaptive variation in na-
tive trees, discussion about suitable seed sources
often emphasizes “local” in a very narrow sense
or within political boundaries, rather than being
based on sound evidence of the scale over which
adaptation occurs.
The requirement to use locally collected seed
has been given such precedence that restoration
projects have occasionally been abandoned be-
cause of a lack of appropriate local seed sources
(Wilkinson, 2001). Use of native species in both
restoration and on farms has also been limited by
a lack of basic information on seed storage and
germination and establishment methods; a reflec-
tion of the historical emphasis on plantation for-
estry with a limited range of exotic species.
2.2. Basic concepts and theory
It is worth considering some basic concepts to ap-
preciate to what extent and at what scale local
adaptation may apply. The forces of natural se-
lection may vary in space, resulting in genotype
× environment interactions for fitness. In the ab-
sence of other forces and constraints, such diver-
gent selection should cause each local population
to evolve traits that provide an advantage under
its local environmental conditions (i.e. its habitat),
regardless of the consequences of these traits for
fitness in other habitats. What should result, in
the absence of other forces and constraints, is
a pattern in which genotypes of a population
would have on average a higher relative fitness
in their local habitat than genotypes from other

GENET IC CONS IDERATI ONS IN ECO SYSTEM RE STORATI ON USING N ATIVE TRE E SPECIE S
29
habitats. This pattern and process leading to it is
local adaptation (Williams, 1966).
However, local adaptation may be hindered by
gene flow, confounded by genetic drift, opposed
by natural selection as a result of temporal envi-
ronmental variability and constrained by a lack of
genetic variation or by the genetic architecture of
underlying traits. Thus, although divergent natu-
ral selection is the driving force, these other forc-
es, in particular gene flow, are integral aspects of
the process of local adaptation. Owing to such
forces, local adaptation is not a necessary out-
come of evolution under spatially divergent se-
lection (Kawecki and Ebert, 2004). Environmental
heterogeneity also favours the evolution of adap-
tive phenotypic plasticity. Where there are no
costs of and constraints on plasticity, a genotype
that produces a locally optimal phenotype in each
habitat would become fixed in all populations.
Adaptive phenotypic plasticity would lead to
adaptive phenotypic differentiation, but without
underlying genetic differentiation. Lack of plas-
ticity is thus a prerequisite for local adaptation.
In summary, factors predicted to promote local
adaptation include: low gene flow (i.e. restricted
pollen or seed dispersal, or strong habitat fidel-
ity), strong selection against genotypes optimally
adapted to other habitats but moderate selection
against intermediate genotypes (most likely un-
der moderate differences between habitats with
respect to traits under selection), little temporal
variation in the forces of selection, small differ-
ences between habitats in size and quality (e.g.
the amount of resources) and costs of or con-
straints on adaptive plasticity.
2.3. Historical perspective of local
adaptation
The extent to which observed morphological and
growth differences in plants are under genetic
control and related to the environment in which
a population occurs, has formed fertile ground for
research. Linnaeus reported as early as 1759 that
yew trees brought to Scandinavia from France
were less winter hardy than indigenous Swed-
ish yews. In his classical research, Turesson (1922)
studied populations of several herbaceous species
in transplant common garden experiments, dem-
onstrating the widespread occurrence of intraspe-
cific, habitat-related genetic variation and intro-
ducing the term “genecology.” Clausen, Keck and
Hiesey (1940) extended study of the expression of
population adaptation to environmental differ-
ences by using climatically different sites over a
range of altitudes. Subsequent research has shown
that such genetically related adaptive variation is
widespread in herbaceous species with low levels
of gene flow under strong selection pressures (see
summary in Briggs and Walters, 1997). There are
many key differences between herbaceous plants
and trees, where long life cycles, wide distribu-
tions and extensive gene flow (pollen and seed
dispersal) would tend to suggest more extensive
scales and patterns of adaptation, with differ-
ences most likely to occur at the geographic and
altitudinal extremes of species ranges.
2.4. The scale of local adaptation
in trees: how local should a
seed source be?
Evidence for strong local adaptation effects, es-
pecially in trees, remains mixed and such adap-
tation is very difficult to predict (Ennos, Worrell
and Malcolm, 1998; Montalvo and Ellstrand, 2000;
Joshi et al., 2001; Hufford and Mazer, 2003; Bis-
choff et al., 2006; Leimu and Fischer, 2008). Prov-
enance and progeny field trials have shown that
while genotype × environment interaction occurs
in many tree species, this may not be expressed
as a home-site advantage (i.e. provenance perfor-
mance is unstable across sites, but not as a result
of greater fitness of local seed source). Geograph-
ical proximity may be a poor indicator of adaptive
fitness (e.g. Betula spp.; Blackburn and Brown,
1988) and also stability, with some provenances
that show stable performance across sites origi-
nating from sites adjacent to unstable performers
(e.g. Kleinschmit et al., 1996).

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The northern hemisphere forestry literature
suggests that latitudinal or altitudinal gradi-
ents, or both, can be important for detecting the
scale of local adaptation, but that other factors
such as habitat, rainfall and topographical dif-
ferences can also be significant (Ennos, Worrell
and Malcolm, 1998). There is evidence for adap-
tive variation over reasonably short distances in a
number of conifer tree species in western North
America, owing to features such as aspect and al-
titude (e.g. Adams and Campbell, 1981; Sorensen,
1994). This is particularly marked in areas with
oceanic climates, where environmental gradients
are much steeper than in more continental sites.
Field, greenhouse and laboratory studies on co-
nifer species in the northwestern United States
show that a significant proportion (typically 25–
45 percent) of the genetic variation within popu-
lations is accounted for by climatic (e.g. rainfall
and temperature) or location (e.g. latitude, alti-
tude, slope aspect, distance from ocean) variables
that reflect environmental factors specific to each
location. There are often differences between
provenances from warmer and colder climates,
the former showing adaptation to the longer
growing season in lower latitudes but suffering
from early or late frosts when moved too far into
higher latitudes. The degree of risk in transplant-
ing across a species’ distribution is correlated
more with environmental changes than with
the geographical distance moved (Adams and
Campbell, 1981; see Insight 5). This suggests that
habitat matching may be a more useful means of
determining where seed should be sourced than
would be an arbitrary distance from the site to be
restored. Provenance trials of a number of tropi-
cal tree species show that most morphological ge-
netic variation occurs within rather than between
provenances. In most of the species studied, rank-
ing reversals (adaptation) or significant geno-
type × environment interactions only occur with
large environmental site differences (e.g. dry vs
wet zones, alkaline vs acidic soils). Unfortunately,
almost nothing is currently known about local
adaptation in temperate southern hemisphere
species.
Currently there are too few studies from too
few regions of the world to allow for predictions
regarding the scale and importance of local ad-
aptation for the myriad of life-history traits and
evolutionary histories of tree species that re-
quire restoration. For example, Leimu and Fischer
(2008) used only 32 species in their local adapta-
tion meta-analysis, none of which were tree spe-
cies. There is also a need for reciprocal transplant
experiments (RTEs), which test the fitness of
“home” and “away” genotypes within the sites
from which the genotypes originate (Primack and
Kang, 1989) and can mimic natural regeneration
by establishing seedlings in a forest at close spac-
ings to encourage early competition and with
minimal intervention (e.g. little or no weeding).
Germplasm selected and tested in forestry trials
or plantations for growth, form and other com-
mercial criteria may be less suited to the more
competitive environment of semi-natural forests
and restoration.
The scale over which species show adaptation
to their environment depends on the degree of
habitat heterogeneity, in particular the specific
habitat characteristics that affect a species, and
the interaction with gene flow. Dispersal levels
may be a useful high-level predictor of the im-
portance of local adaptation, under the premise
that species with long-range gene flow are less
likely to generate strong local adaptation, where-
as restricted gene flow is more likely to generate
genotypes adapted to their local environment.
Extensive gene flow in widely distributed tree
species suggests that local adaptation over a small
geographic scale is unlikely unless selection forces
are very strong.
2.5. Are non-local seed sources
ever appropriate?
In highly modified or degraded landscapes, us-
ing non-local seed may be entirely appropriate
or indeed the only option for restoration. Miti-
gating the negative physical effects associated
with vegetation removal, such as loss of topsoil,

GENET IC CONS IDERATI ONS IN ECO SYSTEM RE STORATI ON USING N ATIVE TRE E SPECIE S
31
altered hydrological flows or increased nutrient
loads, may require specific germplasm that is able
to cope with these conditions. For example, sa-
line scalds in southern Australia that developed
following the removal of deep-rooted perennials
are not generally amenable to restoration using
local species, let alone local seed. In these cases,
planting saline-tolerant varieties of other species
may be the only option to prevent further degra-
dation of valuable agricultural land. The loss of
diversity at genes of major effect may also require
sourcing of seed from non-local populations. For
example, small populations of self-incompatible
plants can be mate-limited if diversity in the in-
compatibility locus is low, requiring seed from
beyond the local area to introduce new mating
types. However, impacts on local species and
communities that may arise from using non-local
seed need to be considered carefully, preferably
prior to restoration and using an appropriate risk
management framework (Byrne, Stone and Mil-
lar, 2011). This should also include analysis of the
risk of not undertaking restoration and allowing
landscape degradation and biodiversity loss to
continue.
2.6. Local seed sources may not
produce restoration-quality
seed
Habitat fragmentation remains a major threat
to biodiversity worldwide through the loss of
populations and consequent altered biotic and
abiotic processes (Bakker and Berendse, 1999;
Eriksson and Ehrlen, 2001; Hobbs and Yates,
2003; Lienert, 2004). Unfortunately, some re-
gions of the world have now reached a tipping
point, such that whole biomes may be in danger
of collapse (Hoekstra et al., 2005). The most im-
mediate consequence of fragmentation for use
of native species in restoration and farm systems
is limitations to seed supply following the loss of
individuals and populations. But several nega-
tive genetic and demographic effects associated
with fragmentation can also have an impact on
the quantity and quality of seed available. The re-
moval of trees and populations from landscapes
directly reduces genetic diversity, most of which
is irreplaceable since genetic mutations accumu-
late slowly over long evolutionary periods (i.e.
tens of thousands to millions of generations).
Diversity is further eroded in small populations
by drift resulting from random sampling within
populations, as well as inbreeding as a result of
trees in remnant populations often being more
highly related than those in larger populations
(Barrett and Kohn, 1991; Ellstrand and Elam,
1993). Reduced fitness and productivity are com-
monly documented effects associated with ge-
netic erosion and inbreeding, both of which can
have an impact on a population’s ability to persist
in stressful situations or changing environments
(Frankham, Ballou and Briscoe, 2002; Hughes et
al., 2008). Other negative outcomes include poor
reproductive success, smaller, poor-quality plants
and increased susceptibility to pests and patho-
gens (Lienert, 2004 and references therein). Over
time, this exposes small populations to decline
through recruitment failure (Figure 2.1) and lim-
its their utility as appropriate seed sources for
restoration. Limited seed supply and poor-quality
seed are two major impediments to the successful
planting of native species and restoration of na-
tive vegetation, especially at the landscape level.
Worldwide analyses of fragmentation impacts
on plant reproduction indicate that some spe-
cies are shifting towards selfing (Aguilar et al.,
2006; Aguilar et al., 2008; Eckert et al., 2010), but
how this translates to seed production depends
largely on reproductive strategy. For example,
species that cannot self or mate with close rela-
tives (self-incompatible) will not produce seed
unless pollinated by distantly- or non-related
trees and small, self-incompatible populations
are often characterized by reduced seed produc-
tion, severely limiting quantities available for
restoration. In contrast, species that can self and
mate with close relatives (self-compatible) con-
tinue to produce seed but this is often less fit, be-
ing smaller, slower to germinate and with poorer
survival (Buza, Young and Thrall, 2000; Young

THE STATE OF THE WORLD’S FOREST GENETIC RESOURCES – THEMATIC STUDY
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et al., 2000; Mathiasen, Rovere and Premoli,
2007). Restoration using this seed is therefore
likely to produce poorer results than expected
and over the long term is less likely to develop
into a self-sustaining population. A requirement
that only local seed be used for restoration can
drive practitioners to use seed from small, inbred
populations that are unlikely to produce posi-
tive long-term restoration outcomes, but rather
create more small, inbred populations, with lim-
ited long-term persistence. One consideration is
that populations restored with a narrow genetic
base may be limited in their ability to respond
to the rapid predicted shifts in climatic variables
(Helenurm, 1998).
2.7. Adaptation and climate
change
There are theoretical reasons that underlie ob-
served patterns of adaptive variation; these also
suggest that many tree species over large areas
may fail to show local adaptation at a very narrow
scale. The prevalence of extensive gene flow may
counteract selection, while the temporal varia-
tion in selective forces that trees experience (e.g.
yearly variation in temperature, frosts or rainfall)
is likely to have a stabilizing effect rather than the
directional selection that would lead to highly lo-
calized adaptation. Given the long life of trees,
the environment is also likely to have altered over
the lifespan of a tree or only a few generations,
such that a particular site no longer experiences
the same conditions under which the trees origi-
nally evolved. These factors explain the relative
lack of adaptation over short distances in many
tree species. Temporal variation in environment
is particularly important for trees, not only with
respect to past adaptation but also in the context
of predicted climate change (e.g. Broadmeadow,
Ray and Samuel, 2005), and thus undue emphasis
on local seed sources may also cause problems.
Figure 2.1.
Simplified representation of how low genetic
diversity and inbreeding can impact on plant
population persistence and seed production in
small populations of plants

GENET IC CONS IDERATI ONS IN ECO SYSTEM RE STORATI ON USING N ATIVE TRE E SPECIE S
33
2.8. Benefits of using larger but
more distant seed sources
Using large populations as primary seed sources
for restoration not only ensures that seed quality
will be higher but also that larger quantities are
available. In many cases a population of 100–200
plants would be large enough to provide good-
quality seed, but more than 400 plants may be
needed for some species. A good restoration out-
come is also more likely if the habitat of the site to
be restored is matched as closely as possible with
that of the nearest large population. Seed from
these large populations could be augmented
with that collected from small populations closer
to the restoration site to capture any useful ge-
netic diversity they may contain (Broadhurst et
al., 2008). To capture as much genetic diversity as
possible from large populations, as many plants
as practically possible should be sampled broadly
across the site, collecting from a range of cohorts,
from various sides of plant canopies without dis-
rupting biotic associations that also rely on this
seed. Breed et al. (2012) reviewed such strate-
gies for sourcing restoration seed (Box 2.1) and
summarized their suitability for mitigating cli-
mate change and habitat fragmentation impacts
(Table 2.1).
The mixing of introduced and native germplasm
raises the issue of outbreeding depression; the
potential problem of reduced vigour as adapted
gene complexes are broken up or the proportion
of locally adapted alleles is reduced. As with local
adaptation, evidence for outbreeding depression
comes from herbaceous species that show highly
localized adaptation (see Hufford and Mazer,
2003) and there is little evidence for its occur-
rence in trees at distances of less than hundreds
of kilometres (e.g. Hardner et al., 1998, Boshier
and Billingham, 2000). For example large-scale
importation of cheap seed from Eastern Europe
has shown problems of maladaptation in Britain.
But it seems unlikely that use of material from
maritime France that matches future climate pre-
dictions (Broadmeadow et al., 2005) and of similar
phylogeographic origins will face such problems,
nor lead to outbreeding depression problems on
introgression with British material.
2.9. Conclusions
Any genetic conservation policy for native trees
should aim at conserving the evolutionary po-
tential of their populations, rather than at pre-
serving a particular genetic structure and status.
The extent and scale of local adaptation in many
tree populations, and thus its practical impor-
tance to restoration efforts, remain in doubt.
While there is a need for more field trials, both
of the traditional provenance or progeny and
RTE types, to provide more information on the
scale of adaptation, planting of native trees con-
tinues apace and demand for seed from certified
sources increases. There is good evidence to sug-
gest that emphasis on a very restricted view of
what is “local” will not lead to better-adapted
tree populations and is more likely to lead to use
of stock of limited genetic diversity than would a
broader approach.
It has been argued that, given the lack of ex-
tensive trials investigating adaptive variation
in native tree populations, the precautionary
principle should be adopted in sourcing germ-
plasm for planting trees (e.g. Flora Locale, 1999;
UKWAS, 2007). This is expressed as the use of lo-
cal seed, although the subsequent view of what
constitutes the local population varies from a
particular forest to large seed zones. However,
given current evidence for trees, i.e. clear dan-
gers from inbreeding and loss of genetic diver-
sity, with extensive gene flow and adaptation at
a broad scale, it seems more logical to apply the
precautionary principle in terms of ensuring the
use of genetically diverse material with the ca-
pacity to adapt to current and future conditions.

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Strict local provenancing: collecting seeds
from plants that are located physically very close
to the revegetation site (e.g. Natural England,
United Kingdom: 5 miles; Western Australian Forest
Management Plan 2004–2014: 15 km).
Relaxed local provenancing: collecting seeds
with a bias towards certain ecological criteria, and
avoiding small population fragments (e.g. Australian
FloraBank: soil type, altitude and climate).
Predictive provenancing (Sgrò, Lowe and
Hoffmann, 2011): use of naturally occurring
genotypes experimentally determined to be adapted
to projected conditions. This technique requires
data on local adaptation of target species (e.g. by
reciprocal transplant experiments), as well as climate
projections for these species at a revegetation site
(e.g. by bioclimatic modelling).
Composite provenancing (Broadhurst
et al.
,
2008): collecting a mixture of seed that attempts
to mimic natural gene-flow dynamics. For example,
recommended proportions of seed collected from
local, intermediate and distant distance-classes could
be determined by estimating the pollen dispersal
kernel for target species.
Admixture provenancing (Breed
et al.
, 2012):
collecting seed only from large populations, focusing
on capturing a wide selection of genotypes from a
diversity of environments with no spatial bias towards
the revegetation site. These seeds are then admixed
for sowing or planting, generating a population
with a mixture of genotypes from a wide array of
provenances.
References
Breed, M.F., Stead, M.G., Ottewell, K.M., Gardner,
M.G. & Lowe, A.J. 2012. Which provenance and where?
Seed sourcing strategies for revegetation in a changing
environment.
Conserv. Genet.
, November 2012. doi:
10.1007/s10592-012-0425-z.
Broadhurst, L.M., Lowe, A., Coates, D.J., Cunningham,
S.A., McDonald, M., Vesk, P.A. & Yates, C. 2008. Seed
supply for broadscale restoration: maximising evolutionary
potential.
Evol. Appl.
, 1: 587–597.
Sgrò, CM, Lowe, A.J. & Hoffmann, A.A. 2011. Building
evolutionary resilience for conserving biodiversity under
climate change.
Evol. Appl.
, 4: 326–337.
Source: Breed, M.F., Stead, M.G., Ottewell, K.M., Gardner, M.G.
& Lowe, A.J. 2012. Which provenance and where? Seed sourcing
strategies for revegetation in a changing environment.
Conserv.
Genet.
, November 2012. doi: 10.1007/s10592-012-0425-z.
Box 2.1.
Summary of alternative strategies for sourcing seed for restoration
TABLE 2.1.
Suitability of provenancing techniques under climate change with habitat fragmentation
Provenancing
technique Adaptive
potential
benefits
Genetic rescue
benefits Low genetic
load Suitable
with high
uncertainty
Economically
efficient Likely
population
success
Strict local x*
Relaxed local x*
Predictive x x x†
Composite x x x x
Admixture x x x x x
* May experience high failure rates, negating the economic benefit.
† Benefit rests on successfully matching genotype fitness with future conditions.
Source: Breed
et al.
(2012).

GENET IC CONS IDERATI ONS IN ECO SYSTEM RE STORATI ON USING N ATIVE TRE E SPECIE S
35
Current threats to the maintenance of genetic
diversity come principally from poor practice in
seed collection; undue emphasis on restricting
the area of collection or poor instruction of col-
lectors can limit the number of trees and hence
genetic diversity sampled, leading to the estab-
lishment of trees with restricted genetic diversity
and limited future adaptive potential. A study
of the few remnant ash and rowan trees in the
denuded Carrifran valley in southern Scotland
showed that large amounts of genetic diversity
are maintained, making them suitable for use in
restoration despite their highly fragmented na-
ture (Bacles, Lowe and Ennos, 2004). In contrast,
some of the locally sourced material planted as
part of the Carrifran wildwood restoration pro-
ject was shown to be low in genetic diversity
(Kettle, 2001), presumably because of poor col-
lection practices, which impose limitations on the
future potential of the population.
It is disturbing to contemplate that some of
the poorest seed sources exist in the very regions
where restoration is most needed and that con-
tinued requirements for using local seed simply
perpetuate the problem. In many regions of the
world with fragmented forest populations, being
able to reliably source large volumes of quality
seed of native species can be challenging. Not
only are there fewer populations from which seed
can be collected, but fragmentation has split con-
tinuous populations into much smaller and more
isolated remnants, which can impact the quality
and quantity of seed (e.g. Lowe et al., 2005). The
implications from this are that (i) remnant vegeta-
tion contains all of the diversity that is left that
is extremely valuable, and (ii) it is important that
most of the diversity that does remain is used
for restoration (i.e. avoid over-collection from a
few populations). In regions where fragmenta-
tion is high, should the rules for using local seed
change? Can we afford the luxury of being too
restrictive about seed sources? Are small, frag-
mented and probably inbred populations so pre-
cious that we cannot source seed from beyond
our comfort zone?
References
Adams, T. & Campbell, R.K. 1981. Genetic adaptation and
seed source specificity. In S.D. Hobbs & O.T. Helgerson,
eds. Reforestation of skeletal soils: Proceedings of a
workshop held November 17–19, 1981, Medford,
Oregon, pp. 78–85. Corvallis, OR, USA, Forest Research
Laboratory, Oregon State University.
Aguilar, R., Ashworth, L., Galetto, L. & Aizen, M.A.
2006. Plant reproduction susceptibility to habitat
fragmentation: review and synthesis through a meta-
analysis. Ecol. Lett., 9: 968–980.
Aguilar, R., Quesada, M., Ashworth, L., Herrerias-Diego,
Y. & Lobo J. 2008. Genetic consequences of habitat
fragmentation in plant populations: susceptible signals
in plant traits and methodological approaches. Mol.
Ecol., 17: 5177–5188.
Bacles, C.F.E, Lowe, A.J. & Ennos, R.A. 2004. Genetic ef-
fects of chronic habitat fragmentation on tree species:
the case of Sorbus aucuparia remnants in a deforested
Scottish landscape. Mol. Ecol., 13: 574–583.
Bakker, J.P. & Berendse F. 1999. Constraints in the restora-
tion of ecological diversity in grassland and heathland
communities. Trends Ecol. Evol., 14: 63–68.
Barrett, S.C.H. & Kohn, J.R. 1991. Genetic and evolution-
ary consequences of small population size in plants:
implications for conservation. In D.A. Falk & K.E.
Holsinger, eds. Genetics and conservation of rare
plants, pp. 3–30. New York, USA, Oxford University
Press.
Bischoff, A., Cremieux, L., Smilauerova, M., Lawson,
C.S., Mortimer, S.R., Dolezal, J., Lanta, V., Edwards,
A.R., Brook, A.J., Macel, M., Leps, J., Steinger, T. &
Müller-Schärer, H. 2006. Detecting local adaptation in
widespread grassland species – the importance of scale
and local plant community. J. Ecol., 94: 1130–1142.
Blackburn, P. & Brown, I.R. 1988. Some effects of expo-
sure and frost on selected birch progenies. Forestry, 61:
219–234.
Breed, M.F., Stead, M.G., Ottewell, K.M., Gardner, M.G.
& Lowe, A.J. 2012. Which provenance and where?
Seed sourcing strategies for revegetation in a changing

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environment. Conserv. Genet., November 2012. doi:
10.1007/s10592-012-0425-z.
Briggs, D. & Walters, S.M. 1997. Plant variation and evolu-
tion (3rd ed.). Cambridge, UK, Cambridge University
Press.
Broadhurst, L.M., Lowe, A., Coates, D.J., Cunningham,
S.A., McDonald, M., Vesk, P.A. & Yates, C. 2008.
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A population of trees or shrubs is autochthonous
if it has regenerated naturally since its arrival af-
ter the last glaciation; any human intervention in
breeding should have occurred with strictly local
material only. For long-lived species such as trees,
autochthony assumes a continuous presence at a
given site since post-glacial immigration (Klein-
schmit, Kownatzki and Gegorius, 2004). This im-
plies a continuity of local genetic diversity after
thousands of years of natural selection. Trees and
shrubs that belong to native species but are im-
ported from other climatic zones or geographic
regions are not autochthonous.
After many years of neglect, the use of na-
tive species in afforestation and landscape pro-
grammes is gaining importance all over the
world, based on the basic underlying ecological
principle that native species and genotypes will
be well adapted to local conditions and will have
co-evolved with other components of local forest
ecosystems. This has led to massive plantations
of indigenous tree and shrub species in Western
Europe, not only in forestry but also for native
woodland restoration and other landscape plant-
ings, such as thickets, wooded banks and hedge-
rows. A major challenge is to ensure that plant-
ing material used represents the genetic variation
and diversity within native species. Several initia-
tives have been developed in various European
countries to promote the use of locally sourced
seeds for the production of planting stock (e.g.
Belgium: Vander Mijnsbrugge, Cox and Van
Slycken, 2005; Germany: Kleinschmit, Leinemann
and Hosius, 2008; Denmark: Kjaer et al., 2009).
Here we describe in detail the programme on the
production of autochthonous planting stock in
Flanders, Belgium.
3.1. Why should autochthonous
diversity be protected?
There is a high demand for “native” planting
stock in Flanders, Belgium, and to a broader ex-
tent in many Western European countries. The
use of native planting material is promoted by
a wide range of public organizations. However,
planting stock of native material in commercial
nurseries is largely not autochthonous. Seeds of
native species are often imported, originating
from foreign provenances, often in Eastern Eu-
ropean countries. This is especially true for shrub
species. For trees, in the European Union, Coun-
cil Directive 1999/105/EC of 22 December 1999
(Council of the European Union, 2000) regulates
the marketing and transport of forest reproduc-
tive material through an obligatory certification
system indicating the origin of the material (al-
though control in practice is not perfect). How-
ever, certification is not obligatory for shrubs, and
Chapter 3
Continuity of local genetic diversity
as an alternative to importing
foreign provenances
Kristine Vander Mijnsbrugge1,2
1 Research Institute for Nature and Forest, Belgium
2 Agency for Nature and Forest, Belgium

THE STATE OF THE WORLD’S FOREST GENETIC RESOURCES – THEMATIC STUDY
PART 2
40
shrub germplasm is commonly imported from
Eastern and Southern Europe where cheaper
seed is available. Nursery managers often do not
know or are not interested in the exact origin of
the seed they obtain. Tree seed may also be im-
ported when seed is not available from officially
approved sources or supplies are too limited to
meet requirements.
Introduction of non-local material can
have numerous negative consequences. Non-
autochthonous planting stock may be poorly
adapted to local growing conditions, which can
lead to negative consequences such as lower
fitness (e.g. McKay et al., 2005; Krauss and He,
2006; Edmands, 2007; Laikre et al., 2010; Vander
Mijnsbrugge, Bischoff and Smith, 2010). Problems
may only become evident many years after seem-
ingly successful establishment. Intraspecific hy-
bridization of local and introduced genotypes
may result in outbreeding depression, i.e. re-
duced fitness in subsequent generations, loss
of genetic diversity and loss of adaptation, and
less adapted characteristics can introgress into
the autochthonous populations. The introduc-
tion of non-local material may also have negative
effects on associated plant and animal species.
Imported hawthorn (Crataegus monogyna) has
been shown to flower several weeks earlier than
native hawthorn, potentially threatening the in-
sects and birds whose reproductive cycles are syn-
chronized with this event (Hubert and Cottrell,
2007). In addition, purity of the species can be
problematic in commercial planting stock. A ge-
netic study on commercially available hawthorn
in Flanders, grown from seeds imported from
Hungary, showed that it comprised a mixture of
C. monogyna and C. monogyna × C. rhipidophylla
(Debeer, 2006).
3.2. Inventory of autochthonous
woody plants
A simple way to ensure the continuity of local ge-
netic diversity is the production of autochthonous
planting stock. For this an overview is needed of
the remnant autochthonous populations still pre-
sent. A survey was conducted to locate remaining
autochthonous populations in Flanders, Belgium,
from 1997 to 2008. The evaluation of autoch-
thony in the field was conducted following the
methodology presented by Maes (1993). In short,
areas of woody vegetation that are indicated as
forest on historical maps are identified. Informa-
tion on flora, soil conditions and geomorphology
further refine the selection of potentially rel-
evant sites. In the field, the woody vegetation is
evaluated according to a set of criteria. The tree
or shrub must be a wild variety and old. No evi-
dence must be seen of plantation (e.g. trees in
lines). The site must be located within the natural
geographic range of the species, and the growth
conditions correspond to the ecological require-
ments of the species. The tree or shrub must be
present on similar sites in the surrounding area. A
variety of plants in the tree, shrub or herb layer is
indicative of undisturbed woodland and ancient
forests. If hedges or wooded banks have been
planted with locally sourced material the plants
can be considered autochthonous.
The findings show that autochthonous woody
plants have become seriously endangered in
Flanders, with only about 6 percent of the cur-
rent forest cover holding autochthonous woody
plants. Several causes for this loss of autoch-
thonous material are evident. Only 11 percent
of Flanders is now forested and what there is is
highly fragmented as a result of centuries of in-
tensive forest use. Small fields have been replaced
by large, open expanses of farmland, with the
consequent disappearance of wooded banks, old
hedges and small forests on farmland.
The inventory data (in Flemish) are accessible
on the internet (www.natuurenbos.be).
3.3. Producing autochthonous
planting stock
The Agency for Nature and Forest (ANB), under
the Flemish Forest Administration, has been col-
lecting seed from inventoried sites since 1998 to

GENET IC CONS IDERATI ONS IN ECO SYSTEM RE STORATI ON USING N ATIVE TRE E SPECIE S
41
produce autochthonous planting stock. Seed is
collected from natural populations present on
inventoried sites (so-called in situ collecting) fol-
lowing general guidelines for appropriate col-
lection methods. Sites adjacent to plantations of
the same species are omitted because of the risk
of cross-pollination from unknown provenances.
Seed is collected from at least 30 seed-bearing
plants per species within each region of prove-
nance. Region of provenance, a term commonly
used in forestry, is an area within which move-
ment of plant material will not negatively affect
the fitness of the populations in the long run.
In Flanders, the surveyed sites are mostly frag-
mented, small and are not managed for seed
production. Therefore, several sites must be vis-
ited to find 30 seed-bearing trees or shrubs for
every species. This implies a time-consuming and
costly effort. The Flemish legislation (Anony-
mous, 2003a), which follows Council Directive
1999/105/EC, allows mixing seed lots within a re-
gion of provenance. This practice guarantees a
good genetic variability in the derived planting
stock. A genetic study on sloe (Prunus spinosa) in
Flanders showed that old autochthonous hedges
dominated by sloe may show low within-popula-
tion genetic diversity. In this case, mixing of seed
lots from different autochthonous locations is
specifically advised (Vander Mijnsbrugge et al.,
in press.
Until now, the planting stock has been grown
in two government nurseries located in Koekelare
and in Brasschaat. However, the decision has been
taken to close them, mainly for financial reasons.
Future planting stock will be grown increasingly
in private nurseries under contract. The autoch-
thonous planting stock is used only in forests
owned by or managed by ANB. As seed collection,
growth and planting are all performed within the
forest administrative boundaries, no certification
or control system is involved.
Since 1998 seeds have been collected also by
public organizations called Regional Landscapes
(“Regionale Landschappen”) that are working to
protect and enhance the local authenticity of ru-
ral landscapes (Table 3.1). Here, all planting stock
is grown in private nurseries under a sales con-
tract. The seeds and derived planting stock are
not certified, and the work of the nursery is not
controlled by any official agency, necessitating a
relationship of trust between the client and the
nursery. Again, this autochthonous planting stock
is used in the Regional Landscapes’ own projects,
mainly landscape plantings such as hedgerows,
wooded banks, on farms, etc., and can also be
sold to local people.
Since 2004 seeds can be collected on invento-
ried sites that are officially approved as a seed
source, primarily under the category “source
identified” (as defined by the Council Directive
on the marketing of forest reproductive mate-
rial). At least 30 seed-bearing trees or shrubs of
the same species must be present on such sites,
with a good score for autochthony. There must
be no non-autochthonous plantations in the vi-
cinity. Autochthonous stands showing traits of sil-
vicultural value are approved under the category
“selected.” Five stands of Alnus glutinosa have
been given this designation. Private nurseries can
collect seeds from these officially approved seed
sources and obtain a certificate from an independ-
ent governmental control agency that proves the
origin of the seeds. The landowner of the col-
lection site can charge those wanting to collect
seeds, although in general private nurseries are
not willing to pay large sums. Major problems
faced by certified in situ collections are the reluc-
tance of landowners to agree to the designation
of a woody population on their property as an
official seed source, the laborious process of ap-
proval of the sites, the small number of sites that
meet the requirements for approval, and lack of
management for high seed production. A major
advantage of the system is that certified planting
stock becomes available to a broader public.
3.4. Seed orchards
Seed orchards hold many advantages over in situ
collecting. They produce large amounts of seed
and at the same time preserve the gene pool of

THE STATE OF THE WORLD’S FOREST GENETIC RESOURCES – THEMATIC STUDY
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42
TABLE 3.1.
Seeds and berries collected between 2006 and 2010 by Regional Landscapes (public organizations)
from autochthonous populations in Flanders
Species Fresh weight (kg)*
2006 2007 2008 2009 2010
Acer campestre
70.5 60.0 70.5 66.3 56.7
Alnus glutinosa
35.5 64.3 41.4 28.0 61.2
Carpinus betulus
116.3 159.3 11.5 141.4 88.0
Cornus sanguineum
26.0 39.8 26.0 11.3 6.4
Corylus avellana
10.2 229.3 26.5 56.7 71.9
Crataegus laevigata
4.7 11.6 4.0 2.8 0.9
Crataegus monogyna
111.8 44.1 59.8 107.6 103.8
Crataegus spp.
397.1 465.1 398.9 458.5 309.6
Euonymus europaeus
10.1 11.2 28.6 24.6 21.7
Fraxinus excelsior
105.3 2.5 62.4 5.6 226.4
Ilex aquifolium
18.1 5.0 5.5 6.0 5.9
Malus sylvestris
– 1.8 – – 0.3
Mespilus germanica
1.5 42.8 5.1 2.5 9.4
Prunus padus
2.4 0.1 0.7 7.0 6.0
Prunus spinosa
218.3 121.1 24.8 113.7 107.9
Quercus robur
608.0 422.0 33.3 390.4 460.9
Rhamnus cathartica
1.0 5.4 4.4 – –
Rhamnus frangula
13.7 20.4 34.0 51.9 56.0
Rosa arvensis
0.3 1.0 1.8 1.0 1.1
Rosa canina
26.2 16.5 35.7 11.3 19.3
Rosa corymbifera
– – 7.8 – –
Rosa spp.
1.0 1.4 – 5.0 5.4
Sambucus nigra
– – – 5.0 2.5
Sorbus aucuparia
109.5 67.0 39.1 135.4 66.9
Tilia cordata
– – – 1.5 –
Viburnum opulus
110.1 110.3 61.6 48.2 72.5
Total 1997.5 1901.9 983.1 1681.4 1760.7
* Uncleaned fresh weight.

GENET IC CONS IDERATI ONS IN ECO SYSTEM RE STORATI ON USING N ATIVE TRE E SPECIE S
43
the autochthonous populations from which the
plants in the orchard originate. A programme
initiated by the Research Institute for Nature
and Forest and ANB for the creation of autoch-
thonous seed orchards started in Flanders in
1999. Seed orchards have been established for all
woody species that are regularly or occasionally
planted. Basic material for these is collected at
the inventoried sites. The objective is to repre-
sent the genetic diversity of the autochthonous
populations present in a region of provenance.
There are four main regions of provenance in
Flanders, with an average area of 3000 km2. Thus,
theoretically, there should be four seed orchards
for every woody species for which planting stock
is desirable, one for each region of provenance.
In practice, the number of orchards established
differs for various reasons, such as the natural
distribution pattern. For example, the nutri-
ent poor soils in the north of Flanders (regions
of provenance “Kempen” [KEM] and “Vlaamse
Zandstreek” [VZA]) are characterized by a spec-
trum of species that differs from that found on
the more nutrient-rich soils in the south (regions
of provenance “Brabants District Oost” [BDO]
and “Brabants District West” [BDW]). Thus, for
example, seed orchards for Eonymus europaeus,
a species found on nutrient rich soil types, have
been established for only the BDO and BDW re-
gions. Few relict populations remain for some
rare and dispersed species such as Tilia cordata,
Ulmus laevis or Malus sylvestris, and as a result
orchards have been created using basic material
from the whole of Flanders. Similarly, orchards
for the whole of Flanders have been established
for seemingly abundant species but for which
autochthonous populations are rare, such as
Quercus petraea or Populus tremula.
The most clearly authenticated autochthonous
trees and shrubs are propagated, mainly vegeta-
tively, from geographically scattered sites within
the region of provenance. The use of vegetatively
propagated plants ensures that they are geneti-
cally identical to the parent tree and ensures
there is no pollution from non-autochthonous
sources. In evolutionary terms, only one genera-
tion of exchange of genetic information is missed.
The disadvantage is that vegetative propagation
is difficult and expensive, particularly for recalci-
trant genera such as Quercus. Experienced green-
house technicians are indispensable. Labour- and
cost-intensive in vitro techniques are not used.
For trees with economic importance, the orchard
clones can serve as parent material for breeding
in future. Every seed orchard contains a minimum
of 50 genotypes per species, collected from at
least five different sites, and up to four ramets
per genotype. An ideal seed orchard contains 200
plants. In addition, the aim is to duplicate each
seed orchard at another location within the re-
gion of provenance.
Once established, the autochthonous seed
orchards are officially approved as seed sources
(category “source identified”) and the seeds from
them can be certified. The first plantations date
from 2003 and planting is ongoing. The major-
ity of orchards are situated on land owned and
managed by ANB, while some have been estab-
lished on municipal land and land owned by
nature conservation organizations. By October
2011 a total of 14339 plants had been planted
in 90 seed orchards at 25 different locations in
Flanders (Table3.2). Shrub species in several or-
chards are fruiting and certified seeds are being
collected by private nurseries and a commercial
seed merchant (there is only one in Flanders). A
major problem facing the nurseries is the tech-
nical and administrative inefficiency of a large
number of small regions of provenance; other
European countries have fewer, larger regions of
provenance. Small countries tend to define small
regions of provenance, mainly because of the ab-
sence of a pan-European consensus on the proper
way to delineate them. The geographic scale of
local adaptation is difficult and time consuming
to measure for long-lived perennials.
3.5. Promotion of use
Flanders has a state-funded system for subsi-
dizing (re)forestation that promotes the use

THE STATE OF THE WORLD’S FOREST GENETIC RESOURCES – THEMATIC STUDY
PART 2
44
of autochthonous provenances (Anonymous,
2003b). A basic subsidy supports the use of na-
tive species, the amount of subsidy depending on
the choice of species (e.g. indigenous oaks receive
the highest subsidy). An additional subsidy, with
a fixed financial value, is granted for the use of
specific autochthonous provenances that are indi-
cated on the list of endorsed provenances, which
TABLE 3.2.
Number of individuals in the autochthonous orchards in Flanders by region of provenance
(October 2011)
Species BDO BDW VZA KEM Flanders Total
Acer campestre
– 198 – – – 198
Carpinus betulus
– – 157 – 303 460
Cornus sanguineum
– 509 – – – 509
Corylus avellana
73 299 205 225 – 802
Crataegus monogyna
572 759 629 – – 1960
Euonymus europaeus
268 362 – – – 630
Fraxinus excelsior
177 100 161 108 – 546
Juniperus communis
– – – 742 – 742
Malus sylvestris
– – – – 199 199
Mespilus germanica
– 216 218 – – 434
Populus tremula
– – – – 96 96
Prunus avium
– – – – 181 181
Prunus insititia
– – – – 20 20
Prunus padus
289 467 – 498 – 1254
Prunus spinosa
– 363 37 – – 400
Quercus petraea
– – – 194 – 194
Quercus robur
– – – 117 – 117
Rhamnus frangula
– 183 238 547 – 968
Rosa arvensis
– 131 – – – 131
Rosa canina
– 203 243 – – 446
Salix alba
76 – 126 – – 202
Sorbus aucuparia
– 315 175 442 – 932
Tilia cordata
– – – – 346 346
Tilia platyphyllos
– – – – 142 142
Ulmus laevis
– – – – 482 482
Viburnum opulus
541 540 405 462 – 1948
Total 1996 4645 2594 3335 1769 14 339
*BDO: Brabants District Oost; BDW: Brabants District West; VZA: Vlaamse Zandstreek; KEM: Kempen.

GENET IC CONS IDERATI ONS IN ECO SYSTEM RE STORATI ON USING N ATIVE TRE E SPECIE S
45
lists all officially approved autochthonous seed
sources and orchards. The list (in Flemish) is ac-
cessible on the internet (www.inbo.be). A major
drawback is that subsidies are only for (re)foresta-
tion, not any other landscape plantations such as
hedges or tree rows or wooded banks.
3.6. Discussion
The Flemish government has invested heavily
in production of autochthonous planting stock,
starting with a laborious inventory, followed by
both in situ collection of seeds and the establish-
ment of seed orchards for many native species,
both trees and shrubs. As a rough estimate, over
recent years about 1 million autochthonous plants
have been grown annually in both government
and private nurseries. The majority of the plants
are from seed collected in situ and grown under
sales contracts. However, officially approved seed
orchards are now starting to produce seed and
certified seed is becoming increasingly available
for all interested forest nurseries.
The programme now faces two issues. The first
concerns communication. When private owners
or public organizations buy planting stock their
decisions are influenced by price, and autochtho-
nous stock is more expensive (albeit sometimes
only slightly) than non-autochthonous planting
stock. As a result there is a tendency to purchase
non-autochthonous planting stock. Targeted
communication is needed to make all stakehold-
ers aware of the value of autochthonous prov-
enances, and the importance of the continuity of
local genetic diversity of autochthonous popula-
tions. A major challenge lies in providing a clear
explanation of the role and importance of genetic
diversity. People readily understand that low ge-
netic diversity leads to fitness problems related to
inbreeding, but do not realize that bringing dif-
ferentiated populations can have negative con-
sequences that may result in diminishing genetic
diversity and fitness.
The second major issue is control. Private nurser-
ies play a pivotal role in production of autochtho-
nous stock. However, their primary purpose is to
make a profit. Inevitably, some nursery managers
may be tempted to increase their profits by sell-
ing non-autochthonous stock as (more expensive)
autochthonous stock. Although genetic studies
can distinguish autochthonous from non-autoch-
thonous material, they require highly skilled staff
and are too expensive and time-consuming to use
as a general control mechanism. Thus, controls
during seed collection and growth in the nursery
are the major (general) tools at hand.
References
Anonymous. 2003a. 3 oktober 2003 – Besluit van de
Vlaamse regering betreffende de procedure tot
erkenning van bosbouwkundig uitgangsmateriaal
en het in de handel brengen van bosbouwkundig
teeltmateriaal. Belgian Law Gazette, 11 November:
54793–54824.
Anonymous. 2003b. 27 juni 2003 – Besluit van de
Vlaamse regering betreffende de subsidiëring van
beheerders van openbare en privé-bossen. Belgian
Law Gazette, 10 September: 45431–45500.
Council of the European Union.. 2000. Council
Directive 1999/105/EC of 22 December 1999 on the
marketing of forest reproductive material. Official
Journal of the European Communities, 15 January
2000, L11: 17–40.
Debeer, L. 2006. Studie van de genetische diversiteit in
het genus Crataegus (meidoorn): interspecifieke hy-
bridisatie en herkomstanalyse. University of Ghent,
Belgium. (Master’s thesis)
Edmands, S. 2007. Between a rock and a hard place:
evaluating the relative risks of inbreeding and out-
breeding for conservation and management. Mol.
Ecol., 16: 463–475.
Hubert, J. & Cottrell, J. 2007. The role of forest genetic
resources in helping British forests respond to
climate change. Edinburgh, UK, Forestry Commision
Scotland.

THE STATE OF THE WORLD’S FOREST GENETIC RESOURCES – THEMATIC STUDY
PART 2
46
Kjær, E.D., Hansen, L.N., Graudal, L., Olrik, D.C.,
Ditlevsen, B., Jensen, V. & Jensen, J.S. 2009.
The Danish program for domestication of native
woody species. In I.S. Kafé, ed. Abstracts from the
workshop: Genetic conservation and management
of sparsely distributed trees and bushes, Sorø,
Denmark, 15–17 September 2008. Forest and
Landscape Working Papers 36/2009. Hørsholm,
Denmark, Forest & Landscape Denmark.
Kleinschmit, J.R.G, Kownatzki, D. & Gegorius, H.R.
2004. Adaptational characteristics of autochthonous
populations – consequences for provenance delinea-
tion. Forest Ecol. Manag., 197: 213–224.
Kleinschmit, J.R.G, Leinemann, L. & Hosius, B. 2008.
Gene conservation through seed orchards – a case
study of Prunus spinosa L. In D. Lindgren, ed. Seed
orchards. Proceedings from a conference at Umeå,
Sweden, September 26–28, 2007, pp. 115–125
(available at: http://www.iufro.org/download/
file/5432/4289/20901-umea07_pdf/). Accessed 23
January 2013.
Krauss, L.S. & He, T.H. 2006. Rapid genetic identifica-
tion of local provenance seed collection zones for
ecological restoration and biodiversity conservation.
J. Nature Conserv., 14: 190–199.
Laikre, L., Schwartz, M.K., Waples, R.S. & Ryman,
N. 2010. Compromising genetic diversity in the
wild: unmonitored large-scale release of plants and
animals. Trends Ecol. Evol., 25: 520–529.
Maes, N. 1993. Genetische kwaliteit inheemse bomen
en struiken. Deelproject: Randvoorwaarden en
knelpunten bij behoud en toepassing van inheems
genenmateriaal. IBN-rapport 20. Wageningen, The
Netherlands, IKC-NBLF, IBN-DLO.
McKay, J.K., Christian, C.E., Harrison, S. & Rice, K.J.
2005. “How local is local?” – a review of practical
and conceptual issues in the genetics of restoration.
Restor. Ecol., 13: 432–440.
Vander Mijnsbrugge, K., Bischoff, A. & Smith, B.
2010. A question of origin: where and how to col-
lect seed for ecological restoration. Basic Appl. Ecol.,
11: 300–311.
Vander Mijnsbrugge, K., Cox, K. & Van Slycken, J.
2005. Conservation approaches for autochtho-
nous woody plants in Flanders. Silvae Genet., 54:
197–205.
Vander Mijnsbrugge, K., Depypere, L., Chaerle, P.,
Goetghebeur, P. & Breyne P. In press. Genetic and
morphological variability among autochthonous
Prunus spinosa populations in Flanders (northern
part of Belgium): implications for seed sourcing.
Plant Ecol. Evol.

GENET IC CONS IDERATI ONS IN ECO SYSTEM RE STORATI ON USING N ATIVE TRE E SPECIE S
47
The pedunculate oak (Quercus robur L.) in
Central Europe was intensively managed in
the past. Basically, the oak stands have been
artificially reforested, using intercropping
practices. Large numbers of acorns were
planted, and some of these were imported from
distant populations to improve the quality of
oak wood expected.
The Slavonian oak, a local provenance of pe-
dunculate oak, is reported to have a distinct
population, mostly located in the Slavonian Plain
in Croatia, southeastern Europe (Mátyás, 1972;
Klepac, 1981). This area was largely unaffected by
humans for almost 400 years from the fifteenth
century because of frequent wars and military ac-
tions. Forests started to be harvested in the late
nineteenth century. The stands were composed
of 200–300-year-old huge oaks (up to 40 metres
tall and yielding 40–50 m3 of wood) with excel-
lent wood quality. The largest tree on record had
a breast-height diameter of 260 cm and yielded
64 m3 of building timber and was sent to the
Exposition Universelle in Paris in 1900. As a result
of these qualities, this provenance was in high
demand in Europe.
Historical documents show that huge num-
bers of acorns were harvested all over Slavonia
and taken to distribution centres in the former
Austrian-Hungarian Empire. These centres distrib-
uted the acorns throughout Hungary (Kolossváry,
1975) and to other European states. Excellent
growth in many stands has been reported since
the 1880s in Austria, Czech Republic, France,
Germany, Hungary and many other parts of
Europe (Koloszár, 1982; Sabadi, 2003).
A survey of chloroplast DNA diversity in Europe
(Petit et al., 2002) has shown that the specific
haplotypes of the Balkan strain are most common
in the Slavonian oak stands, and many planted
stands elsewhere in Europe have varying propor-
tions of these haplotypes (Gailing et al., 2007)
which are indicative of Slavonian origin. The
Slavonian oak stands have not only been acclima-
Figure I4-1.
A 112-year-old Slavonian oak stand
Insight 4
Historical genetic contamination in
pedunculate oak (Quercus robur L.)
may favour adaptation
Sandor Bordacs
Central Agricultural Office, Department of Forest and Biomass Reproductive Material, Hungary

THE STATE OF THE WORLD’S FOREST GENETIC RESOURCES – THEMATIC STUDY
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48
tized to local conditions but also entered local oak
gene pools across Europe. For example, Slavonian
oak genes for early and late bud burst have been
identified provenances in Germany (Gailing et al.,
2007). This kind of genetic contamination might
be beneficial, especially in Central Europe where
most plantations have been established and
where intensive climate warming is predicted in
the next 50 years. “Imported” southern genes for
traits such as late bud burst and drought and heat
tolerance may help local oak populations adapt
to the changing climate.
References
Gailing, O., Wachter, H., Schmitt, H.-P., Curtu, A.-L.
& Finkeldey, R. 2007. Characterization of differ-
ent provenances of Slavonian pedunculate oaks
(Quercus robur L.) in Münsterland (Germany) with
chloroplast DNA markers: PCR-RFLPs and chloroplast
microsatellites. Allg. Forst Jagdztg, 178: 85–90.
Klepac, D. 1981. Les forets de chene en Slavonie. Rev.
For. Franc., 33: 87–104.
Kolossvary, Sz., ed. 1975. Az erdögazdálkodás története
Magyarországon [The history of forestry in Hungary].
Akadémiai Kiado, Budapest.
Koloszár, J. 1982. A szlavon tölgy (Quercus robur
ssp. slavonica /Gay./ Maty.) termöhelyi igénye
es erdömüvelési jelentösége. University of West
Hungary, Sopron, Hungary. (Ph.D. thesis)
Mátyás, V. 1972. A szlavon tölgy (Quercus robur
ssp. slavonica /Gay./ Maty.) erdészeti jelentösége
Magyarországon. Erd. Kut., 68: 63–77.
Petit, R.J., Csaikl, U.M., Bordács, S., Burg, K., Coart,
E., Cottrell, J., Deans, J.D., Dumolin-Lapegue,
S., Fineschi, S., Finkeldey, R., Gillies, A., Glaz, I.,
Goicoechea, P.G., Jensen, J.S., König, A.O., Lowe,
A.J., Madsen, S.F., Mátyás, G., Munro, R.C.,
Pemonge, M.H., Popescu, F., Slade, D., Tabbener,
H., Taurchini, D., van Dam, B., Ziegenhagen, B.
& Kremer, A. 2002. Chloroplast DNA variation in
European white oaks. Phylogeography and patterns
of diversity based on data from over 2600 popula-
tions. Forest Ecol. Manag., 156: 1–3, 5–26.
Sabadi, R. 2003: The position of Pedunculate Oak in
Spaˇcva, Europe, and the world. In D. Klepac & K.
Jemri’
c ˇ
Corkalo, eds. A retrospective and perspective
of managing forests of pedunculate oak in Croatia.
HAZU Centre for Scientific Research Vinkovci,
Zagreb.

GENET IC CONS IDERATI ONS IN ECO SYSTEM RE STORATI ON USING N ATIVE TRE E SPECIE S
49
Seed zones and seed movement guidelines con-
tribute to the restoration of native ecosystems
by ensuring adapted and resilient plant popula-
tions. Seed zones have a long history in the Pacific
Northwest of the United States. Plantation for-
estry was initiated in the early twentieth century
with the establishment of the Wind River Nursery
by the United States Forest Service in 1910. The
nursery was established in southwestern Wash-
ington State to reforest and restore large areas
of bare land and understocked forests result-
ing from large forest fires and logging. Initially,
foresters did not pay particular attention to the
source of forest tree seed. Seed came from eas-
ily accessible locations, often lower elevation
forests near population centres and at logging
operations. By the 1930s, however, it was be-
coming evident that not all plantations were as
productive as they could be, particularly those
at higher elevations, when compared with adja-
cent naturally regenerated stands. The gradual
decline of trees from non-local sources was also
evident in two pioneering research studies begun
in 1912: the Wind River Arboretum, which tested
trees from around the world for their suitability
to the Pacific Northwest, and the Douglas-Fir He-
redity Study, which addressed questions of type
and location of Douglas-fir (Pseudotsuga men-
ziesii (Mirb.) Franco) parents from which to collect
seed. Mortality and poor growth of trees from
off-site sources increased as stands aged, with a
particularly sharp increase in the years after an
extreme cold-weather event.
These observations led in the 1940s to the es-
tablishment of the first seed collection zones and
seed collection guidelines for Douglas-fir. A sys-
tem to certify the stand origin of forest tree seed
was established by the mid-1960s, and in 1966
seed zone maps for Washington and Oregon were
published. These maps were widely used and
have served their purpose of ensuring adapted
planting stock for reforestation and restoration.
In the meantime, researchers have learned much
more about geographic patterns of genetic vari-
ation in adaptive traits for a variety of forest tree
species, primarily from short-term genecological
studies such as those by Campbell (1986) and by
Sorensen (1992). Genecological studies consider
genetic variation as found in common garden
trials, and relate that variation to the climates or
physiography of seed sources. Consistent, sensi-
ble correlations between genetic variation and
seed-source environments indicate that a trait
has responded to natural selection and may be of
adaptive importance. Based on results from gene-
cological studies, seed zones in Washington and
Oregon were revised, primarily enlarging them in
latitudinal directions, but mostly maintaining ele-
vation limits (Randall, 1996; Randall and Berrang,
2002). This is because forest trees in temperate
Insight 5
The development of forest
tree seed zones in the Pacific
Northwest of the United States
Brad St Clair
United States Department of Agriculture Forest Service, Pacific Northwest Research Station,
United States

THE STATE OF THE WORLD’S FOREST GENETIC RESOURCES – THEMATIC STUDY
PART 2
50
and boreal regions are primarily adapted to mini-
mum winter temperatures, which are largely as-
sociated with elevation. Adaptation to aridity is
also important in some regions, and may be par-
ticularly important in the tropics. Species may also
differ in the scale and patterns of genetic varia-
tion and adaptation, and the revised seed zones
in Oregon and Washington take these differences
into account. Some species, such as Douglas-fir,
are tightly adapted to their environments and
may be considered specialist species. Other spe-
cies, like western red cedar (Thuja plicata Donn
ex D. Don), are more generally adapted and may
be considered generalist species. Consequently,
Douglas-fir has many seed zones with relatively
narrow elevation bands of 150 m, whereas west-
ern red cedar has fewer seed zones with elevation
bands of 450 to 600 m (Figure I5-1). An additional
benefit of seed zones is that they contribute to
maintaining genetic diversity and structure of for-
est trees at landscape scales that are likely most
important for adaptation.
Seed zones and seed-movement guidelines
have helped to ensure productive, healthy and di-
verse forests in the Pacific Northwest for the past
half-century and more. What have we learned?
First, even with limited or no knowledge of
genetic structure of a species, a reasonable as-
sumption is that native populations are at least
approximately adapted to their local environ-
ments (Savolainen, Pyhäjärvi and Knürr, 2007).
The question then becomes, how local is local?
Start somewhere. Make assumptions about the
climatic or other environmental variables that
are most important for adaptation, and deline-
ate seed zones based on those assumptions. In
the Pacific Northwest, initial seed zones were
based on climatic variables of cold and drought,
Figure I5-1.
Seed zones for Douglas-fir and western red cedar in Oregon and Washington State, United States
Source: adapted from Randall (1996); Randall and Berrang (2002).
for Douglas-fir western red cedar

GENET IC CONS IDERATI ONS IN ECO SYSTEM RE STORATI ON USING N ATIVE TRE E SPECIE S
51
vegetation types and physiography, especially el-
evation. Genecology studies indicated that many
of those zones were too conservative, particularly
in the north–south direction and particularly for
some species that were later determined to be
generalists. Revised seed zones took into account
this new knowledge, but in the meantime, origi-
nal seed zones based primarily on climate served
their purpose.
Second, short-term genecology studies are val-
uable for indicating genetic structure important
for adaptation and for delineating seed zones
and seed-movement guidelines. These may be
followed by longer-term reciprocal transplant
studies or provenance tests to evaluate long-
term adaptive responses, including estimat-
ing productivity given climates at the locations
of seed sources and planting sites (see Wang,
O’Neill and Aitken, 2010). An important finding
from these studies is that each species must be
considered individually, and that the patterns
and scale of adaptation are not always obvious
beforehand.
Finally, during the last few decades, scientists
and land managers have recognized that cli-
mates are changing and have begun to consider
management responses. Knowledge of genetic
variation in adaptive traits is important for un-
derstanding responses of native populations to
climate change and for evaluating management
options to adapt to climate change, including
planting populations adapted to future climates
and ensuring genetic diversity for future evolu-
tion. The primary lesson from the development of
seed zones in the Pacific Northwest is that, rather
than waiting for the genetic knowledge to accu-
mulate, it is better to act based on the best avail-
able knowledge, which may be from other species
in other regions, and then to adjust management
responses based on new knowledge from genetic
studies.
References
Campbell, R.K. 1986. Mapped genetic variation of
Douglas-fir to guide seed transfer in southwest
Oregon. Silvae Genet., 35: 85–96.
Randall, W.K. 1996. Forest tree seed zones for western
Oregon. Salem, OR, USA, Oregon Department of
Forestry.
Randall, W.K. & Berrang, P. 2002. Washington tree
seed transfer zones. Olympia, WA, USA, Washington
Department of Natural Resources.
Savolainen, O., Pyhäjärvi, T. & Knürr, T. 2007. Gene
flow and local adaptation in trees. Annu. Rev. Ecol.
Evol. Syst., 38: 595–619.
Sorensen, F.C. 1992. Genetic variation and seed transfer
guidelines for lodgepole pine in central Oregon.
Research Paper PNW-RP-468. Portland, OR, USA,
USDA Forest Service, Pacific Northwest Research
Station.
Wang, T., O’Neill, G.A. & Aitken, S.N. 2010.
Integrating environmental and genetic effects to
predict responses of trees to climate. Ecol. Appl., 20:
153–163.

GENET IC CONS IDERATI ONS IN ECO SYSTEM RE STORATI ON USING N ATIVE TRE E SPECIE S
53
The needs of a large and growing human popu-
lation have led to very high levels of habitat de-
struction. In more than half the world’s biomes
20–50 percent of land area has been converted
to human uses. The majority of conversion is for
agriculture, which is expanding in 70 percent of
countries, declining in 25 percent, and static in
5 percent (FAO, 2003). Although biogeographic
regions differ markedly in the extent of habitat
conversion to agriculture (Klein Goldewijk, 2001),
in all regions at least 25 percent of the area had
been converted to other land-uses by 1950. Tropi-
cal dry forests are the biome most affected, with
almost half replaced by agriculture (Mace et al.,
2005). Nearly 25 percent of tropical rain forest has
been fragmented or entirely cleared (Mace et al.,
2005), while temperate broadleaf and Mediterra-
nean forests have experienced 35 percent or more
conversion. However, global assessments show a
decline in the rate of forest loss from 1990 to 2005
(Chazdon, 2008).
A major issue in land-use change is habitat frag-
mentation, defined as the reduction in area of a
specific habitat type and division of the remain-
ing habitat into smaller and spatially separated
habitat patches as a result of replacement by an-
thropogenic land-uses, such as agriculture, human
settlements or plantation forestry. The degree
of fragmentation also varies between regions,
with forest biomes in Africa and Europe twice as
likely to be classified as “fragmented forest” com-
pared with North and South America (Wade et al.,
2003). Habitat fragmentation generally results in
a complex landscape mosaic of native and human-
dominated habitat types, which may have serious
consequences for many species. This paper exam-
ines the impacts of fragmentation on the genetic
viability of tree populations and how habitat con-
nectivity and landscape functionality relate to the
conservation and use of native tree species.
4.1. Genetic problems related
to fragmentation
Maintenance of genetic diversity in trees is vital
for the continued fitness, resilience, adaptation
and evolution of their populations (Ellstrand,
1992; Garner, Rachlow and Waits, 2005). Habitat
fragmentation can lead to loss of allelic diversity
through increased inbreeding and reduced ef-
fective population size as a result of the genetic
isolation of populations. Specifically, inbreeding
may result from increased self-pollination, or
where remaining trees are related through recent
common ancestry (biparental inbreeding; Young,
Boyle and Brown, 1996). Genetic isolation and in-
breeding can lead to reduced fitness or inbreed-
ing depression through: (1) lack of effective ferti-
lization; (2) expression of deleterious alleles (Sork
et al., 2002); and (3) general reduction in hete-
rozygosity (Ellstrand, 1992). Inbreeding may have
Chapter 4
Fragmentation, landscape
functionalities and connectivity
Tonya Lander1 and David Boshier2,3
1 Natural History Museum, London, United Kingdom
2 Department of Plant Sciences, University of Oxford, United Kingdom
3 Bioversity International, Italy

THE STATE OF THE WORLD’S FOREST GENETIC RESOURCES – THEMATIC STUDY
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54
especially dire consequences for species that were
mainly outcrossing, such as many tree species (Ell-
strand, 1992; Husband and Schemske, 1995; but
see Williams and Savolainen, 1996; Young, Boyle
and Brown, 1996).
Importantly, ecological factors sometimes
pose a more imminent conservation threat than
genetic degradation (Caughley, 1994). For ex-
ample, global declines in pollinators, associated
with land-use change (Ricketts, 2001; Steffan-
Dewenter, Munzenberg and Tscharntke, 2001;
Baum et al., 2004; Kremen et al., 2007) and frag-
mentation (Frankham, 1995; Eckert et al., 2010),
may disrupt mutualistic relationships and consti-
tute a problem not only for species survival, but
also for continued ecosystem function and crop
production (Biesmeijer et al., 2006). If ecologi-
cal or demographic risks are the most pressing,
an undue focus on the genetic consequences of
fragmentation can represent a missed opportuni-
ty to address key ecological risks and environmen-
tal factors of immediate concern (Asquith, 2001).
Generalizations about the potential genetic ef-
fects of fragmentation must be evaluated in the
light of evolutionary history, life history and mat-
ing systems to provide a more complete, albeit
complex, understanding (Kramer et al., 2008). In
this context impacts on pollinators are as impor-
tant to understanding genetic impacts of fragmen-
tation on trees. Studies of fauna and flora suggest
that some species appear more vulnerable to frag-
mentation than others. For example, species with
large range area requirements, primary habitat
specialists (Tilman et al., 1994; Laurance et al.,
2001) and those with low population growth rates
or poor dispersal ability may be especially vulnera-
ble. Similarly, species with low population density,
such as tropical trees, may be more vulnerable be-
cause populations are already small in number and
spatially diffuse, although these species may also
have adaptations that allow persistence at low
density, such as pollination mechanisms adapted to
obligate long-distance pollination (Kramer et al.,
2008). Some species of both fauna and flora are
also particularly vulnerable to edge effects, where
land at the edge of the habitat patch is altered and
the environment becomes more extreme and less
amenable (Woodroffe and Ginsberg, 1998).
Pollen flow is the primary mode of gene flow
in plants (Ellstrand, 1992; Young, Boshier and
Boyle, 2000; Slavov, DiFazio and Strauss, 2002;
Bittencourt and Sebbenn, 2008) and knowing how
this changes as a result of fragmentation is vital
to understanding fragmentation impacts on trees.
Generally in plant species, levels of pollen flow be-
tween fragments appear to be affected by inter-
specific differences in longevity, generation time
and pre-fragmentation abundance, the range of
sexual and asexual reproductive systems (Young,
Boyle and Brown, 1996; Cascante et al., 2002; Kolb
and Diekmann, 2005), habitat specificity (Davies,
Margules and Lawrence, 2000), plant height (Kolb
and Diekmann, 2005), pollination and seed disper-
sal syndromes. Studies also show that the impacts
of fragmentation on pollen flow are more varied,
complex and subtle than original theoretical pre-
dictions. Given this complexity, it is unsurprising
that many studies have found small or no clear ge-
netic effects. Kramer et al. (2008) suggest that four
key assumptions in fragmentation studies must be
re-evaluated: (1) fragment edges delimit popula-
tions; (2) genetic declines manifest quickly enough
to be detected; (3) species respond similarly to
fragmentation; and (4) genetic declines supersede
ecological consequences. For tree species, for ex-
ample, the assumption that pollen dispersal stops
at fragment edges is contradicted by evidence
that in many cases pollination between fragments
is not at all rare (e.g. Young and Merriam, 1994;
Nason and Hamrick, 1997; Dow and Ashley, 1998;
Streiff et al., 1999; Apsit, Hamrick and Nason, 2001;
White, Boshier and Powell, 2002; Latouche-Halle
et al., 2004; Nakanishi et al., 2004; Lander, Boshier
and Harris, 2010). Thus, quite ordinary pollen dis-
persal may be sufficient to link trees in scattered
forest fragments into a functioning metapopula-
tion. In this case the potential negative genetic
effects of small population size would not be real-
ized. This positive view must be balanced by evi-
dence of altered patterns of pollen flow, whereby
connectivity is maintained but biparental inbreed-
ing increases or reduced pollen pool diversity is

GENET IC CONS IDERATI ONS IN ECO SYSTEM RE STORATI ON USING N ATIVE TRE E SPECIE S
55
sampled in mating. Genetic signals may also re-
quire several generations to appear, which could
amount to hundreds of years in the case of long-
lived tree species (Kramer et al., 2008). Moreover,
even if forest fragments are not reproductively iso-
lated or suffering immediate losses of genetic di-
versity, there may be quantitative pollen limitation
(O’Connell, Mosseler and Rajora, 2006) or seed
dispersal limitation, which could limit recruitment
(Kramer et al., 2008).
4.2. Management of fragmented
landscapes
Protected areas
Worldwide, countries have designated protected
areas to conserve predominantly terrestrial native
ecosystems and biodiversity features (Mace et al.,
2005). Biomes differ widely in the percentage of
total area under protection. Of the lands classi-
fied in the four highest IUCN protection catego-
ries, flooded grasslands, tundra, temperate conif-
erous forests, mangroves and boreal forests have
the highest percentage area under protection.
This may be because these biomes are among the
least useful for competing land-uses such as ag-
riculture. Temperate grasslands, Mediterranean
forests and tropical coniferous forests are the
least protected biomes. Many protected areas
exist within landscapes that are fragmented to
a greater or lesser degree, and may not be large
enough to be viable in the long-term. As such,
despite their protected status they are subject to
the same biological issues that face any remnant
fragment of a native ecosystem.
A continuing debate relating to reserve design
is whether it is biologically more effective to set
aside a single large reserve area or several small
ones (Diamond, 1975). Generally, reserve de-
sign has been based on theoretical estimates of
extinction risk and colonization rates and, most
practically, land availability. Population dynamics
models suggest that the reserve design that mini-
mizes extinction risk is species- and case-specific,
depending on dispersal ability, environmental
factors, and extinction and colonization patterns
(McCarthy et al., 2011). However, most models
do not take into account the influence of uncer-
tainty in extinction risk on optimal reserve design.
Mathematically, rather than minimizing the ex-
pected extinction risk, a better objective may be
to maximize the chance that extinction risk is ac-
ceptably small (McCarthy et al., 2011). In practice,
the creation of reserves remains limited by land
availability, resources to manage reserves, local
support and participation, continuity of political
will and capacity to protect reserve areas, and
competition for other land-uses, such as food pro-
duction. This has usually resulted in a bias of pro-
tected areas to upland areas and an absence on
lowland fertile soils. Coupled with deforestation
and fragmentation, often superimposed on habi-
tat heterogeneity, the result is a disproportionate
loss of ecosystems, species, populations and geno-
types adapted to lowlands and fertile soils.
The shortcomings of conservation methods that
rely on exclusion of people from reserves are in-
creasingly recognized. Problems stem from land-
tenure conflicts, displacement of local people and/
or their activities and development needs, the
costs of reserve management and protection, and
opportunity costs for countries where reserves are
located (Wells and Brandon, 1992; Brockington
and Schmidt-Soltau, 2004). To mitigate these prob-
lems some initiatives have experimented with al-
lowing human activities inside reserves or inside
buffer areas around reserves. Other initiatives
have taken the view that if local people benefit
from the reserve they will be motivated to protect
it and so have actively encouraged local people
to use reserves; this has been called conservation
through use (CTU). However, after 15 years there
is limited evidence that CTU initiatives achieve
species and ecosystem conservation at the same
time as improving local livelihoods (Belcher and
Schreckenberg, 2007; Barrance, Schreckenberg
and Gordon, 2009). A multidisciplinary approach is
needed to investigate the potential for integrat-
ing conservation and development and, more spe-
cifically, which species are or could be sustainably

THE STATE OF THE WORLD’S FOREST GENETIC RESOURCES – THEMATIC STUDY
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56
conserved in such systems, from both biological
and human management perspectives. Efforts to
maintain genetic diversity and adaptive capacity
within species are irrelevant if current manage-
ment drastically reduces the possibility of popula-
tion persistence.
Remnant trees
As discussed above, habitat fragmentation and
physical isolation do not always impede pollen
flow and may increase it (but see Cascante et al.,
2002; Hamrick, 2004; Sork and Smouse, 2006; By-
rne et al., 2007). Despite large-scale studies, we
still do not know at what distance forest frag-
ments become genetically isolated, although new
research is showing that the question is perhaps
increasingly irrelevant. Rather than complete iso-
lation, evidence currently points to alterations of
mating patterns with increased distance. In some
cases, single or “isolated” trees receive pollen
from a wide spatial and genetic array of pollen
donors, and more individual pollen donors than
trees in groups (Hamrick, 2004). Single trees may
also be less likely to receive pollen from their near-
est neighbours than trees in groups (Chase et al.,
1996; White, Boshier and Powell, 2002; Dick, Etch-
elecu and Austerlitz, 2003), although in other spe-
cies the reverse appears to be true (Ward et al.,
2005). The extent to which such single trees exhibit
selfing appears to depend on the presence and
strength of any self-incompatibility system within
the species. Other paternity studies in fragmented
landscapes have shown that, while the majority of
pollen dispersal events are of the order of tens or
hundreds of metres in both wind-pollinated (e.g.
Dow and Ashley, 1996; Sork et al., 2002) and insect-
pollinated species (e.g. Kwak, Velterop and van
Andel, 1998; Konuma et al., 2000; Lander, Boshier
and Harris, 2010), pollen can travel tens or hun-
dreds of kilometres (e.g. 6–14 km in Ficus spp., Na-
son and Hamrick, 1997; Nason, Herre and Hamrick,
1998; 3.2 km in Dinizia excelsa, Dick, Etchelecu and
Austerlitz, 2003; see also Kwak, Velterop and van
Andel, 1998; Chuine, Belmonte and Mignot, 2000;
Hamrick and Nason, 2000; Burczyk, Lewandowski
and Chalupka, 2004). In addition, pollen disper-
sal distances have been shown to exceed pollina-
tor flight distances as a result of pollen carryover
(Ellstrand, 1992; Ghazoul, Liston and Boyle, 1998).
Such data suggest that remnant trees and small
patches of trees can be effective and important
in maintaining genetic connectivity across frag-
mented landscapes and in conserving genetic di-
versity (Lander, Boshier and Harris, 2010).
Corridors
As a result of the dominance of island biogeogra-
phy-based ideas that (1) habitat and non-habitat
are clearly distinguishable and (2) non-habitat
is wholly hostile for organism travel (MacArthur
and Wilson, 1967; Ricketts, 2001; Vandermeer and
Carvajal, 2001; Jules and Shahani, 2003), manage-
ment of fragmented landscapes has frequently
focused on the impacts of spatial isolation of in-
dividuals or species (Sork and Waits, 2010). The
land between habitat patches has been consid-
ered ecologically uniform and generally hostile
(Ricketts, 2001; Vandermeer and Carvajal, 2001),
with the probability of organism survival and dis-
persal treated as a function of habitat fragment
size and linear distance between fragments (isola-
tion by distance; Jules and Shahani, 2003). Given
this focus, research is frequently framed in terms
of how probable it is that an organism will be
able to pass through a certain area to move be-
tween habitat patches.
Programmes to mitigate the potential nega-
tive effects of fragmentation have tended to
focus on increasing landscape “connectivity,”
defined as the degree to which the landscape fa-
cilitates or impedes movement of organisms be-
tween habitat patches (Adriaensen et al., 2003).
Increasing connectivity between habitat patches
is expected to increase effective population size,
reduce inbreeding and facilitate migration, dis-
persal and colonization (Li et al., 2010; Hagerty
et al., 2011). In a landscape classified in a binary
way into habitat and non-habitat, the logical ap-
proach to increasing connectivity between habi-
tat patches is to build bridges, called corridors or
stepping stones (Villalba et al., 1998; Adriaensen
et al., 2003). Corridors are narrow strips of habi-

GENET IC CONS IDERATI ONS IN ECO SYSTEM RE STORATI ON USING N ATIVE TRE E SPECIE S
57
tat built or conserved to connect habitat patches,
while stepping stones are small patches of habitat
scattered through the non-habitat area between
larger patches of habitat (Levin, 1995; Nason and
Hamrick, 1997; Lowe et al., 2005). The hypothesis
is that the more similar the corridor area is to na-
tive habitat, the more likely it is that an organism
will move through it. Although much theoretical
and empirical attention has been given to biolog-
ical corridors and stepping stones, whether they
are effective or not is unresolved (e.g. Simberloff
et al., 1992; Pullinger and Johnson, 2010; Richard
and Armstrong, 2010) and may often be related
to a lack of clarity over identification of the target
species and their specific connectivity needs.
Conservation outside protected areas:
an integrated landscape approach
There is a growing perception that non-habitat ar-
eas outside reserves and outside remnant patches
of native habitat may provide non-ideal rather
than fatal environments (Gustafson and Gardner,
1996; Moilanen and Hanski, 1998; Arnaud, 2001;
Vandermeer and Carvajal, 2001; Bender, Tischen-
dorf and Fahrig, 2003; Coulon et al., 2004; Lander,
Boshier and Harris, 2010). Numerous empirical
studies have shown that the type of non-habitat
between habitat patches affects patterns of in-
sect and other animal dispersal and hence seed
and pollen dispersal (Ghazoul, Liston and Boyle,
1998; Franklin et al., 2000; Davies, Melbourne and
Margules, 2001; Ricketts, 2001; but see Bruna and
Kress, 2002; Baum et al., 2004; Darvill et al., 2006).
Thus the focus of conservation management can
change from (1) how much land can be set aside,
(2) how to minimize the linear distance between
habitat patches or (3) how to create habitat
bridges, and turn instead to measures of separa-
tion between habitat patches that incorporate
variation in how easily the target organism passes
through the different land-use types in the matrix
(i.e. permeability; Spear et al., 2010; Doerr, Bar-
rett and Doerr, 2011; Hagerty et al., 2011; Lander
et al., 2011).
Although the permeability concept recognizes
the potential for variation in non-habitat resist-
ance to pollinator movement, it is still based on
a binary landscape model where the question
is about the study species’ presence in, absence
from or travel between native habitat patches.
Some recent models of organism movement in
fragmented landscapes are not concerned with
the designation of different parts of a landscape
as habitat or non-habitat, but rather focus on the
quantity and accessibility of resources and threats
in that landscape. Thus, land outside traditionally
defined native habitat ceases to be an area to
pass through and is investigated in its own right
for its capacity to provide habitat services as well
as its ability to support or inhibit movement (e.g.
Lander et al., 2011). The entire landscape in this
case may be considered a patchwork of partial
habitats of varying quality (Kremen et al., 2007).
A growing body of research suggests that this
“partial habitat” or “resource model” view may
be both accurate and useful. For example, urban
gardens, rough grassland and clover leys, none of
which are native habitat, provide vital habitat ser-
vices for bumblebees in agro-ecosystems (Goulson
et al., 2010). Similarly, models that incorporate
resource availability in various land-use types,
such as floral resources for bees in both agricul-
tural fields and native vegetation, have high ex-
planatory value in predicting bee abundance and
species richness (Winfree, 2010). Thus, landscape
models that recognize the potential habitat ser-
vices that different, apparently non-habitat, land-
uses may provide could be a useful basis for inter-
preting empirical data and developing landscape
management strategies.
4.3. The use of native species
in ensuring functionality
in fragmented landscapes
Against this background, land managers are asked
to select, design, manage and link landscapes that
will be effective in conserving biodiversity. Refor-
estation of degraded tropical forest lands has
frequently involved the establishment of single-
species plantations of fast-growing exotic species

THE STATE OF THE WORLD’S FOREST GENETIC RESOURCES – THEMATIC STUDY
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58
(e.g. Pinus, Eucalyptus or Acacia) which generate
mainly financial benefits, whereas ecological res-
toration that maximizes biodiversity may produce
few, short-term economic benefits (Lamb, Erskine
and Parrotta, 2005). If, as is often the case, res-
toration must be balanced with financial returns,
plantations of native species, in monoculture or
as species mixtures, may provide more biological
value than plantations of exotic species (Chazdon,
2008) while still providing better financial returns
than “pure” ecological restoration projects. The
economic and biological value of plantations of
native species may be increased by underplanting
the trees with shade-tolerant agricultural cash
crops or species that produce non-timber forest
products. Compared with monocultures, mixed
plantations can deliver higher production, protec-
tion against disease and pest damage (ecological
resilience) and greater security in uncertain fu-
ture markets (financial resilience).
In highly modified landscapes where specific
corridor development is called for, studies support
a broad vision of corridor design where a range of
land-uses may be combined to provide a perme-
able and ecologically functional landscape rather
than the traditional approach of building con-
tinuous habitat corridors connecting intact forest.
Corridor design, management and monitoring
should involve assessment of different land-use
types in terms of their ability, individually and
in combination, to support movement of target
species and provide the resources target species
need. Assessments of corridor function will be
linked to the characteristics of the target species,
sustainable land-use aims of the area and vari-
ability between species and ecosystems in their
resilience to disturbance. The balance of land-use
types may need modification to maintain or im-
prove connectivity.
This approach to landscape management is
based on identifying land-uses that provide both
habitat services and social and economic returns;
the balance between these needs is clearly con-
text dependent. For example, in an area of high
forest cover, land-uses may be assessed prin-
cipally for gene flow, whereas in much more
highly deforested areas a fuller complement of
benefits may be sought from particular systems,
with their specific location in the corridor zone
also being important. Thus, in the highly defor-
ested dry forest zone of western Honduras the
traditional Quezungual fallow system (Kass et
al., 1993), in which farmers manage naturally
regenerated shrubs, fruit trees and timber trees
among their crops, is likely to provide a variety
of genetic conservation benefits for a range of
native tree species without the need for estab-
lishing specific biological corridors. Other com-
plex systems, such as traditional shaded coffee
or jungle rubber, may also rate highly for genetic
conservation benefits. In contrast, simpler agro-
forestry systems such as pasture trees and living
fences offer fewer genetic conservation benefits
and are unlikely to prove effective mediators
of pollen flow for species without a self-incom-
patibility mechanism. The emphasis generally
should be on maintenance and improvement of
economically and socially viable landscapes that
promote connectivity (for genes, species and
ecological processes) and conservation of biodi-
versity more generally.
Importantly, assessments of the genetic con-
servation benefits of agro-ecosystems are more
likely to be species specific than management-
system specific, and need to take into account the
farming system, density of trees and their origin
(natural regeneration or planted). For example,
maintaining native timber trees over large areas
of coffee is likely to have beneficial genetic ef-
fects for gene flow, population numbers and con-
servation of particular populations. However, if
the same system were used in only a small area
it could lead to a reduced genetic base in seed
production through related (biparental) mating.
Thus, the area or management unit should be
measured in numbers of participating households
or numbers of land units in which land-uses ben-
eficial to target species conservation are practised
(Boshier, Gordon and Barrance, 2004). Given the
speed with which land-use may change in re-
sponse to market prices, this measure in itself may
require monitoring.

GENET IC CONS IDERATI ONS IN ECO SYSTEM RE STORATI ON USING N ATIVE TRE E SPECIE S
59
Identifying the factors that leave some species
genetically susceptible to human disturbance re-
quires extensive reproductive and regeneration
ecology and genetic data. The lack of informa-
tion, resource limitations and the need for more
immediate action in many situations necessi-
tates pragmatic best-guess approaches to iden-
tify which land-uses may favour gene flow for
which species and which will not. The ability to
extrapolate from results from model species to
make more general recommendations for spe-
cies management groups (combining ecological
guild, spatial distribution and reproductive biol-
ogy) depends on the existence of basic biological
information (e.g. incompatibility and pollination
mechanisms, dispersal and seedling regeneration)
that enables species to be classified (Jennings et
al., 2001).
Consideration of available information sug-
gests that the following species types are un-
likely to show genetic conservation benefits from
tree-based agro-ecosystems: outcrossing species
that are self-compatible; slow-growing species
that reproduce only when they are large (or, in
the extreme, monocarpic species, i.e. those that
flower only once in their life); species with poor
regeneration under human disturbance; species
with highly specific pollinators or seed dispers-
ers susceptible to disturbance; rare species with
low population densities; and species with highly
clumped distributions. Inevitably, such generali-
zations will be qualified by the range of factors
that have been shown to influence patterns of
genetic variation in trees.
4.4. Conclusions: policy and
practice
• We need clear objectives in conservation
planning that clearly identify target species,
ecosystems and biogeographical regions.
• We need to continue to improve our
understanding of the ecology of target
species and ecosystems so that it is possible
to make decisions about management that
will: (i) create landscapes where habitat and
ecosystem services are provided alongside
economic outputs; (ii) increase landscape-
scale genetic connectivity between remnant
populations; and (iii) maintain ecosystem,
species and genetic diversity.
• Evidence suggests that for many tree
species, populations and individuals, gene
flow may be high across some fragmented
landscapes with little apparent forest
cover. The view of forest fragmentation as
producing genetic isolation may be more a
human perception than a true reflection of
actual gene flow. It is therefore important
to recognize the complementary role that
maintenance of trees on farms is already
playing to in situ conservation. Trees in
a whole range of agroforestry and other
land-use systems may play an important but
varied role in the long-term genetic viability
of many native tree species, facilitating
gene flow between existing reserves,
conserving particular genotypes not found
in reserves, maintaining minimum viable
populations and acting as intermediaries
and alternative host habitat for pollinators
and seed dispersers (Harvey and Haber,
1999). Underestimating the capacity of
many species to persist in large numbers
in these agro-ecosystems under current
practices could lead to the misdirection
of limited conservation resources toward
species not under threat (Boshier, Gordon
and Barrance, 2004). Agroforestry tree
populations may represent a considerable
conservation resource, which, if taken into
consideration, may show that species that
are currently assumed to be threatened by
habitat loss are thriving (Vandermeer and
Perfecto, 1997).
However, although they undoubtedly con-
tribute to reproduction in remnant forests, the
benefits and effects are more complex than pre-
dicted and vary from species to species. Uneven
representation and overrepresentation in pollen
pools and mating may lead to non-random mat-

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60
ing, with reductions in genetic diversity in subse-
quent generations. We should not overestimate
the extent to which agro-ecosystems will benefit
the genetic conservation of forest tree species.
In addition to some of the complications raised
here, it is evident that many of the tree species
found in agro-ecosystems are already present in
adequate numbers in existing reserves. Similarly,
some of the species threatened by low population
numbers are not of the type that will easily per-
sist in such systems. The greatest potential role of
agroforestry and other agro-ecosystems will be in
highly deforested areas where reserves are very
small or nonexistent and where the trees main-
tained in these systems represent an important
part of a particular population’s or species’ gene
pool. In such circumstances, the fact that many
tree species that live in such disturbed vegetation
can be conserved through existing practices can
free resources for the conservation of more criti-
cally threatened species needing more conven-
tional, resource intensive approaches.
• We need to ascertain which land-uses
are favourable to connectivity and
conservation and which are antagonistic.
In the fragmented forests of central Chile,
Lander et al. (2011) found that support
for subsistence farms and modification of
management to reduce the size of pine
plantation clearfells would be likely to
have significant positive impacts on the
viability of pollinator populations and the
probability of pollinators moving across the
landscape between native forest fragments.
Goulson et al. (2010) and Winfree (2010)
found that urban gardens, rough grassland
and floral resources in agricultural
fields provided vital habitat services for
pollinators in landscapes dominated by
agriculture. Agricultural land itself can
provide habitat services, depending on
the diversity of crops, size of individual
fields, use of agrochemicals, application
of integrated crop- and pest-management
systems and management of waterways
and soils (Sustainable Agriculture Network,
2010). Although the process of identifying
favourable land-uses has begun, this is a rich
area of study.
• We need a broader view of conservation
that recognizes that reserves can be only
part of the solution to our conservation
concerns and embrace the possibility that
anthropogenic land-uses may provide
valuable and necessary ecosystem and
habitat services. In the binary view, where
only native habitats, or those land-uses most
similar to native habitats, are recognized as
providing ecosystem services, land managers
will tend to focus on distances between
habitat patches, degree of habitat patch
aggregation and corridor-type connectivity
(Doerr, Barrett and Doerr, 2011). This lack
of differentiation between non-habitat
land-use types limits management options
and contributes to polarization of the
conservation debate, leaving decision
makers with the unenviable task of choosing
between economic activity or setting aside
land for conservation. If we move beyond
the expectation that organisms will move
in a directed manner between areas that
have been designated as habitat and focus
instead on the ecological requirements of
target species or the ecological attributes
of the land-uses in the wider landscape,
we may understand how best to manage a
mosaic of habitats of varying quality. This
type of landscape management strategy
could be both more effective biologically
and less expensive than traditional
conservation based on land set aside for
conservation.
• The complementary benefits of different
land-use practices for genetic conservation
must be further evaluated, recognized
and promoted. There is a need to
raise awareness among development
professionals of the value of natural
regeneration as both a conservation and
socioeconomic resource. The emphasis on
a limited range of species, often exotics,

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61
by development agencies may reduce the
potential genetic benefits of such systems,
besides creating potential problems of
invasiveness. However, there is also a need
for conservation planners, more accustomed
to in situ methods, to consider the possibility
that tree populations found outside
protected areas have a role in biodiversity
conservation (Boshier, Gordon and Barrance,
2004). This in turn requires the direct
involvement of development organizations
in biodiversity conservation and an effective
interaction between them and traditional
conservation organizations to ensure both
conservation and development benefits.
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In forest trees, as in all organisms, new genetic
variants are generated by mutation. The adap-
tive value of the new variants is initially tested by
strong selection pressure during the production
of sporophytes and gametophytes in the local
environmental conditions in which the mutations
appeared. However, the potential benefits of ge-
netic mutations can be tested under different en-
vironmental conditions through gene flow.
Of the five main evolutionary forces (mutation,
recombination, selection, genetic drift and gene
flow), gene flow is the only one that can generate
new genetic variation through the direct or indi-
rect combination of genetic variants and occurs
at a landscape scale. Gene flow in sessile, long-
lived organisms like trees depends strongly on the
movement of gametes in the form of pollen grains
(pollen flow) and zygotes, usually as seeds (seed
dispersal). Several tree species can also propagate
vegetatively through broken twigs, root suckers
or layers, distributing genetic information identi-
cal to that of the original tree. In most species pol-
len, seeds and vegetative propagants are moved
by vectors, such as wind, animals, water or human
beings.
Gene flow is defined as “the proportion of
newly immigrant genes moving into a popula-
tion” (Endler, 1977). Movement of genes within
populations is termed gene movement (Devlin,
Roeder and Ellstrand, 1988). However, anthro-
pogenic impacts have modified the movement
of genetic variants not only among populations
but also, frequently, within them. Moreover, the
physical dimension of a “biological population”
(individuals that exchange genetic informa-
tion and share a common evolutionary path) is
difficult to define in nature. As a result, for practi-
cal purposes, such as for restoration, gene flow is
considered to include not only exchange of genes
among populations but also the local reproduc-
tive system dynamics within them.
We differentiate between “gene flow” and
“effective gene flow” (see Lowe, Harris and
Ashton, 2004). Effective gene flow at the pollen
level is influenced by movement of the pollen
grain from the male flower to the female flower,
germination in the pistil or equivalent structure
and fertilization of the ovule to form a new zy-
gote (seed). Pollen can move between flowers of
the same tree, between flowers of different trees
within the same population and/or between flow-
ers of different trees from different populations.
Effective gene flow at the seed level is influenced
by the movement of the seed, its germination and
establishment as a sapling. Seed can be dispersed
to different places within the same population, to
different populations and/or to newly available
habitats (i.e. colonization).
Pollen flow is much more extensive than seed
dispersal (see, for example, Petit et al., 2005).
However, it has been shown that seed dispersal
can be more effective than pollen dispersal at
maintaining genetic connectivity in exceptional
circumstances, as in the case of Fraxinus excelsior
fragments in an ancient deforested landscape
(Bacles, Lowe and Ennos, 2006).
Chapter 5
Gene flow in the restoration
of forest ecosystems
Leonardo Gallo and Paula Marchelli
Unit of Ecological Genetics and Forest Tree Breeding, INTA Bariloche, Argentina

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5.1. Genetic effects at different
scales
Gene flow connects populations and influences
colonization, and therefore constitutes one of the
main processes to be considered when managing
and/or restoring forest ecosystems. Human
activity continuously modifies the environment
at various overlapping spatial and temporal
scales. Matching the two dynamic processes of
effective gene flow and environmental change is
essential to promote viable and adaptable forest
ecosystems.
Gene flow integrates many different ele-
ments and functions of the ecosystem through
various selection gradients. Any restoration ac-
tivity should take into consideration the complex
equilibrium that rules gene flow in the desired
species. Assessment of this equilibrium should be
made at the appropriate spatial and temporal
scales, given that effective gene flow in tree spe-
cies occurs over ranges from tens of metres to sev-
eral kilometres and generation intervals are long.
The general increase in genetic diversity within
populations due to gene flow can be considered
as an advantage for the adaptation and adapt-
ability of the forest ecosystems. This is especially
true when the immigrant genes and/or genotypes
are better adapted to the local environmental
conditions than the local genotypes. However,
gene flow can introduce undesirable genes and/
or genotypes if the local population is already
well adapted and in evolutionary equilibrium
with the environmental conditions. If the influx
of maladapted genes is large relative to the size
of the extant population, the local genetic ad-
aptation can be undermined, although natural
selection would be expected to weed out poorly
adapted individuals over time. However, the long-
distance gene flow commonly found in forest tree
species may augment the ability of populations
to respond to climate change through a general
increase in genetic variation in the population
(Kremer et al., 2012).
5.2. Considerations in restoration
and management
Forest restoration and management activities
may have an active or passive impact on the gene
flow occurring in the system. Active gene flow
management (AGFM) relates to the movement
of genetic material from one location to another
by human beings. Passive gene flow management
(PGFM) relates to the modification of the land-
scape and environment to facilitate the natural
movement and recruitment of genetic variants
into the population to be restored. It could in-
clude, for example, the reintroduction of native
animals functioning as pollen and/or seed vec-
tors. Both types of interventions can and should
be used at the same time in some restoration situ-
ations. In PGFM, human activities modify not only
the landscape in which natural gene flow takes
place and/or new genotypic variants establish but
also the local environmental conditions. In some
cases the improvement of the receptor environ-
ment could facilitate effective natural dispersion
by increasing the likelihood of establishment of
the desired genetic variants.
In some special cases, AGFM can be implemented
even across continents. In the last century, south-
ern beech species (Nothofagus spp.) originating
in Chile and Argentina were used to restore some
English landscapes because they grow better than
native Fagaceae species (L. Gallo, personal obser-
vation), while maintaining landscape functional-
ity and visual appearance (Poole, 1987). A female
clone of Salix × rubens, a hybrid between S. fragilis
and S. alba, was introduced into the Patagonian
region of South America from Eurasia and over the
last 100 years has colonized huge areas, in some
cases more than 100000km2, mainly through nat-
ural distribution of broken twigs by water (Budde
et al., 2011). It has also hybridized and intro-
gressed with the native species, S. humboldtiana,
competing for its natural habitat and diluting its
gene pool (Bozzi et al., 2011). Another remarkable
case of human influence on the distribution pat-
tern of genetic diversity in forest tree species is the
vegetative propagation and distribution of an elm

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clone of Ulmus minor var. vulgaris (Ulmus procera)
by the Romans, who used it to support grape vines
in their vineyards in France, Spain and England (Gil
et al., 2004). Araucaria araucana was probably in-
troduced in some areas of northern Patagonia by
the Mapuche people, who used its seeds as food
(Gallo, Letourneau and Vinceti, 2004; Marchelli et
al., 2010).
At the landscape scale, the effect of human
activities on gene flow is a consequence of for-
est ecosystem modifications mainly through frag-
mentation and introduction of artificial barriers.
Both of these mainly affect vectors.
The genetic consequences of habitat fragmen-
tation depend on whether it affects gene flow; if
it restricts gene flow, fragmentation is highly del-
eterious in the long term (Frankham et al., 2002).
When fragmentation occurs, movement of pollen
and seed is determined by the distances between
the fragments (particularly in wind-pollinated
species) or the environmental conditions of the
landscape between the fragments (particularly
in insect-pollinated species). The effects of forest
fragmentation on the behaviour of pollinators
and animal vectors that disperse seed differ de-
pending on species and cannot be generalized. In
some cases fragmentation can reduce gene flow
distances (Powell and Powell, 1987) and in others
has been shown to increase gene flow (White,
Boshier and Powell, 2002).
In some regions of the world gene flow is af-
fected by unidirectional movement of the vec-
tor during pollination or seed dispersal. In such
cases the effect of fragmentation depends essen-
tially on the location of the fragments. This is the
case for fragmented populations of Patagonian
Cypress, Austrocedrus chilensis, growing in a
xeric region where less than 2 percent of the
wind blows a north–south (or south–north) dur-
ing pollination and more than 75 percent blows
west–east (Figure 5.1). This dioecious species has
been shown to have pollination distances of over
Figure 5.1.
Fragment of Austrocedrus chilensis (Patagonian cypress) with about 100 hundred individuals,
separated from a neighbouring fragment by just 1200 m and occurring on an arid grassland steppe
matrix with 350 mm of mean annual precipitation. The orientation of the fragments is north–south
and therefore gene flow is severely restricted since wind persistently blows in west–east.

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1000m in the fragmented margins of it s natural
distribution (Colabella, 2011). However, fragments
lying less than 1200 m apart on a north–south axis
were found to be genetically isolated from each
other using isozymes (Gallo and Pastorino, 2010)
and microsatellite markers (Arana et al., 2010). In
addition to the reproductive isolation, the small
effective population size of these fragments re-
sulted in genetic drift that is expressed in the fixa-
tion and loss not only of some neutral alleles but
also of some adaptive and continuously varying
traits. In this case the vector, namely the wind,
acts as a dynamic barrier because of its persistent
directionality (“reproductive isolation by wind,”
Gallo and Pastorino, 2010).
Poverty and lack of ecosystem protection and
management controls affect the pollen and seed
dispersal of many forest tree species. For exam-
ple, hunting of animals that are seed or pollen
vectors reduces their population size and there-
fore reduces gene flow. In contrast, introduc-
tion of novel vectors can offset other impedi-
ments to gene flow. For example, introduction
of African bees, which can fly long distances in
fragmented landscapes, resulted in greater pollen
flow between Dinizia excelsa (Fabaceae) trees in
fragmented Amazonian rainforest than that re-
corded in pristine forest without the African bees
(Dick, Etchelecu and Austerlitz, 2003). In some
wind-pollinated species, fragmentation increases
the speed and free movement of wind and con-
sequently the dispersal distances of pollen and
seeds (Young et al., 1993; Bacles, Lowe and Ennos,
2006).
Gene flow can also be restricted or completely
interrupted by artificial physical barriers (plan-
tations of introduced species, buildings, wind-
breaks etc.). However, in Australia significant
gene flow has been reported between remnant
natural populations of Eucalyptus loxophleba
and introduced plantations of the same species
(Sampson and Byrne, 2008). In the Canary Islands,
the natural regeneration in artificial plantations
of Pinus canariensis were found to have greater
genetic diversity than the planted adult trees as
a result of pollen flow coming from surrounding
natural stands of the same species (Navascués and
Emmerson, 2007).
As stated before, effective gene flow requires
the recruitment of the transported genetic vari-
ants in the new environment, and habitat dis-
turbances can impede or facilitate this process.
For example, despite evidence of extensive pol-
len flow between western continuous forests of
Araucaria araucana and fragmented eastern pop-
ulations (Gallo et al., 2004), no regeneration was
found in several of the fragments because of the
severity of anthropogenic impacts, particularly
the impacts of grazing livestock, animals intro-
duced for hunting and collection of seed by hu-
mans. Thus, effective gene flow was zero and the
sustainability of the whole system is threatened
although gene flow between continuous and
fragmented populations exists (Gallo et al., 2004).
A very well known case in which human activity
affects gene flow among plant species is the crea-
tion of environments that encourage the devel-
opment of interspecific hybrids (Arnold, 1997). In
such human-induced “hybrid habitats” the paren-
tal species can barely survive but the interspecific
hybrids thrive (e.g. Campbell and Wasser, 2007).
In many cases, gene flow between the parental
species could not have been realized without the
environmental disturbance, as has been shown in
Prosopis chilensis and P. flexuosa (Mottuora et al.,
2005).
Gene flow is strongly related to the mating
system and therefore to the fitness of individu-
als and populations. Many forest tree species
have a very strong spatial genetic structure, even
in large, continuous forests. Related individuals
tend to grow in groups because of limited dis-
persal of seeds around the mother tree and/or
spatial directionality of the vectors. This structure
depends mainly on two opposing forces: selection
pressure and gene flow. The natural vectors of
pollen and seeds determine how large the real-
ized mating system can be. Human activities mod-
ify both: the spatial structure of the remaining
adult genotypes defining the distances between
them and the environment in which seed has to
germinate and seedlings have to establish. In spe-

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Effective gene flow
Gene flow should be evaluated on the recruited
regeneration after passive or active restoration. The
following considerations therefore depend on there
being effective gene flow, i.e. establishment and
evolutionary adaptation of the genetic variants.
Genetic diversity
Only genetic diversity can assure current adaptedness
and future population adaptation. Therefore, its
magnitude and distribution have to be known in both
the population to be restored and the population or
trees to be used as donors of genes and/or genotypes
(propagation material). If some fragments are
subjected to genetic diversity restrictions as a result
of demo-genetic processes (bottle necks, genetic
drift, biparental inbreeding etc.), material from other
populations with more genetic diversity (preferentially
from the same seed transfer zone) should be used to
assure evolutionary stability.
Genetic connectivity
Restoration activities must ensure the movement
of genetic information between trees or fragments
within a biological population (individuals that share
an evolutionary path). Dispersion curves and dispersal
distances for pollen, seed or vegetative parts must be
estimated. New molecular methods are increasingly
more precise, cheaper and easy to apply than
traditional methods and facilitate this work. The effect
of potential physical barriers established by humans
(e.g. forestry) should be considered when managing
passive gene flow for genetic connectivity.
Climate-change scenarios for the restoration
area
In the current global scenario of climate change, even
well-adapted populations should be restored using
material from populations that are better adapted
to expected future environmental conditions at the
location being restored. In some cases active or passive
gene flow might be promoted, in general, from drier
environments towards more humid sites.
Isolated trees
In highly fragmented landscapes efforts should be
made to maintain existing isolated trees and to
integrate them in the local socioeconomic system,
since they act as genetic bridges (receiving pollen
and dispersing seeds) and as ecological corridors for
pollinators between fragments and/or continuous
forests. Knowing the pollen movement distance of the
species involved allows development of a network
of isolated trees that can help maintain the genetic
connectivity in such agricultural landscapes.
Unidirectionality of vector movements
The relative location of the fragments restored must
be considered, especially with wind-pollinated species.
In insect-pollinated species “attractor trees” (e.g.
abundantly flowering trees for generalist insects) should
be established at strategic locations within the agricultural
landscape to guide the pollinators’ movements.
Impact on animal vectors
Many animals act as vectors of pollen or seeds,
especially in tropical and subtropical forests. When
restoration activities are implemented, illegal hunting
activities should be forbidden and traditional hunting
activities of local communities should be organized
and controlled to ensure that the vector population is
maintained at a viable level. When restoration activities
are undertaken in strongly fragmented agricultural
landscapes, special consideration should be given to
managing the use of pesticides and herbicides in the
surrounding cultivated areas. The type and amount of
chemicals to be used, the timing of their application and
location of buffer zones should be agreed with local
communities and/or agricultural enterprises to minimize
negative impacts on pollinator-insect populations.
Stand genetic structure
Sustainable stand density should be taken into account
when restoring degraded forest populations. Effective
population number and probable biparental inbreeding
has to be monitored when passive restoration
strategies are implemented.
Box 5.1.
Fundamental considerations for gene-flow management in restoration activities

THE STATE OF THE WORLD’S FOREST GENETIC RESOURCES – THEMATIC STUDY
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72
cies with very short pollination distance, manage-
ment practices can severely alter gene flow and
through it the mating system. For example, pollen
of the southern beech (Nothofagus nervosa) ef-
fectively disperses less than 35 m but an average
of nine pollen parents contribute to each mother
tree, indicating that density of the stand is cru-
cial for the movement of the pollen (Marchelli,
Smouse and Gallo, 2012). When forest manage-
ment activities reduce stand density, a reduction
of the genetic diversity in subsequent generations
of the managed forest is expected, mainly due to
the increase of biparental inbreeding.
References
Arana, M.V., Gallo, L.A., Vendramin, G.G., Pastorino,
M.J., Sebastiani, F. & Marchelli, P. 2010. High ge-
netic variation in marginal fragmented populations
at extreme climatic conditions of the Patagonian
Cypress Austrocedrus chilensis. Mol. Phylogenet.
Evol., 54: 941–949.
Arnold, M.L. 1997. Natural hybridization and evolution.
Oxford Series in Ecology and Evolution. Oxford, UK,
Oxford University Press.
Bacles, C.F.E, Lowe, A.J. & Ennos, R.A. 2007. Effective
seed dispersal across a fragmented landscape.
Science, 311: 628.
Bozzi , J., Marchelli, P., Thomas, L.K., Ziegenhagen,
B., Leyer, I. & Gallo, L. 2011. Impact of introduced
willows on the genetic structure of Salix humboldti-
ana of north Patagonia. In Biolief 2011. 2nd World
Conference on Biological Invasions and Ecosystems
Functioning, November 21–24, 2011, Mar del Plata,
Argentina. Mar del Plata, Argentina, Grupo de
Investigación y Educación en Temas Ambientales.
Budde, K., Gallo, L., Marchelli, P., Mosner, E., Liepelt,
S., Ziegenhagen, B. & Leyer, I. 2011. Wide spread
invasion without sexual reproduction? A case study
on European willows in Patagonia, Argentina. Biol.
Invasions, 13: 45–54.
Campbell, D.R. & Waser, N.W. 2007. Evolutionary
dynamics of an Ipomopsis hybrid zone: confronting
models with lifetime fitness data. Am. Nat., 169(3):
298–310.
Colabella, F. 2011. Análisis del flujo polínico en una
población fragmentada de Ciprés de la Cordillera
(Austrocedrus chilensis). Universidad Nacional de San
Martín, Argentina. (Tesis para optar al título de Lic.
en Biotecnología)
Devlin, B., Roeder, K. & Ellstrand, N.C. 1988.
Fractional paternity assignment, theoretical develop-
ment and comparison to other methods. Theor.
Appl. Genet. 76: 369–380.
Dick, C.W., Etchelecu, G. & Austerlitz, F. 2003. Pollen
dispersal of tropical trees (Dinizia excelsa: Fabaceae)
by native insects and African honeybees in pristine
and fragmented Amazonian rainforest. Mol. Ecol.,
12: 753–764.
Endler, J.A. 1977. Geographic variation, speciation,
and clines. Monographs in Population Biology 10.
Princeton, NJ, USA, Princeton University Press.
Hybridization
Active restoration activities can unintentionally
introduce material from high genetically differentiated
populations into the population to be restored, altering
the natural gene-flow equilibrium. The potential effects
of hybridization between remnant and introduced
material have to be considered since outbreeding
depression may have negative effects on adaptation.
Box 5.1. (continued)
Fundamental considerations for gene-flow management in restoration activities

GENET IC CONS IDERATI ONS IN ECO SYSTEM RE STORATI ON USING N ATIVE TRE E SPECIE S
73
Frankham, R., Ballou, J. & Briscoe, D. 2002.
Introduction to conservation genetics, 1st ed.
Cambridge, UK, Cambridge University Press.
Gallo, L.A. & Pastorino, M.J. 2008. El efecto de la frag-
mentación en la pérdida de la diversidad genética y
su estrategia de conservación. Evidencia de deriva
genética en Ciprés de la Cordillera (Austrocedrus
chilensis). In B. Vinceti, compilador. Estudios de casos
en conservación y utilización de recursos genéticos
forestales, con especial incidencia en Latinoamérica.
Contribución al Curso Internacional “Conservación
y utilización de los recursos genéticos forestales”.
Rome, Bioversity International.
Gallo, L., Letourneau, F. & Vinceti, B. 2004. A model-
ling case study: options for forest genetic resources
management in Araucaria araucana ecosystems. In
B. Vinceti, W. Amaral & B. Meilleur, eds. Challenges
in managing forest genetic resources for livelihoods:
examples from Argentina and Brazil, pp. 187–210.
Rome, International Plant Genetic Resouces Institute.
Gallo, L., Izquierdo, F., Sanguinetti, L.J., Pinna, A.,
Siffredi, G., Ayesa, J., Lopez, C., Pelliza, A.,
Stritzler, M., Gonzalez-Peñalba, M., Maresca, L.
& Chauchard, L. 2004: Arauacaria araucana forest
genetic resources in Argentina. In B. Vinceti, W.
Amaral & B. Meilleur, eds. Challenges in manang-
ing forest genetic resources for livelihood: examples
from Argentina and Brazil. Rome, International Plant
Genetic resources Institute.
Gil, L., Fuentes-Utrilla, P., Alvaro Soto, M., Cervera,
M.T. & Collada, C. 2004. Phylogeography: English
elm is a 2,000-year-old Roman clone. Nature, 431:
1053.
Kremer, A., Ronce, O., Robledo-Arnuncio, J.,
Guillaume, F., Bohrer, G., Nathan, R., Bridle,
J.R., Gomulkiewicz, R., Klein, E., Ritland, K.,
Kuparinen, A., Gerber, S. & Chueler, S. 2012.
Long-distance gene flow and adaptation of forest
trees to rapid climate change. Ecol. Lett., 15(4):
378–392. doi: 10.1111/j.1461-0248.2012.01746.x.
Lowe, A., Harris, S. & Ashton, P. 2004. Ecological ge-
netics: design, analysis and application. Oxford, UK,
Blackwell Publishing.
Marchelli, P., Baier, C., Mengel, C., Ziegenhagen,
B. & Gallo, L. 2010. Biogeographic history of the
threatened species Araucaria araucana (Molina) K.
Koch and implications for conservation: a case study
with organelle DNA markers. Conserv. Genet., 11(3):
951–963.
Marchelli, P., Smouse, P.E. & Gallo, L.A. 2012. Short-
distance pollen dispersal for an outcrossed, wind-
pollinated southern beech (Nothofagus nervosa
(Phil.) Dim. et Mil.). Tree Genet. Genomes, 8(5):
1123–1134.
Mottuora, M.C., Finkeldey, R., Verga, A.R. & Gailing,
O. 2005. Development and characterization of
microsatellite markers for Prosopis chilensis and
Prosopis flexuosa and cross-species amplification.
Mol. Ecol. Notes, 5: 487–489.
Navascués, M. & Emerson, B.C. 2007. Natural recovery
of genetic diversity by gene flow in reforested areas
of the endemic Canary Island pine, Pinus canariensis.
Forest Ecol. Manag., 244: 122–128.
Petit, R.J., Duminil, J., Fineschi, S., Hampe, A., Salvini,
D. & Vendramin, G,G. 2005. Comparative organi-
zation of chloroplast, mitochondrial and nuclear di-
versity in plant populations. Mol. Ecol., 14: 689–701.
Poole, A.L. 1987. Southern beeches. Wellington,
Department of Scientific and Industrial Research,
Science Information Publishing Centre.
Powell, A.H. & Powell, G.V.N. 1987. Population dynam-
ics of male euglossine bees in Amazonian forest
fragments. Biotropica, 19: 176–179.
Sampson, J.F. & Byrne, M. 2008. Outcrossing between
an agroforestry plantation and remnant native
populations of Eucalyptus loxophleba. Mol. Ecol.,
17: 2769–2781.
White, G.M, Boshier, D.H. & Powell, W. 2002.
Increased pollen flow counteracts fragmentation
in a tropical dry forest: an example from Swietenia
humilis Zuccarini. PNAS, 99(4): 2038–2042.
Young, A.G., Merriam, H.G. & Warwick, S.I. 1993. The
effects of forest fragmentation on genetic variation
in Acer saccharum Marsh. (sugar maple) popula-
tions. Heredity, 71: 277–289.

GENET IC CONS IDERATI ONS IN ECO SYSTEM RE STORATI ON USING N ATIVE TRE E SPECIE S
75
Restoration of genetic diversity implies the man-
agement of different levels of genetic exchange,
including hybridization.10 The most common un-
derstanding of hybridization is mating between
related species, but consequences of intraspecific
hybridization must also be considered. When
successful mating occurs naturally between in-
dividuals from two populations, or groups of
populations, that are distinguishable on the ba-
sis of one or more heritable characters, natural
hybridization takes place (Harrison, 1990). One
of the main potential evolutionary outcomes of
intra- or interspecific hybridization is introgres-
sion, which means the movement of the genes
from one population or species into the other
as a result of successive backcrosses (Anderson,
1949). But there are other potential theoreti-
cal outputs, such as increased genetic diversity
through the generation of new gene combina-
tions and genotypes, heterosis, hybrid speciation,
reinforcement of the reproductive barriers that
favour parental speciation, and stabilization of
hybrid zones (Carney, Wold and Rieseberg, 2000).
Recently, hybridization has been highlighted as a
way to regain traits that had been lost and per-
haps to replace damaged alleles with functional
copies from related species (Rieseberg, 2009).
An often mentioned negative consequence of
hybridization is the genetic dilution of a rare
10 Hybridization in the text is taken in a wide sense and as a long
term process including F1, F2, backcrosses, etc., among different
populations of the same species and/or of different species.
population through introgression and exogamic
or outbreeding depression caused by dilution of
the local adaptation and hybrid breakdown (e.g.
disruption of well co-adapted gene complexes)
(Hufford and Mazer, 2003). Additionally, in many
hybridizing systems a strong environmental influ-
ence has been observed in the hybrids’ genera-
tion and fitness. Ecosystem degradation alters en-
vironmental conditions and consequently affects
some biologically important traits such as gene
flow, gamete production and flowering phenol-
ogy that promote hybridization (e.g. Lamont et
al., 2003; Mottuora et al., 2005). In those altered
“hybrid environments” hybrids have an adaptive
advantage over their parents (Arnold, 2006).
6.1. The impact of restoration
Restoration activities can impose genetic con-
nectivity by moving seeds and establishing plants
from the donor population directly into the de-
graded population or by favouring gene flow
between them (active or passive restoration), cre-
ating conditions for hybridization between popu-
lations or species that were not previously in con-
tact. This can have positive or negative impacts,
depending on the situation.
During the genetic restoration process the oc-
currence of natural hybrids can be promoted or
avoided, depending on their expected fitness, the
degradation level of the ecosystem and the final
objectives of the restoration.
Chapter 6
The role of hybridization in
the restoration of forest ecosystems
Leonardo Gallo
Ecological Genetics and Forest Tree Breeding, INTA Bariloche, Argentina

THE STATE OF THE WORLD’S FOREST GENETIC RESOURCES – THEMATIC STUDY
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6.2. Promoting hybridization
Under very strong selection pressures, increased
genetic variation and/or the generation of new
genotypes through hybridization can be adap-
tively advantageous. For example, a controlled
introgression programme has been implemented
to rescue the genetic information and the ecolog-
ical and economic benefits of American chestnut
(Castanea dentata), which was devastated by an
exotic pathogen, chestnut blight (Cryphonectria
parasitica), at the beginning of the twentieth cen-
tury. A programme of controlled hybridization
and introgression incorporated resistance genes
from the Chinese chestnut (Castanea mollisima)
into the genome of the American chestnut with
remarkable success. Recently, candidate genes
for developing resistance have been identified
through the use of advanced molecular technolo-
gies (Barakat et al., 2009).
Controlled hybridization programmes may
become important means to confront climate
change and counteract the negative effects of
drought. Several genetically different donor
populations having drought tolerance could in-
troduce into the degraded ecosystems potentially
genes that could confer better adaptation to fu-
ture climatic conditions.
6.3. Avoiding hybridization
If the divergence between the hybridizing popu-
lations is caused by differences in local selection
(local adaptation), maladapted hybrids would be
expected (Hufford and Mazer, 2003) and hybridi-
zation should be avoided or mitigated.
Such maladaptation has been observed in the
natural hybridization between two Patagonian
southern beeches, Nothofagus nervosa and N.
obliqua (Gallo, Marchelli and Breitembücher,
1997), where selective logging in the past had
substantially reduced the population of N. ner-
vosa and altered the natural pollen competition
equilibrium, promoting the generation of mala-
dapted hybrids (Gallo, 2004). First-generation
hybrids have reduced fitness and the system has
reached a particular equilibrium where backcross-
es are also limited (the “evolutionary novelty”
hybridization model described by Arnold, 1997).
In such situations, restoration activities should
include the reconstruction of the original species
structure of the forest.
If hybrids are well adapted, a large proportion
of the gametes produced by the few individuals
of the degraded population will generate hybrids
and through introgressive backcrosses their ge-
netic information will tend to be diluted; a pro-
cess known as genetic assimilation. At a regional
scale, this is the problem with the European
black poplar (Populus nigra) and the eastern cot-
tonwood (Populus deltoides), introduced from
the United States. Genetic rescue activities have
been carried out to save the few pure individuals
of black poplar since the hybrid (P. × canadensis)
competes for river niches and introgresses its ge-
netic information into the black poplar (Smulders
et al., 2008). Such a situation can be also expected
in intraspecific hybridization when restoring de-
graded populations.
Landscape fragmentation can increase gene
flow, depending on the species’ pollination mecha-
nisms, and with it the negative genetic exchange of
diverging populations. The use of physical barriers
(intermediate forestations) or even removing the
contaminating trees could be possible solutions.
6.4. Seed sources and seed-zone
transfer
When gene flow is restricted among isolated
fragments in a heterogeneous landscape, the
occurrence of strong local adaptation processes
and/or genetic drift divergence has to be consid-
ered. In such cases, the use of local propagation
material is recommended for active restoration
programmes.
Gene flow and fluctuating selection pressure
can reduce the probability of development of
highly localized ecotypes (McKay et al., 2005),
especially where there is extensive pollen flow

GENET IC CONS IDERATI ONS IN ECO SYSTEM RE STORATI ON USING N ATIVE TRE E SPECIE S
77
in long-lived species (Moreno, 2012). Moreover,
in many ecosystems site disturbance introduces a
new challenge for local adaptation and restora-
tion (O´Brien and Krauss, 2010), to which global
climate change impacts should be added. Seed-
zone transfer guidelines should be determined.
To minimize the probability of outbreeding de-
pression, and when lacking a scientifically-based
delineation of seed zones, seed collections should
be made near the restoration site, if populations
of sufficient size and genetic quality are available
as seed sources (Hufford and Mazer, 2003), in or-
der to match climatic and environmental condi-
tions (McKay et al., 2005).
The combination of a network of common-gar-
den field trials across a range of representative
site conditions with molecular genetic diversity
and gene flow analyses could help to better de-
termine operational genetic management units
(Pastorino and Gallo, 2009) that should be taken
into account in any restoration activities.
References
Anderson, E. 1949. Introgressive hybridization. New
York, USA, John Wiley and Sons.
Arnold, M.L. 1997. Natural hybridization and evolution.
Oxford Series in Ecology and Evolution. Oxford, UK,
Oxford University Press.
Arnold, M.L. 2006. Evolution through genetic exchange.
Oxford, UK, Oxford University Press.
Barakat, A., DiLoreto, D.S., Zhang, Y., Smit, C., Baier,
K., Powell, W.A., Wheeler, N., Sederoff, R. &
Carlson, J.E. 2009. Comparison of the transcrip-
tomes of American chestnut (Castanea dentata) and
Chenese chesnut (Castanea mossissima) in response
to the chestnut blight infection. BMC Plant Biol., 9:
51.
Carney, S.E., Wold, D.E. & Rieseberg, L.H. 2000.
Hybridisation and forest conservation. In A. Young,
Boshier, D. & T. Boyle, eds. Forest conservation
genetics, principles and practice, pp. 167–182.
Collingwood, Victoria, Australia, CSIRO Publishing,
and Wallingford, UK, CABI Publishing.
Gallo, L. 2004. Modelo conceptual sobre la hibridación
natural interespecifica entre Nothofagus nervosa
y N. obliqua. In C. Donoso, A. Premoli, L. Gallo &
R. Ipinza, eds.Variación intraespecífica en espe-
cies arbóreas de los bosques templados de Chile
y Argentina, pp. 397–408. Editorial Universitaria,
Chile.
Gallo, L.A., Marchelli, P. & Breitembücher, A. 1997.
Morphologycal and allozymic evidence of natu-
ral hybridization between two southern beeches
(Nothofagus spp.) and its relation to heterozygosity
and height growth. Forest Genet., 4(1): 13–21.
Harrison, R.G. 1990. Hybryd zones: windows on evolu-
tionary process. In D. Futuyma & J. Antonovics, eds.
Oxford Surveys in Evolutionary Biology 7: 69–128.
Hufford, K.M. & Mazer, S.J. 2003. Plant ecotypes:
genetic differentiation in the age of ecological resto-
ration. Trends Ecol. Evol., 18: 147–155.
Lamont, B.B., He, T., Enright, N.J., Kraus, S. & Miller,
B.P. 2003. Anthropogenic disturbance promotes
hybridization between Banksia species by altering
their biology. J . Evol. Biol., 16: 551–557.
McKay, J.K., Christian, C.E., Harrison, S. & Rice, K.J.
2005. “How local is local?”– A review of practical
and conceptual issues in the genetics of restoration.
Restor. Ecol., 13: 432–440.
Moreno, C. 2012. Estudio del flujo génico mediado por
polen en poblaciones fragmentadas de Araucaria
araucana. Universidad Nacional del Comahue,
Argentina. (Tesis pata optar al grado de Doctor en
Biología)
Mottuora , M.C., Finkeldey, R., Verga, A.R. & Gailing,
O. 2005. Development and characterization of
microsatellite markers for Prosopis chilensis and
Prosopis fl exuosa and cross-species amplification.
Mol. Ecol. Notes, 5: 487–489.
O´Brien, E.K. & Krauss, S.L. 2010. Testing the home-
site-advantage in forest trees on disturbed and
undisturbed sites. Restor. Ecol., 18: 359–372.

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Pastorino, M. & Gallo, L. 2009. Preliminary operational
genetic management units of a highly framented
forest tree species of southern South America. Forest
Ecol. Manag., 257: 2350–2358.
Rieseberg, L.H. 2009. Replacing genes and traits
through hybridization. Curr. Biol., 19: 119–122.
Smulders, M.J.M, Beringe, R., Volosyanchuk, R.,
Vande Broeck, A., van der Schoot, J., Arens, P.
& Vosman, B. 2008. Natural hybridisation between
Populus nigra L. and P. × canadensis Moench. Hybrid
offspring competes for niches along the Rhine
river in the Netherlands. Tree Genet. Genomes, 4:
663–675.

GENET IC CONS IDERATI ONS IN ECO SYSTEM RE STORATI ON USING N ATIVE TRE E SPECIE S
79
Genetic variation within and among popula-
tions has not been studied in the vast majority
of tree species. This makes it difficult to plan ef-
fective germplasm collection strategies for forest
restoration and species conservation purposes.
This paper provides guidelines for collection
of propagation material for forest restoration
when knowledge of genetic variation between
and within populations is lacking. Rare species
that occur as scattered trees or in small cohorts
growing hundreds of metres apart are unsuitable
sources of propagation material for stand estab-
lishment since they never form stands in nature.
Therefore, this paper refers to commonly occur-
ring species.
Forest restoration projects may also aim to con-
serve the genetic diversity of the species used, but
for a specific treatment of sampling for gene con-
servation in sensu stricto, the reader is referred
to Eriksson (2005a). Evolutionary factors are pre-
sented to help understand the existing genetic
variation between and within populations.
7.1. Evolutionary factors
In nature there is constant interaction among evo-
lutionary factors, with the result that it is hard to
know or predict the genetic variation in a species
(Mayr, 1988; Eriksson, 2005b). Evolutionary factors
are briefly discussed here to elucidate their role
in evolution. (See Box 7.1 for definition of terms.)
Natural selection requires that there is genetic
variation for traits contributing to fitness (Endler,
1986). Changes in gene frequencies are depend-
ent on existing conditions; future conditions have
no influence over them. Thus, there is no goal or
predetermined direction of selection. Moreover,
most fitness traits are complex. For a tree they
may consist of growth rhythm, growth rate, toler-
ance of adverse climatic factors, tolerance of pests
or diseases, and uptake and utilization of nutri-
ents. It is highly unlikely that natural selection
influences these components individually; rather
it is the whole individual that is the “target” in
natural selection. For most traits of significance in
evolution we find a bell-shaped curve for the dis-
tribution of individuals. In many cases of natural
selection the individuals in the centre of the dis-
tribution are favoured (stabilizing selection) but
in some cases the individuals in one of the tails
of the distribution are favoured (directional se-
lection). Selection that favours individuals in the
two tails of the distribution is known as disruptive
selection. This type of selection probably occurs
rarely within populations.
Genetic drift causes random losses of genes,
the rate of loss increasing with decreasing num-
ber of reproductive individuals in a population.
At a constant population of ten fruiting trees for
ten generations genetic variation (or more cor-
rectly “additive variance”) falls to approximately
half of the original (Eriksson, 1998). A popula-
tion of 20 individuals would lose approximately
Chapter 7
Collection of propagation material
in the absence of genetic knowledge
Gösta Eriksson
Department of Plant Biology and Forest Genetics, Swedish University for Agricultural Sciences,
Sweden

THE STATE OF THE WORLD’S FOREST GENETIC RESOURCES – THEMATIC STUDY
PART 2
80
20percent of its additive variance. The difference
in loss between populations of 500 and 1000 is
only 0.5 percent per generation. This means that
not much is gained by having thousands of fruit-
ing trees in gene-resource populations. A popula-
tion of 50 fruiting trees has been regarded as a
satisfactory size for a gene-resource population
and for sourcing propagation material. With such
a population size the loss of additive variance is
just 1 percent per generation.
Gene flow caused by pollen and seed transfers
between populations can be considerable in wind-
pollinated species but less in species with limited
pollen and seed dispersal (Govindaraju, 1988).
Gene flow is such a strong levelling factor that
only one migrant per generation prevents differ-
entiation in neutral genes (genes not influencing
fitness) in a randomly mating population.
Mutations occur at low frequencies, and there-
fore do not exert any strong influence on evo-
lution in the short term. Mutations are of great
significance in the long term because they create
genetic variation for the other evolutionary fac-
tors to act upon.
The impact of within-population and between-
population variation of natural selection, genetic
drift and gene flow are summarized in Table 7.1.
Stabilizing natural selection within populations
leads to less overlapping of the adjacent popu-
lations. This in turn means that there will be a
sharpening of the differences between popula-
tions. Since the loss of genes as a result of genetic
drift is random, different genes will be lost in dif-
ferent populations, which also leads to increased
variation between populations. The effect of gene
flow on variation within and between populations
is evident from the definition of gene flow. Thus,
gene flow reduces differences between popula-
tions but increases the within-population vari-
ation. Understanding the principles behind the
distribution of variation helps to determine how
many different populations should be consid-
ered as sources of propagation material in order
to adequately capture the genetic variation of a
species. It can guide species restoration efforts or
help to prepare for environmental changes with
diverse propagation material.
7.2. Methods for sampling
diversity
Before materials are selected for forest restora-
tion, it is useful to examine the abiotic and biotic
factors at the reforestation site. Abiotic factors
such as climate, nutrient availability and photo-
periodic conditions might be easy to determine,
while biotic factors such as pests and diseases
might be hard to identify in advance. At least
one, but usually more, of these factors is signifi-
cant at a particular reforestation site. For exam-
ple, photoperiodic conditions are of significance
at high latitudes where this factor varies consider-
ably. With only basic knowledge of genetic diver-
sity and biotic and abiotic factors we have to rely
on educated guesses to develop a sampling strat-
egy for our target species. Generally, sampling in
the absence of genetic knowledge should take
place at localities with closest similarity with the
reforestation sites.
7.3. Genetic variation
Hamrick, Linhart and Mitton (1979) hypothesized
that life history traits such as species distribution,
stage in ecosystem – pioneers versus climax spe-
cies – or wind pollination vs insect pollination
would influence genetic variation within and
between species (Figure 7.1). However, in his
TABLE 7.1.
The impact of natural selection, genetic drift
and gene flow on genetic variation within and
between populations
Variation within
populations Variation between
populations
Natural selection decrease increase
Genetic drift decrease increase
Gene flow increase decrease

GENET IC CONS IDERATI ONS IN ECO SYSTEM RE STORATI ON USING N ATIVE TRE E SPECIE S
81
study of adaptive traits in several tree species,
Baliuckas (2002) reported only weak support for
this hypothesis. Until more research results are
obtained, Hamrick, Linhart and Mitton’s (1979)
hypothesis may still be used to guide sampling.
Sampling existing adaptations is simple if all
variation is included in every population and
not between them. Then, it suffices to sample
enough trees to avoid inbreeding. However, ab-
sence of genetic variation among populations is
not known to occur in any tree species other than
the red pine (Pinus resinosa Ait.).
Because the evolutionary factors act in a com-
plex way, we have to simplify the possible effect
of their interactions on genetic diversity. The ef-
fect of genetic drift can be excluded if mating is
random, leaving the two opposing factors natu-
ral selection and gene flow. Figure 7.2 illustrates
the possible between-population differentia-
tion for various combinations of gene flow and
disruptive natural selection. A prerequisite for
Figure 7.1.
The expected effects of life history traits on
genetic variation within and between populations
Source: Hamrick, Linhart and Mitton (1979)
Figure 7.2.
The possible genetic variation between populations in random-mating species, as affected by varying
strengths of disruptive natural selection and gene flow

THE STATE OF THE WORLD’S FOREST GENETIC RESOURCES – THEMATIC STUDY
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82
differentiation between populations is that the
species experiences the biotic and abiotic site
conditions in its area of distribution as hetero-
geneous. In the absence of disruptive selection
there may be some differentiation of populations
for random reasons (position 1 in Figure 7.2). The
larger the variation of the site conditions the
greater the differentiation between populations;
differentiation will be greatest in the absence of
gene flow (position 2). When both gene flow and
natural selection are strong, populations may still
differentiate (position 3).
It is obvious that species covering large ar-
eas, such as Norway spruce (Picea abies L. Karst.)
and Scots pine (Pinus sylvestris L.) in Europe
or Douglas-fir (Pseudotsuga menziesii (Mirb.)
Franco) and Lodgepole pine (Pinus contorta
Douglas) in North America, face extreme varia-
tion in site conditions. It is not only the climate
that varies in their distribution areas but also
soil conditions. Such a large variation in site
conditions causes population differentiation
(Dietrichson, 1961; Eiche, 1966; Rehfeldt, 1989;
Lindgren et al., 1994). Since these species also
are wind pollinated they are examples of spe-
cies with high disruptive selection and large gene
flow (position 3 in Figure 7.2). These four spe-
cies experience large differences in climate over
their distribution areas. For this type of species,
numerous populations should be considered as
sources of propagation material for reforesta-
tion. Still larger numbers of populations have to
be included if the species grow on contrasting
edaphic conditions within a climatic zone.
Red pine is distributed over a large area in east-
ern North America without much population dif-
ferentiation (Mosseler, Egger and Hughes, 1992).
Strangely, there also seems to be very limited
variation within populations. This lack of varia-
tion has been attributed to a genetic bottleneck
after the last glaciation. Red pine is an example
of a species with weak disruptive selection (po-
sition 1 in Figure 7.2). Since there is one almost
homogenous population, in theory seed could be
collected from anywhere in the range of the spe-
cies for planting anywhere. In reality, it would be
prudent to exercise some caution. Gene flow does
not occur in this species.
There are few examples of forest tree species
that would experience natural selection in the ab-
sence of gene flow (position 2 in Figure 7.2). The
Fraser fir (Abies fraseri (Pursh) Poir.), may have
such a genetic structure. The species comprises
seven populations in North Carolina, Tennessee
and Virginia, United States, mainly at more than
1500 m above sea level (Pauley and Clebsch, 1990).
However, a large part of the differentiation is
probably due to genetic drift. For such a species,
all populations should be designated as gene re-
source populations. Bearing in mind global warm-
ing, planting at higher elevation than the present
distribution of Fraser fir is recommended if fund-
ing for such an approach could be raised. This rec-
ommendation is also valid for other species with
similar characteristics to those of the Fraser fir.
Evolution = change of genetic constitution
Adaptation = the process of genetic change of a
population, owing to natural selection, resulting in a
better adaptedness
Adaptedness = the degree to which an organism is
able to live and reproduce in a given environment
Adaptability = the ability to respond genetically or
phenotypically to changed environmental conditions
Evolutionary factors
Natural selection = improvement of adaptedness via
differential transfer of genes to the next generation
Genetic drift = random loss of genes in small
populations
Gene flow = migration to a recipient population from
another population with a different gene frequency
Mutation = a chemical or structural change of DNA
Random mating = each tree in a population has an
equally large chance to take part in the fertilization
as all other trees in this population
Box 7.1.
Definition of terms

GENET IC CONS IDERATI ONS IN ECO SYSTEM RE STORATI ON USING N ATIVE TRE E SPECIE S
83
7.4. Avoidance of genetic drift
Once it is decided which populations should be
used for collection of material, sampling should
encompass enough trees to avoid narrowing the
genetic variation in the stand to be established.
Insufficient variation might lead to loss of the
whole new stand as a result of pests, diseases or
adverse abiotic factors. Because a restored stand
is expected to be self-perpetuating, it is impor-
tant to avoid inbreeding.
Recommendations have been formulated to
guide sampling for propagation purposes (e.g.
Dawson and Were, 1997). It is commonly sug-
gested to collect germplasm from a minimum of
30 trees. To avoid offspring of related trees oc-
curring in the sample, a minimum distance of
50–100m between the collected trees is suggest-
ed. Broad genetic diversity of the propagation
material will not only ensure a viable population,
but will most likely be advantageous for adapta-
tion to changing environmental conditions.
Selfing and other forms of inbreeding in cross-
fertilizing trees cause strong inbreeding depres-
sion; stronger the closer the relatedness. Thus,
selfing leads to much stronger depression than
cousin matings. In one of the oldest field trials
with a selfed forest tree, Norway spruce, the in-
breeding depression of stem volume at 60 years
of age was substantial and amounted to approxi-
mately 50 percent reduction in growth (Eriksson,
Schelander and Åkebrand, 1973).
The European white elm (Ulmus laevis Pall.) in
southern Finland is one example of a species expe-
riencing natural selection in the absence of gene
flow (position 2 in Figure 7.2). This species has high
between-population variation in Finland, mainly
attributed to genetic drift (Vakkari, Rusanen and
Kärkkäinen, 2009). Between-population varia-
tion mediated by genetic drift does not result
in high adaptedness in all small populations. To
obtain good reforestation material it might be
useful to put together trees in seed orchards or
clone archives, as suggested for conservation of
the European white elm in Finland, in which two
to ten clones from each of 19 populations were
planted in ex situ plantations for seed produc-
tion (Vakkari, Rusanen and Kärkkäinen, 2009).
Progenies from such plantations will be exposed
to natural selection, frequently resulting in change
of the genetic constitution. In this way the effects
of genetic drift will be reduced and the genetic
variation will be increased. The number of clones
in such plantations should preferably be 50 but an
absolute minimum of 20 must be met.
Considering predicted global warming, measures
could be taken to mitigate the effects of a rapid en-
vironmental change (Figure 7.3). Each subpopula-
tion should ideally consist of 50 trees. The illustrated
principle may also be applied in conditions where it
is desirable to support migration of a species.
Figure 7.3.
The oval area represents the current distribution
of a species along a climatic gradient from
a warm climate at the bottom to a cooler
climate at the top. The green circles are gene
resource subpopulations. Materials from the
subpopulations should be transferred to a
cooler climate to mitigate the impact of climatic
warming. It might be wise to establish some
subpopulations outside the present range of
distribution (dark green circle).

THE STATE OF THE WORLD’S FOREST GENETIC RESOURCES – THEMATIC STUDY
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7.5. Conclusion
In summary, estimates of the distribution of ge-
netic variation within and between populations
enable genetically solid conservation and also
promote adaptation in species restoration efforts.
Consideration of the variation between popula-
tions is especially important for species with high
disruptive selection and limited gene flow (po-
sition 2 in Figure 7.2), followed by species with
high disruptive selection and large gene flow (po-
sition 3 in Figure 7.2). For species with low disrup-
tive selection and limited gene flow (position 1
in Figure 7.2), much of the genetic variation oc-
curs within the population and less between the
populations, which means that restoration mate-
rial from a few populations would be sufficient.
References
Baliuckas, V. 2002. Life history traits and broadleaved
tree genetics. Acta Universitatis Agriculturae
Sueciae: Silvestria 258. Uppsala, Sweden, Swedish
University of Agricultural Sciences.
Dawson, I. & Were, J. 1997. Collecting germplasm for
trees – some guidelines. Agroforest. Today, 9: 6–9.
Dietrichson, J. 1961. Breeding for frost resistance. Silvae
Genet., 10: 172–179.
Eiche, V. 1966. Cold damage and plant mortality in
experimental provenance plantations with Scots pine
in northern Sweden. Studia Forestalia Suecica 36.
Umeå, Sweden, Skogshögskolan.
Endler, J.A. 1986. Natural selection in the wild.
Princeton, NJ, USA, Princeton University Press.
Eriksson, G. 1998. Evolutionary forces influencing vari-
ation among populations of Pinus sylvestris. Silva
Fenn., 32: 173–184.
Eriksson G. 2005a. Selection of target species and
sampling for genetic resources in absence of
genetic knowledge. In T. Geburek & J. Turok, eds.
Conservation and management of forest genetic
resources in Europe, pp. 391–411. Zvolen, Slovakia,
Arbora Publishers.
Eriksson, G. 2005b. Evolution and evolutionary factors,
adaptation, adaptability. In T. Geburek & J. Turok,
eds. Conservation and management of forest
genetic resources in Europe, pp. 199–211. Zvolen,
Slovakia, Arbora Publishers.
Eriksson, G., Schelander, B. & Åkebrand, V. 1973.
Inbreeding depression in an old experimental planta-
tion of Picea abies. Hereditas, 73: 185–194.
Govindaraju, D.R. 1988. Relationship between dispersal
ability and levels of gene flow in plants. Oikos, 52:
31–35.
Hamrick, J.L., Linhart, Y.B. & Mitton, J.B. 1979.
Relationships between life history characteristics and
electrophoretically detectable genetic variation in
plants. Annu. Rev. Ecol. Syst., 10: 173–200.
Lindgren, D., Ying, C.C., Elfving, B. & Lindgren, K.
1994. Site index variation with latitude and altitude
in IUFRO Pinus contorta provenance experiments in
western Canada and northern Sweden. Scand. J.
For. Res., 9: 270–274.
Mayr, E. 1988. Toward a new philosophy of biology.
Observations of an evolutionist. Cambridge, MA,
USA, Harvard University Press.
Mosseler, A., Egger, K.N. & Hughes, G.A. 1992. Low
levels of genetic diversity in red pine confirmed by
random amplified polymorphic DNA markers. Can. J.
For. Res., 22: 1332–1337.
Pauley, E.F. & Clebsch, E.C. 1990. Patterns of Abies
fraseri regeneration in a Great Smoky Mountains
spruce–fir forest. Bull. Torrey Bot. Club, 117:
375–381.
Rehfeldt, G.E. 1989. Ecological adaptations in Douglas-
fir (Psuedotsuga menziesii var. glauca): a synthesis.
Forest Ecol. Manag., 28: 203–215.
Vakkari, P., Rusanen, M. & Kärkkäinen, K. 2009. High
genetic differentiation in marginal populations of
European white elm (Ulmus laevis). Silva Fen., 43(2):
185–196.

GENET IC CONS IDERATI ONS IN ECO SYSTEM RE STORATI ON USING N ATIVE TRE E SPECIE S
85
Despite ongoing pressures to clear tropical for-
ests, there is also substantial interest in their res-
toration and tropical forest cover is increasing in
certain regions (Asner et al., 2009). The motiva-
tion for restoring tropical forests comes from an
interest in enhancing or restoring the delivery of
ecosystem services (e.g. sequestering carbon, min-
imizing erosion, improving water quality, main-
taining hydrological cycling and harbouring bio-
diversity) and the maintenance of natural capital.
Most tropical forest restoration efforts focus on
reintroducing tree species to accelerate forest re-
covery (Holl, 2012). There are three principal meth-
ods of tree propagation used to restore former
agricultural lands in the tropics: (1) planting seed-
lings grown in nurseries from seed; (2) vegetative
propagation of individuals, either directly onsite or
in nurseries; and (3) direct seeding into a restora-
tion site. Which of these strategies to use depends
on the goals of the project, the natural rate of re-
covery and the ecology of the system (Holl, 2012).
Germinating and establishing seedlings from
seed in nurseries is the predominant form of tree
propagation and the most widely used method
throughout the tropics. The two other techniques
are emerging as viable potential alternatives be-
cause they are less labour-intensive and cheaper.
The establishment of trees by vegetative means
has typically centred on using small cuttings from
young branches and shoots of trees or larger
branches that are often referred to as stakes.
Direct seeding refers to the mass seeding of a spe-
cies or group of species into a restoration site at
the onset of a project, or more-targeted seeding
of typically mid- to late-successional species at
later stages in the recovery process.
In this review we outline each propagation
strategy and provide a summary of their relative
advantages and disadvantages. In presenting case
studies we draw heavily upon the research that the
authors have performed in southern Costa Rica, as
all three propagation methods have been evaluat-
ed in studies at the same research sites (Figure 8.1).
These case studies have been conducted on land
formerly used for cattle grazing or for growing
coffee for at least 18 years. Sites are in the tropical
premontane rain-forest zone, range in elevation
from 1000–1500 m above sea level, and receive a
mean annual rainfall of about 3500–4000 mm with
a dry season from December to March.
8.1. Establishing tree seedlings
from seed in nurseries
Establishing tree seedlings in nurseries from seed
is by far the most common strategy used to prop-
agate trees for restoration. Studies have either
focused on establishing a broad range of spe-
cies to create a baseline forest community (e.g.
Chapter 8
Evaluation of different tree
propagation methods in ecological
restoration in the neotropics
R.A. Zahawi1 and K.D. Holl2
1 Las Cruces Biological Station, Organization for Tropical Studies, Costa Rica
2 Environmental Studies Department, University of California, Santa Cruz, United States

THE STATE OF THE WORLD’S FOREST GENETIC RESOURCES – THEMATIC STUDY
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86
Butterfield, 1995; Parrotta and Knowles, 2001;
Rodrigues et al., 2009) or on using a few species
as nurse trees to create infrastructure at a resto-
ration site and accelerate the natural process of
forest recovery (e.g. Holl et al., 2011). Researchers
typically either work with a tree nursery to obtain
large numbers of seedlings (Carpenter, Nichols
and Sandi, 2004; Holl et al., 2011), harvest their
own seeds from an adjacent forest and germinate
and establish them in shade houses (e.g. Butter-
field, 1995, 1996; Elliott et al., 2003; Carpenter,
Nichols and Sandi, 2004) or transplant seedlings
that have established in a natural setting (often
referred to as “wildlings”) directly into a restora-
tion site (Parrotta and Knowles, 2001). The diver-
sity of species in a nursery is often limited to a
few native and exotic commercially viable species,
although the range is increasing in some regions
as restoration efforts become more widespread.
For example, 60–80 species are often available in
nurseries in southeastern Brazil, where there are
extensive restoration projects in the Atlantic for-
est region (Rodrigues et al., 2009).
Establishing seedlings in nurseries from seed
can yield large numbers of individuals. Seedlings
are typically transplanted when they reach 20–40
cm in height (Carpenter, Nichols and Sandi, 2004;
Holl et al., 2011). In former pasture lands that are
dominated by aggressive African pasture grasses,
surrounding above-ground vegetation should be
cleared periodically for 2–3 years to reduce com-
petition and shading (Butterfield, 1995; Holl et
al., 2011); if this is not done (and sometimes even
if it is), seedling mortality can be very high. When
a full canopy cover is obtained, maintenance is
no longer necessary as competition from pasture
grasses and other ruderal vegetation is decreased
as a result of shading.
Figure 8.1.
Location of the experimental sites used to test tree-propagation methods in Costa Rica

GENET IC CONS IDERATI ONS IN ECO SYSTEM RE STORATI ON USING N ATIVE TRE E SPECIE S
87
Establishment of stands with either pure or
mixed tree species has been broadly demonstrat-
ed to be successful in the tropics (e.g. Butterfield,
1995; Haggar, Wightman and Fisher, 1997; Lamb,
1998; Montagnini, 2001; Parrotta and Knowles,
2001; Calvo-Alvarado, Arias Richter, 2007; Butler,
Montagnini and Arroyo, 2008). Survival and
growth of different species can vary widely, how-
ever, and it appears that some are more able than
others to tolerate the stressful microclimate and
nutrient conditions found in degraded tropi-
cal landscapes (Butterfield, 1995; Parrotta and
Knowles, 2001; Carpenter, Nichols and Sandi,
2004). The successful establishment of a given spe-
cies can be highly site-specific (Butterfield, 1996).
A number of authors have suggested that some
large-seeded and shade-tolerant species are bet-
ter introduced at later stages in succession, once
an overstory canopy cover has developed and
conditions are more favourable to seedling estab-
lishment (Parrotta and Knowles, 2001; Martínez-
Garza and Howe, 2003; Cole et al., 2011).
Lack of genetic variability among seedlings is a
concern in many tropical forest restoration pro-
jects. Due to logistical constraints, most nursery
endeavours (commercial and non-commercial)
have often harvested seed from fewer than ten
mother trees (Butterfield, 1995; Carpenter, Nichols
and Sandi, 2004), and this can have strong rami-
fications for the long-term fitness of populations
(Carpenter et al., 1995). Once a canopy cover is es-
tablished, recruitment of naturally dispersed spe-
cies can be quite high, resulting in rapid forest re-
covery (Jones et al., 2004; Butler, Montagnini and
Arroyo, 2008), although the community composi-
tion of species can vary widely depending upon
the nurse species planted (Parrotta and Knowles,
2001; Carnevale and Montagnini, 2002) and the
availability of local propagules (Holl, 2007).
Case study
Holl et al. (2011) established a long-term resto-
ration study spread across 100 km2 in southern
Costa Rica in 2004–2006. The 14 study sites are
located between the Las Cruces Biological Station
(8°47’7’’ N; 82°57’32’’ W) and the town of Agua
Buena (8°44’42’’ N; 82°56’53’’ W). Each site incor-
porates three 50 × 50 m treatment plots – two
active restoration plots and one passive or control
restoration plot. Seedlings of four tree species
(Terminalia amazonia (J.F. Gmel.) Exell, Vochysia
guatemalensis Donn. Sm., Erythrina poeppigiana
(Walp.) Skeels and Inga edulis Mart.) were estab-
lished in two planting designs at each site – a
plantation-style planting where the entire 50 × 50
m area was planted, and an island planting where
trees were planted in six different-sized patches
within the 50 × 50 m plot. The four species cho-
sen are characterized by high regional survival,
rapid growth and extensive canopy development
(Nichols et al., 2001; Carpenter, Nichols and Sandi,
2004; Calvo-Alvarado, Arias and Richter, 2007).
Terminalia amazonia and V. guatemalensis are na-
tive timber species that produce valuable timber
and favour establishment of native woody spe-
cies in their understory (Cusack and Montagnini,
2004). Erythrina poeppigiana and Inga edulis are
naturalized, fast-growing nitrogen-fixing species.
Both are widely used in agricultural intercropping
systems to provide shade and increase soil nutri-
ents, and have extensive branching architecture;
I. edulis also produces fruit that can attract birds
(Pennington and Fernandes, 1998; Nichols et al.,
2001; Jones et al., 2004). All four species were pur-
chased from a local nursery.
All sites were cleared of above-ground vegeta-
tion with machetes prior to planting. Seedlings av-
eraged 20–30 cm in height when planted. Ruderal
vegetation was cleared every two to three months
at all sites for 2.5 years. Establishment of planted
seedlings was highly successful (>90 percent) and
some sites reached canopy closure within two to
three years (Holl et al., 2011). However, growth
rates were highly variable among sites and some
have yet to develop a fully closed canopy even six
years into the study. The reasons behind this dis-
parity are unclear, however, but are probably re-
lated to prior land use. Seed dispersal and tree re-
cruitment at this stage in the study (six years after
planting) is largely comprised of early-successional
species with few mid- to late-successional spe-
cies (N = 55 species by 2010 survey; Cole, Holl and

THE STATE OF THE WORLD’S FOREST GENETIC RESOURCES – THEMATIC STUDY
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Zahawi, 2010; Zahawi et al., 2013). Overall, the
strategy of planting a few widely available and
hardy nurse species to accelerate natural forest
recovery has been highly successful. However, the
broad variation in both establishment and growth
of planted seedlings, as well as the huge variation
in seed dispersal and subsequent seedling estab-
lishment among sites, strongly underscore the im-
portance of broadly replicating restoration studies
across the landscape to avoid reaching erroneous
conclusions based on a few sites.
8.2. Establishment by vegetative
propagation
Vegetative propagation has been an integral
technique for the establishment of trees in tropi-
cal agriculture, especially in the humid tropics,
for many decades. A few commercial species of
trees are also propagated vegetatively, such as
Pochote (Pachira quinata (Jacq.) W. S. Alverson;
also known as Bombacopsis quinatum (Jacq.)
Dugand) (Hunter, 1987), beechwood (Gmelina
arborea Roxb.) (Romero, 2004) and teak (Tectona
grandis Linn. f.) (Husen and Pal, 2007). Whereas
vegetative propagation has been used extensively
in tropical agriculture and silviculture, it has re-
ceived relatively little attention as a method for
tree propagation in tropical restoration thus far
(but see Perino, 1979; Ray and Brown, 1995; Chap-
man and Chapman, 1999; Granzow de la Cerda
and Garth, 1999; Zahawi, 2005).
There are two main forms of vegetative propa-
gation: (1) cuttings, which are typically 20–40
cm long, taken from young branches or shoots
of trees; and (2) stakes, which are typically 2–2.5
m long, taken from branches that are pollarded
from trees or extant live fence rows.
The establishment of trees from cuttings has
several advantages, including ease of transport,
availability in considerable quantities (once a
mother tree is located, a considerable number
of cuttings can be harvested), speed of planting
(particularly if planted directly into the restora-
tion site) and cost effectiveness. The application
of cuttings for tree propagation in restoration ac-
tivities has been limited to a few studies that have
focused on the methodology (e.g. Ray and Brown,
1995; Itoh et al., 2002; Bonfil-Sanders, Mendoza-
Hernandez and Ulloa-Nieto, 2007), but cuttings
have been widely used for enrichment planting
of dipterocarp forests in Indonesia (Kettle, 2010).
Most studies have reported mixed success, with
high failure rates of a number of species despite
the application of rooting hormones. A few stud-
ies have evaluated the possibility of using cut-
tings to propagate rare or endangered species
(Danthu, Ramaroson and Rambeloarisoa, 2008;
Ratnamhin, Elliott and Wangpakapattanawong,
2011), with some success with some species.
Itoh et al. (2002) found that rooting ability
of 100 tropical trees in Malaysia was related to
plant family and the growth characteristics of the
species; fast-growing species that were generally
of smaller mature stature typically rooted more
readily. The ability to establish is also related to
the type of cutting used; mature branches (har-
vested further down a stem) establish more readi-
ly than apical cuttings (Dick et al., 1998; Danthu et
al., 2002) and leafy cuttings appear more success-
ful at rooting than leafless cuttings (Brennan and
Mudge, 1998; Dick et al., 1998). Seasonality of
timing when cuttings are harvested can also influ-
ence establishment success (Danthu, Ramaroson
and Rambeloarisoa, 2008).
In contrast to cuttings, stakes have been widely
used in agricultural practice throughout south-
ern Mexico and Central America, especially in the
humid tropics. Although the predominant use of
the technique has been to establish live fences,
stakes have also been used as host plants for ag-
ricultural crops such as vanilla and black pepper,
and in some instances for erosion control (Perino,
1979; Sauer, 1979; Budowski, 1987; Budowski and
Russo, 1993). In addition to these functions, trees
often provide other benefits such as nitrogen fixa-
tion to improve soil quality, shade for coffee, fod-
der for cattle and firewood (Budowski and Russo,
1997; Martínez-Betancourt, Ramírez-Molinet
and Rodríguez-Durán, 2000; Harvey et al., 2005).
Such species are also widespread throughout

GENET IC CONS IDERATI ONS IN ECO SYSTEM RE STORATI ON USING N ATIVE TRE E SPECIE S
89
the landscape and have a proven track record
of being hardy, withstanding not only the harsh
conditions found in pastures but also tolerating
extensive and repeated pollarding and other ag-
ricultural practices.
Stakes are typically planted as 2–3-m-tall
branches ranging in diameter from 4 to 12 cm in-
serted directly into a planting site at a depth of
20–30 cm (Budowski and Russo, 1993; Martínez-
Betancourt, Ramírez-Molinet and Rodríguez-
Durán, 2000; Zahawi, 2005), although stakes
at tall as 4–4.5 m can also be established read-
ily (Zahawi, 2008). Accordingly, some degree of
above-ground vertical stratification can be cre-
ated at the time of planting. Establishment suc-
cess appears to vary widely and is dependent
on a number of variables, including geographic
location, elevation, rainfall and planting season
(Budowski and Russo, 1993; Alonso et al., 2001;
Zahawi, 2005). Initial stake size (both height and
diameter) affects survival and growth and can also
have an impact on biomass production rates (da
Costa et al., 2004; Zahawi, 2005; 2008; Zahawi and
Holl, 2009). Stakes also develop greater above-
and below-ground biomass than seedlings in the
initial years after planting; below-ground archi-
tecture is also distinctly different, with stakes pro-
ducing extensive lateral roots while lacking a cen-
tralized taproot (Zahawi and Holl, 2009). Whereas
most farmers consider it important to plant stakes
just after a full moon (Budowski and Russo, 1993),
the effect of moon phase has been examined em-
pirically in only one study; only slight differences
were found in a few growth indicators but not for
survival (Alonso et al., 2002).
Although stakes have long been used in agri-
cultural practice and there is often widespread
local knowledge of how to establish the species
that are utilized in a given location, much of the
information on species establishment, such as
timing and seasonality of planting, has not been
published or quantified experimentally (but see
Alonso et al., 2001; Zahawi, 2005). This informa-
tion is traded among practitioners and stakehold-
ers, and has occasionally been compiled in anec-
dotal form (e.g. Sauer, 1979; Budowski and Russo,
1993). Although the literature documents several
hundred species that can establish vegetatively
(Budowski and Russo, 1993; Martínez-Betancourt,
Ramírez-Molinet and Rodríguez-Durán, 2000;
Harvey et al., 2005), farmers overwhelmingly rely
on only a few species with widespread distribu-
tion and use; however, species choice does vary
regionally and by country. Farmers’ species selec-
tion is focused naturally on features that are im-
portant to them, e.g. species that are not toxic
to livestock, can hold barbed wire and are able
to withstand regular pollarding (Sauer, 1979;
Budowski and Russo, 1993). In contrast, restora-
tion ecologists would likely focus on species with
fruit that attract frugivores, an ability to shade
out pasture grasses, extensive canopy architec-
ture and rapid growth rates. Accordingly, studies
are needed to better document the establish-
ment needs and abilities of species of interest
to restoration. In addition, an evaluation of the
functional traits shared among species that estab-
lish vegetatively would be particularly useful and
would facilitate the search for potential forest
species that could be of value to restoration.
Case study
Plots were established at three field sites in Costa
Rica to evaluate growth and survival of stakes
and compare their performance with stand-
ard nursery-raised seedlings (Zahawi and Holl,
2009). At each site, stakes were harvested from
20–30 individual fence trees of each of ten spe-
cies from nearby live fence rows (less than 3 km
away from the trial site) and planted vegetatively
in rows at 1.5 m intervals. Species were chosen
based on their common use as live fence rows in
the area. Stakes were approximately 2 m tall at
planting. A pointed pole was inserted into the
ground to open a 15–20-cm-deep hole. The stake
was then inserted in the hole and soil was lightly
compacted around its base. All stakes were
planted in July (wet season) and were monitored
for three years for survival and above-ground
development.
Survival differed between species, ranging from
more than 90 percent to less than 30 percent;

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90
survival of some species was highly site-specific.
For most species, stakes with greater initial diam-
eter had a greater probability of survival. Species
varied enormously in above-ground biomass de-
velopment, and canopy cover ranged from less
than 2 m2 to more than 10 m2 in the third year.
Variability between sites was high. Not surprising-
ly, sites where survival and growth of stakes were
high were the same sites where establishment
rates for planted seedlings were high (Holl et al.,
2011). In comparing planting strategies between
the two studies, three-year-old Erythrina poeppi-
giana stakes had greater canopy cover than sap-
lings of the same age, although their height was
similar. Several species established from stakes
produced fruit in the second and third year after
planting. This is not surprising given that stakes
are pollarded from reproductive adult trees, con-
ferring an advantage over planting seedlings that
can take decades to produce fruit and attract
seed dispersers.
8.3. Direct seeding
Direct seeding is by far the cheapest way of re-
introducing vegetation (Lamb, Erskine and Par-
rotta, 2005; Cole et al., 2011), but tree seeds can
be hard to acquire and the rate of success is highly
variable. Seeds are typically harvested from trees
or the forest floor in nearby forests (Doust, Er-
skine and Lamb, 2006; Sampaio, Holl and Scariot,
2007; Cole et al., 2011), although in some cases
they can be purchased (Engel and Parrotta, 2001).
Seeds are either placed on the soil surface or bur-
ied. Many tropical forest tree seeds are recalci-
trant (i.e. they rapidly lose viability when dried),
making storage impossible, and the technique of
bulking up seed in the greenhouse or field plots,
which is commonly used for temperate tree spe-
cies, is not feasible for tropical forest trees.
As with the afore-mentioned propagation
methods, genetic variability of seed stock is often
low. Collecting tropical seed from a wide variety
of species can be difficult; many tropical forest
trees do not set seed every year and individuals
of a given species are often widely dispersed. As a
result, it is not uncommon to harvest seeds from
fewer than five mother trees and in some cases
only two or three (Doust, Erskine and Lamb, 2006;
Sampaio, Holl and Scariot, 2007; Garcia-Orth and
Martinez-Ramos, 2008; Cole et al., 2011). This can
have strong effects on germination and survival,
depending upon the quality of the seed source,
and could result in reduced genetic diversity in
future generations.
Direct seeding can be applied in two principal
ways: (1) at the onset of the recovery process of
the site; and (2) at later stages in the recovery pro-
cess, typically after a canopy cover has formed.
Direct seeding at the initial stages of recovery
has been tested in several small-scale experimen-
tal studies, but has not been considered a viable
restoration option at a large scale because of the
high rate of failure and the challenge of acquiring
and storing sufficient seed (Ray and Brown, 1995;
Engel and Parrotta, 2001; Woods and Elliott, 2004;
Doust, Erskine and Lamb, 2006; Sampaio, Holl and
Scariot, 2007). In most cases some establishment
occurs, but the variability across species and sites
is highly unpredictable. Seeds and recently ger-
minated seedlings typically succumb to a host of
setbacks, including pathogen attack, predation
and desiccation (Augspurger, 1984; Nepstad, Uhl
and Serrao, 1990; Chapman and Chapman, 1999;
Engel and Parrotta, 2001; Cole, 2009; Gallery,
Moore and Dalling, 2010; Cole et al., 2011). Small
seedlings can also be difficult to see among ruder-
al vegetation and may be removed during routine
vegetation clearing. Predators typically remove
a larger proportion of smaller seed than larger
seed, and larger-seeded species tend to have
greater establishment success because they have
larger amounts of stored resources (Camargo,
Ferraz and Imakawa, 2002; Jones, Peterson and
Haines, 2003; Doust, Erskine and Lamb, 2006;
Vieira and Scariot, 2006a). In turn, burial appears
to increase seed survival and germination com-
pared with surface placement (Woods and Elliott,
2004; Doust, Erskine and Lamb, 2006; Garcia-Orth
and Martinez-Ramos, 2008). Lastly, seasonal tim-
ing of planting can have a considerable effect on

GENET IC CONS IDERATI ONS IN ECO SYSTEM RE STORATI ON USING N ATIVE TRE E SPECIE S
91
long-term survival, especially in areas with a pro-
longed dry season (Ray and Brown, 1995; Vieira
and Scariot, 2006b).
Direct seeding seems to be most effective for
larger-seeded and later-successional species that
are introduced as part of enrichment planting
after the canopy has closed (Nepstad, Uhl and
Serrao, 1990; Hooper, Condit Legendre, 2002;
Cole et al., 2011). These species are often under-
represented in the initial stages of forest recovery
because of their short-range dispersal. Studies
comparing establishment at different succes-
sional stages indicate that, although germination
rates are similar among stages, long-term survival
is usually higher in sites that have tree canopy
cover (Bonilla-Moheno and Holl, 2010; Cole et al.,
2011). In contrast, Camargo, Ferraz and Imakawa
(2002) found higher survival of large-seeded spe-
cies in open highly degraded sites and in pastures
than in young and mature forest in lowland areas
in Brazil.
To date, we know of no large-scale tropical
forest restoration projects that introduced forest
trees through direct seeding. However, some au-
thors have suggested that seeding species with
relatively high germination and survival rates at
the early seedling stage should be a component
of a mixed restoration strategy that includes seed-
ing, planting seedlings and allowing for natural re-
generation of different species depending on their
life history (Cabin et al., 2002; Sampaio, Holl and
Scariot, 2007; Bonilla-Moheno and Holl, 2010). In
turn, larger-seeded, later-successional species may
be introduced in small patches in forests with an
overstorey, as introducing such species over large
areas is probably not feasible because of lack of
seeds.
Case study
In our study area in southern Costa Rica, we evalu-
ated the ability to establish from direct seeding of
six mid- to late-successional tree species in three
distinct habitats: recently abandoned pasture,
young plantation (approximately three years old)
as described earlier and young secondary forest
(approximately eight years old). The direct seed-
ing study was replicated across four research sites
(Cole et al., 2011). Species were sown to an aver-
age depth of 3 cm, and germination and survival
were monitored for two years. Germination rates
after two years ranged from near complete fail-
ure in one species to 26–31 percent for four spe-
cies and 94 percent in the sixth species. Overall
germination was similar among the three habi-
tats. However, survival was higher in plantations
(75 percent) than in the other two habitats (~45
percent). Plantations also had greater overall bio-
mass production at the end of the study, which
appeared to be due to higher nitrogen availabil-
ity as two of the four plantation trees were ni-
trogen-fixing species. Results indicate that direct
seeding of later-successional species into young,
recovering habitats with some degree of oversto-
rey cover can circumvent their dispersal limitation
and contribute to higher species diversity in the
forest.
8.4. Choosing an appropriate
restoration strategy
A first stage in any restoration project is to clearly
identify the goals. These goals and specific ob-
jectives will necessarily need to be developed
along with a consideration of the resources (e.g.
financial, labour, sources of seeds or seedlings)
available to achieve these goals and the natural
resilience of the target ecosystem (Holl and Aide,
2011). A goal of most tropical forest restoration
projects will be to restore the species composi-
tion and processes of the forest before it was dis-
turbed. However, given the competing needs of
providing for human livelihoods and maximizing
certain ecosystem services, there will be trade-offs
concerning which goals will be prioritized, such
as species diversity, carbon sequestration, erosion
control, or providing wood or food products used
by humans.
The degree of passive recovery of degraded
tropical lands is highly variable, depending on the
ecology of the system, land-use history and the
surrounding landscape mosaic (reviewed in Holl,

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2007). Therefore, a critical first step in a restora-
tion project is to determine which species will re-
sprout or colonize naturally and, therefore, may
not need to be introduced (Holl and Aide, 2011).
Second, it is a wise investment of resources to
conduct smaller field trials prior to planting large-
scale projects, as propagation methods and spe-
cies differ in their success rates from one location
to another. For projects that span large regions,
it is important to conduct pilot studies at multi-
ple sites given the high variability in success over
even relatively small spatial scales (Butterfield,
1996; Zahawi and Holl, 2009; Holl et al., 2011) as
a result of numerous factors, including land-use
history, soil physical and chemical properties, soil
microbial communities, competition with existing
vegetation and differences in microclimates. All
these recommendations take time and money to
implement, but in the long run will help to en-
sure the most efficient allocation of restoration
resources and will minimize the risk of large-scale
restoration failure.
Selecting an appropriate tree introduction
method requires knowledge of the natural his-
tory of the species available. While this informa-
tion is lacking for many species, the number of
studies screening germination rates (e.g. Sautu et
al., 2006), seedling survival rates (e.g. Butterfield,
1995) and even cuttings (Itoh et al., 2002) has in-
creased in the past two decades. Species that have
low seed germination rates, have complex germi-
nation triggers or produce small numbers of seeds
are not well suited to direct-seeding efforts, given
the large losses that typically occur as a result of
predation, herbivory and pathogens in the field.
Similarly, only certain species have known vegeta-
tive propagation abilities.
It is also important to consider how many spe-
cies will be introduced. Many tropical restoration
efforts plant a small number of tree species (of-
ten fewer than ten) to facilitate colonization and
establishment of a typically highly diverse native
flora and fauna, although in a few studies more
tree species (20–30) have been planted to repre-
sent a range of growth rates and dispersal guilds
(Lamb, 2011). It is much less common to plant
more than 30 species, given the necessary knowl-
edge for propagation and resources needed, al-
though some restoration efforts strive for diverse
plantings (e.g. 60–80 species; Rodrigues et al.,
2009) that include vines and shrubs. Finally, some
species may not be commercially available so it
is necessary to collect seed and then determine
whether it is more efficient to introduce each spe-
cies directly or establish them first as seedlings in
a nursery.
The issue of introducing sufficient genetic
variability into a restoration site is of concern in
all the propagation methods discussed above.
Forestry literature highlights the importance of
using diverse genetic sources, as well as selecting
for high-quality genotypes (reviewed in Carnus et
al., 2006; Kettle, 2010). While some restoration
projects harvest material from many individuals,
it is not uncommon, with all the propagation
techniques described above, to harvest stock
from only a few individuals, particularly when the
number of source trees is limited. Studies exam-
ining the potential implications of such narrow
selections (e.g. Carpenter et al., 1995; Dick et al.,
1998) present compelling results, with high vari-
ability in the establishment and growth of indi-
viduals from different genetic stocks.
In many cases, the cost of different propaga-
tion methods will be an overriding considera-
tion, given that most projects are financially con-
strained. For active restoration projects, direct
seeding represents the most economical route
and can be 20 to 30 times less expensive to carry
out than traditional nursery plantings (Engel and
Parrotta, 2001; Cole et al., 2011). Cuttings repre-
sent a similar cost-effectiveness to direct seeding
if they are directly planted out upon harvesting;
however, this is often not the case. When estab-
lished in nurseries, cuttings represent a similar
cost investment to establishing from seed; ac-
cordingly, this method should only be applied
to species that have demonstrated low seed fe-
cundity, or species that are rare or endangered.
The cost of using stakes is intermediate between
direct seeding and using cuttings established in
nurseries, with cost estimates ranging from two

GENET IC CONS IDERATI ONS IN ECO SYSTEM RE STORATI ON USING N ATIVE TRE E SPECIE S
93
to ten times cheaper than nursery stock (Zahawi
and Holl, 2009). Growing and planting seedlings
is usually the most expensive strategy but it is also
the most commonly used and the most widely
tested methodology.
Logistical considerations, such as challenges of
moving propagative material and the availability
of propagation facilities, also affect species selec-
tion. Seeds and direct-harvested cuttings are the
easiest propagules to transport. Stakes are not
only cumbersome but care must be taken when
transporting them so as not to damage the cor-
tex, which can impair their establishment ability
(Zahawi, personal observation). Accordingly, using
stakes is appropriate only when vegetative mate-
rial is available relatively close to a restoration
site or the need to use this method outweighs the
increased cost of transporting individuals of cer-
tain species. Seedlings are intermediate in terms
of ease of transport but require shade-house
facilities to propagate, which implies additional
costs. Clearly, the relative costs and logistical con-
siderations of different strategies will vary across
restoration projects, depending on availability of
seed and nursery facilities, transport distances for
vegetative propagules and other local conditions.
Each of the three strategies has its own advan-
tages and disadvantages, and it is likely that in
most cases a combination of the different propa-
gation methods is the best restoration approach.
In addition, site-specific conditions, the surround-
ing landscape and other factors specific unique to
a given restoration area will necessarily dictate
the most appropriate strategy.
Acknowledgements
The authors would like to thank J.L. Reid and
D. Douterlungne for helpful comments on earlier
versions of this manuscript.
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Inga: utilization. Kew, UK, Royal Botanic Gardens.
Perino, J.M. 1979. Rehabilitation of a denuded water-
shed through the introduction of kakawate (Gliricidia
sepium). Sylvatrop, 4: 49–68.
Ratnamhin, A., Elliott, S. & Wangpakapattanawong,
P. 2011. Vegetative propagation of rare tree spe-
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306–310.
Ray, G.J. & Brown, B.J. 1995. Restoring Caribbean dry
forests – evaluation of tree propagation techniques.
Restor. Ecol., 3: 86–94.
Rodrigues, R.R., Lima, R.A.F., Gandolfi, S. & Nave, A.G.
2009. On the restoration of high diversity forests: 30
years of experience in the Brazilian Atlantic Forest.
Biol. Conserv., 142: 1242–1251.
Romero, J.L. 2004. A review of propagation programs for
Gmelina arborea. New Forest., 28: 245–254.
Sampaio, A.B., Holl, K.D. & Scariot, A. 2007. Does res-
toration enhance regeneration of seasonal deciduous
forests in pastures in central Brazil? Restor. Ecol., 15:
462–471.
Sauer, J.D. 1979. Living fences in Costa Rican agriculture.
Turrialba, 29: 255–261.
Sautu, A., Baskin, J.M., Baskin, C.C. & Condit, R. 2006.
Studies on the seed biology of 100 native species
of trees in a seasonal moist tropical forest, Panama,
Central America. Forest Ecol. Manag., 234: 245–263.
Vieira, D.L.M. & Scariot, A. 2006a. Effects of logging,
liana tangles and pasture on seed fate of dry forest
tree species in Central Brazil. Forest Ecol. Manag.,
230: 197–205.
Vieira, D.L.M. & Scariot, A. 2006b. Principles of natural
regeneration of tropical dry forests for restoration.
Restor. Ecol., 14: 11–20.
Woods, K. & Elliott, S. 2004. Direct seeding for forest res-
toration on abandoned agricultural land in northern
Thailand. J. Trop. Forest Sci., 16: 248–259.
Zahawi, R.A. 2005. Establishment and growth of living
fence species: an overlooked tool for the restoration
of degraded areas in the tropics. Restor. Ecol., 13:
92–102.
Zahawi, R.A. 2008. Instant trees: using giant vegeta-
tive stakes in tropical forest restoration. Forest Ecol.
Manag., 255: 3013–3016.
Zahawi, R.A. & Holl, K.D. 2009. Comparing the perfor-
mance of tree stakes and seedlings to restore aban-
doned tropical pastures. Restor. Ecol., 17: 854–864.
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Testing applied nucleation as a strategy to facilitate
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10.1111/1365-2664.12014.

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The role that restoration plays in species conser-
vation is increasingly recognized in global forums.
For example, the recent Conference of the Par-
ties to the Convention on Biological Diversity
(COP-10) highlighted ecological restoration as
a significant opportunity for achieving global
conservation goals (CBD, 2010a). But some of
the fundamental challenges to achieving global
restoration targets, such as those set out in the
Global Strategy for Plant Conservation 2011–2020
(CBD, 2010b), are in need of broader recognition.
Contemporary restoration programmes aim to
restore biodiverse plant communities. In prac-
tice this means the return of tens to hundreds of
species in many ecosystems. Large-scale plant re-
introductions (hundreds to tens of thousands of
hectares) must be underpinned by the effective
use of seeds of wild species. This in turn requires
sufficient biological and technical knowledge of a
large number of species to enable the collection,
storage and germination of seeds and establish-
ment of seedlings.
9.1. Landscape-scale restoration
requires large quantities
of seed
Options for the active return of plant species to
degraded sites include direct seeding, planting
of seedlings and the spreading of appropriately
managed topsoil containing seeds (Koch, 2007;
Rokich and Dixon, 2007). Each of these meth-
ods can be used exclusively or in combination,
depending on the size of the restoration pro-
gramme, the available physical and biological
resources and the biological characteristics of the
available plant material (e.g. the seed-storage
characteristics). For all three options, seeds are
fundamental, being spread to site through their
incorporation into returned topsoil, broadcast
by hand, machine planted (e.g. drill seeding or
aerial seeding) or sown in a nursery for seedling
production. Properly handled topsoil can be very
effective at restoring plant communities (Koch,
2007; Rokich and Dixon, 2007). However, at most
restoration sites seed-containing topsoil is limited
or unavailable. For restoration at the landscape-
scale, direct seeding is often the most viable
means of initiating the return of biodiverse plant
communities (Merritt and Dixon, 2011).
A reliable supply of seeds is critical to success-
ful restoration. What is not always recognized
are the constraints surrounding the quantity of
seed required to achieve restoration goals and its
availability (Merritt and Dixon, 2011). Insufficient,
inconsistent and uncoordinated seed supply can
be a significant limiting factor in restoration pro-
grammes. Even at the local or regional scale, fac-
tors such as the availability of seeds, the technical
knowledge, training and licensing of the seed col-
lectors, the cost of seeds, and the biological and
technical knowledge necessary to correctly pro-
cess, store, break dormancy and deliver seeds to
restoration sites contribute to seed-supply short-
falls. At the landscape scale, these factors can be
greatly compounded by the very large quantities
of seeds needed for restoration.
Many restoration programmes are planned or
underway across the globe, aimed at restoring
Chapter 9
Seed availability for restoration
David J. Merritt and Kingsley W. Dixon
Kings Park and Botanic Garden, West Perth, Australia

THE STATE OF THE WORLD’S FOREST GENETIC RESOURCES – THEMATIC STUDY
PART 2
98
thousands or even tens of thousands of hectares,
often in poorly studied ecosystems with little avail-
able information on seed attributes or restoration
technology. With current restoration technologies
the amount of seed required for such programmes
can be calculated to be in the hundreds of tonnes,
far exceeding the seed-collecting capacities of
government agencies, non-governmental organi-
zations (NGOs) and commercial operations, as well
as the available seed resource that can be practi-
cally and ethically collected from wild plant popu-
lations. Seed availability is thus one of the most
significant challenges to large-scale restoration
programmes (Broadhurst et al., 2008; Rodrigues,
Lima et al., 2009; Gibson-Roy et al., 2010; Merritt
and Dixon, 2011; Tischew et al., 2011).
There are many examples of the quantities
and costs of seeds required for landscape-scale
restoration. In the agricultural zone of south-
west Western Australia, a 14 million ha agri-
cultural zone within a Mediterranean-climate,
biodiversity hotspot, over 93 percent of the
landscape has been cleared of native vegeta-
tion over the past 60 years, resulting in numer-
ous sustainability and productivity problems, in-
cluding dryland salinity, soil erosion and weed
invasion (Prober and Smith, 2009). To combat
landscape salinization an estimated 20-70 per-
cent of the landscape would need to be returned
to deep-rooted, woody perennial vegetation
(Prober and Smith, 2009). Using a conservative
seeding rate of just 1.5kg/ha (Jonson, 2010), at-
tempting to restore plant communities to just
20 percent of this landscape would require 4200
tonnes of seeds. In tropical forests in Borneo,
restoration projects plant between 500 and
2500 seedlings/ha (Kettle et al., 2011). Even at
a planting density of just 500 seedlings/ha, over
7 billion seedlings would be required to restore
the estimated 14.3 million ha of degraded for-
est (Kettle et al., 2011). In the United States, the
Bureau of Land Management (BLM) purchased
125 tonnes of seed of forb species in one year
for the Great Basin Restoration Initiative (Shaw
et al., 2005) and in 2007 the BLM spent US$50
million on seeding grass species in the Great
Basin, Mojave and Sonoran Deserts (Knutson et
al., 2009). On a similar scale, the cost of seed
purchases for restoration of 20000 ha of land
disturbed through mining activity in the semi-
arid Pilbara grasslands of Western Australia has
been estimated to exceed AUS$100 million at
current prices for wild-collected seeds (Merritt
and Dixon, 2011).
9.2. Seeding rates necessary
to delivery restoration
outcomes
The quantity of seed required to ensure an ac-
ceptable level of seedling establishment can vary
substantially across different biomes. Ideally,
seeding rates are based on known parameters
and data, including seed size, viability, germina-
tion and establishment rate (Gibson-Roy et al.,
2010). These parameters of seed quality are cap-
tured in the concept of “pure live seed” (a meas-
ure of the purity, viability and germination capac-
ity of a seed batch), an accreditation tool used for
evaluating seeds produced via commercial farm-
ing of wild species in the United States and some
parts of Europe (Jones and Young, 2005). If infor-
mation on seed quality is not gathered prior to
seeding, it is not possible to determine the success
(or otherwise) of direct seeding through monitor-
ing and documentation of seedling emergence to
determine the proportion of seeds that emerge.
In restoration practice, seed-quality analysis
prior to seeding, and monitoring of the results
following seeding, is often not done. Commonly
there is little published information available to
guide setting of seeding rates for local projects,
or criteria to evaluate success, reducing the incen-
tive for practitioners to strive for improvements
in seed-use efficiency. Many studies of direct
seeding are done on a very small scale (e.g. a
few square metres) to investigate the effects of
seed addition and/or seed treatments on seedling
emergence and establishment. These studies do
not always report seeding rates on a weight/area
basis (e.g. kg/ha), but rather the addition of a de-

GENET IC CONS IDERATI ONS IN ECO SYSTEM RE STORATI ON USING N ATIVE TRE E SPECIE S
99
fined number of seeds into a small plot or simply
employ an unknown number of seeds. However,
some examples of seeding rates for different bi-
omes are available that can be used to substanti-
ate the amount of seed required for restoration
(Table 9.1).
9.3. Constraints to seed
supply for landscape-scale
restoration
In most restoration projects the majority of seeds
are collected from wild plant populations. This
presents some challenges, given that many wild
populations, particularly those surrounding ag-
ricultural, pastoral and urban lands, are highly
fragmented, often degraded and under stress.
The amount of seed available to collect from wild
sources can fluctuate significantly from year to
year because of such factors as the environmental
conditions experienced by the maternal plants,
pollen flow, a requirement for disturbances (e.g.
fire) that promote mass flowering and fruiting
of some species and species biology (Jones and
Young, 2005). Relying solely on seeds collected
from the wild will increasingly result in supply
shortfalls as the demand for seeds increases to
match the scale of restoration.
The timing of seed collection is critical to a
successful outcome as there is often a window
of only a few days or weeks between when
seeds are ready to collect and when they are
dispersed and no longer available for collec-
tion. It is important to consider the phenology
of seed development, particularly the timing of
seed maturation, to ensure that the collected
seeds are suitably resilient to post-harvest han-
dling. Seeds should be collected as near as pos-
sible to the point of natural dispersal to ensure
that quality, desiccation tolerance (for ortho-
dox seeds) and longevity are maximized (Hay
and Smith, 2003). Kodym, Turner and Delpratt
(2010) demonstrated, for example, that several
species of Lepidosperma (Cyperaceae), which
contribute important understorey components
of temperate Australian woodland, shed vi-
able seeds quickly but retain non-viable seeds
for some months. Further, in these species vi-
able and non-viable seeds look similar, meaning
that incorrect timing of collection (i.e. too late)
could result in only non-viable seeds being col-
lected, while giving the impression that viable
seeds are available (Kodym, Turner and Delpratt,
2010). The importance of collection timing has
also been recently highlighted for tropical spe-
cies. The reproductive ecology of important
trees species, including dipterocarps, presents
TABLE 9.1.
Examples of seeding rates used in restoration programmes in different biomes
Region Biome Seeding rate
(kg seed/ha) Source
Australia Mediterranean woodland 1.5 Jonson (2010)
Australia Arid grassland 5–7 Merritt and Dixon (2011)
Australia Temperate grassland 50–110 Gibson-Roy
et al.
(2010)
Germany Semi-natural grassland 20–100 Baasch, Kirmer and Tischew (2012);
Kirmer, Baasch and Tischew (2012)
Northwestern Europe Ex-arable grassland 10–100 Kiehl
et al.
(2010)
Northwestern Europe Grassland 20–40 Török
et al.
(2011)
United Kingdom Calcareous grassland 1–40 Stevenson, Bullock and Ward (1995)
United States Continental sagebrush 2–8 Williams
et al.
(2002)

THE STATE OF THE WORLD’S FOREST GENETIC RESOURCES – THEMATIC STUDY
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particular challenges for large-scale seed supply
(Kettle, 2010). Dipterocarp seed production is
sporadic and unpredictable, with mass flowering
and fruiting events (seed masting) being a com-
mon but infrequent occurrence (Kettle, 2010;
Kettle, 2011). The window for seed collection
is short, usually a few weeks, and seed-masting
events are separated by years of low seed pro-
duction (Kettle, 2011). Many tropical forest spe-
cies produce recalcitrant seeds (Sacande, 2004),
including many of the species important for
timber. Recalcitrant seeds do not survive desic-
cation and cannot be stored for more than a
few weeks or months (Berjak and Pammenter,
2008). Storage behaviour of recalcitrant seed
means that it is not possible to take advantage
of seed-masting events through the collection
and storage of seeds for use in years of low pro-
duction. Recalcitrant seeds must be germinated
immediately and the seedlings held in a nursery
for planting into restoration sites (Kettle, 2010;
Kettle, 2011).
A need to source local provenance seeds for
restoration can also create challenges. Seed of
local provenance is best defined as seed that is
genetically representative of a species growing
within a particular climate, habitat, soil type
and profile in the landscape. Seed provenance is
important to restoration as local genotypes are
assumed to be better adapted to local environ-
mental conditions and, therefore, more likely
to establish (Krauss and Koch, 2004; McKay et
al., 2005; Bischoff, Steinger and Müller-Schärer,
2010; Jonson, 2010; Mijnsbrugge, Bischoff and
Smith, 2010). Sourcing seeds of local provenance
can be problematic, particularly in highly frag-
mented landscapes where small, remnant patch-
es of vegetation are separated by large areas of
land cleared for agriculture, infrastructure and
residential development. In these localized areas
the demand for seeds can easily exceed the sup-
ply and there may be some risks of detrimental
effects on the viability of the source vegetation
caused by overharvesting of seeds (Broadhurst et
al., 2008).
9.4. Approaches to improving
seed availability for
restoration
Developing appropriate seed-banking
procedures
Seed banking is a crucial link in the restoration
chain. Correct handling and storage allows ortho-
dox seeds to be banked over many seasons and
allows practitioners to capitalize on high-seeding
years, providing a resource for large restoration
projects. Careful control of the storage environ-
ment will ensure that seed viability is maintained.
Flexibility in the available storage conditions is
preferable, and seeds should be stored under con-
ditions appropriate to their storage behaviour,
dormancy type and designated storage duration
(Merritt and Dixon, 2011). Recognizing that all
seeds go through a storage phase prior to use in
restoration and putting in place the intellectual
and infrastructural capital required to curate the
seeds appropriately will ensure that the quality
of the seed resource is maintained. At present
seeds for use in restoration are stored almost
exclusively by end users, including the commer-
cial seed industry, mining companies, NGOs and
community-based groups. As a result, storage
facilities holding seeds for restoration are com-
monly low on technology, have limited access to
knowledge and training in modern seed science,
have little or no capacity for problem solving or
research and, in the case of the commercial seed
merchant, are profit-driven, meaning only those
plant species that are profitable (i.e. those pro-
ducing seeds that are easily accessible, robust to
the storage conditions and more reliable at the
establishment phase) will be sought, traded and
employed in restoration. Inadequate resourcing
of restoration seed banks is a rapidly emerging
bottleneck hampering landscape-scale restora-
tion. Restoration seed banks must be developed
by adapting principles and technologies put in
place for seed banks conserving biodiversity and
food crops, with the crucial difference that the
volume of seed required to address biodiverse

GENET IC CONS IDERATI ONS IN ECO SYSTEM RE STORATI ON USING N ATIVE TRE E SPECIE S
101
landscape-scale restoration compels restoration
seed banks to store hundreds of tonnes of seeds
(Merritt and Dixon, 2011).
Improving seedling establishment
A major limitation to the effectiveness of direct
seeding is the poor conversion of seeds into es-
tablished seedlings (James, Svejcar and Rinella,
2011; Merritt and Dixon, 2011). Failed seedling
establishment is a significant contributing factor
to the huge quantities of seeds required for resto-
ration and the inability to re-establish functional
plant communities. Across a range of habitats,
commonly less than 10 percent (and often as low
as 3 percent) of seeds delivered to site germinate
and establish (Merritt and Dixon, 2011). In Medi-
terranean southwest Australia, emergence rates
of 1–17 percent have been reported for a range
of Banksia woodland native species (Turner et al.,
2006; Rokich and Dixon, 2007). Similarly, in the
arid grasslands of the United States, 7–17 per-
cent establishment of germinated seeds was re-
corded for three grasses, and modelling of seed
fates across four restoration sites calculated the
probability of a seed producing an established
seedling to be less than 0.06 (James, Svejcar and
Rinella, 2011). Low seedling establishment is also
reported for tropical forests. A restoration trial
using three mature forest species to seed land
previously used for slash-and-burn agriculture in
Mexico’s Yucatan Peninsula found on average that
5–41 percent of seeds germinated and emerged,
and that 3–35 percent of these seedlings estab-
lished (Bonilla-Moheno and Holl, 2010). In central
Amazonia, seedling emergence of 12–33 percent
has been reported across 11 native tree species
seeded into abandoned pasture lands (Camargo,
Ferraz and Imakawa, 2002). Seed losses accrue not
just through failed germination and establish-
ment, but also through wind and water erosion
and predation (Holl et al., 2000; Doust, 2011).
Research and technological development is
needed to reduce the wastage of seeds during
delivery and establishment. Seed-enhancement
treatments must be explored to increase seed
germination performance and seedling estab-
lishment. Seed-enhancement treatments include
priming, coating and pelleting. Much of this tech-
nology is routinely applied through the agricultur-
al and horticultural biotechnology industries, but
as yet has not been widely adopted in the native
seed industry. However, priming has been demon-
strated to increase seedling emergence of native
grass species under field conditions (Hardegree
and Van Vactor, 2000), and simple techniques of
on-farm seed priming are used for cereals and
legumes to improve crop establishment (Harris et
al., 1999). Seed pelleting has been demonstrated
to increase seedling emergence of Banksia wood-
land species in southwest Western Australia, as
well as decreasing predation and losses through
wind erosion (Turner et al., 2006). Seedling estab-
lishment rates can also be improved by correctly
timing seed delivery to site and employing simple
treatments such as incorporation of seeds into the
soil (Turner et al., 2006).
Increasing seed supply
In variously termed seed orchards, seed farming
or seed-production areas, growing wild plant spe-
cies specifically to harvest their seeds for restora-
tion is receiving increasing attention as a part of
the solution to seed-supply shortfalls. Options for
seed-production areas include the setting aside
of wild populations of plants for dedicated seed
collection, the growing of plants in pots under
nursery conditions for annual harvesting of seeds
(Koch, 2007; Gibson-Roy et al., 2010) or the de-
velopment of purpose-designed broadacre seed
farms where plants are grown using agricultural
cultivation and harvesting techniques (Shaw et
al., 2005). Some common challenges to devel-
oping viable seed-production enterprises for a
wide range of species include a limited knowl-
edge of seed-propagation and plant-husbandry
requirements, and the need for rigorous seed
certification and quality-control procedures and
to effectively manage genetic considerations,
including the potential provenance variation
of source-plant material and the genetic conse-
quences of seed production (Gibson-Roy et al.,
2010; Tischew et al., 2011). Other issues relate

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102
to the inadvertent selection processes inevita-
bly introduced via source-plant seed collection,
maternal-plant growth and survival and harvest-
ing techniques (Mijnsbrugge, Bischoff and Smith,
2010). Nevertheless, programmes of research, de-
velopment and commercial supply through large-
scale, certified wild-seed production are in place
for large-scale restoration programmes such as
those under the Great Basin Restoration Initiative
of the United States11 (Shaw et al., 2005). On a
similarly large scale, the SALVERE Project,12 across
central Europe, includes research into the seed-
production potential of semi-natural grasslands
as a source of seeds for restoration. The Millen-
nium Seed Bank’s UK Native Seed Hub Project13
aims to establish seed production for lowland
meadows and semi-natural grassland across the
United Kingdom in partnership with the commer-
cial and restoration sectors. At a more regional
scale, the potential for NGOs and local communi-
ties to develop and manage seed production ar-
eas to increase the supply of understorey species
has been successfully demonstrated for the Grassy
Groundcover Restoration Project across south-
eastern Australia (Gibson-Roy et al., 2010). This
project produced 92 kg of seeds of approximately
200 native herbaceous species over two years for
the restoration of ex-agricultural land (Gibson-
Roy et al., 2010).
9.5. Conclusion
Seeds are fundamental to large-scale restoration,
being the only viable means of reintroducing
plants at the 100–1000 km2 scale. But obtaining
seeds of wild species is a significant challenge to
landscape-scale restoration. Key areas of seed
biology and technology underpin restoration,
and optimizing each step in the chain of seed us-
age in restoration, from collection to delivery to
11 http://www.blm.gov/id/st/en/prog/gbri/technology/native_
plants.html
12 http://www.salvereproject.eu
13 http://www.kew.org/news/uk-seed-hub.htm
site, is necessary. Knowledge of these key areas
is complex when dealing with biodiverse plant
communities and species-specific information.
Seed-enhancement techniques for each species
must be tailored to site-specific needs for ef-
fective restoration. This includes consideration
of abiotic factors such as the landform stability,
slope, aspect and the available growing medium
(soil conditions which are often heavily different
to those prior to disturbance) and hydrological
aspects, including the reliability and seasonality
of rainfall and soil-moisture retention properties.
The unification of science-based seed knowledge
with the infrastructure to support large-scale
seed management and the development of ef-
fective working relationships between seed sci-
entists, restoration practitioners, the commercial
seed industry and the local community will ensure
seeds are used to their full potential.
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establishment of semi-natural grasslands in Central
and Northwestern Europe. Basic Appl. Ecol., 11(4):
285–299.
Kirmer, A., Baasch, A. & Tischew, S. 2012. Sowing of
low and high diversity seed mixtures in ecological
restoration of surface mined-land. Appl. Veg. Sci.,
15(2): 198–207.
Knutson, K.C., Pyke, D.A., Wirth, T.A., Pilliod, D.S.,
Brooks, M.L. & Chambers, J.C. 2009. A chron-
osequence feasibility assessment of emergency
fire rehabilitation records within the Intermountain
Western United States – Final Report to the Joint
Fire Science Program. Project 08-S-08. US Geological
Survey Open-File Report 2009-1099 (available at:
http://pubs.usgs.gov/of/2009/1099/).
Koch, J.M. 2007. Alcoa’s mining and restoration pro-
cess in south western Australia. Restor. Ecol., 15:
S11–S16.

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Kodym, A., Turner, S. & Delpratt, J. 2010. In situ
seed development and in vitro regeneration of
three difficult-to-propagate Lepidosperma species
(Cyperaceae). Aust. J. Bot., 58(2): 107–114.
Krauss, S.L. & Koch, J.M. 2004. Rapid genetic delinea-
tion of provenance for plant community restoration.
J. Appl. Ecol., 41(6): 1162–1173.
McKay, J.K., Christian, C.E., Harrison, S. & Rice, K.J.
2005. “How local is local?” – a review of practical
and conceptual issues in the genetics of restoration.
Restor. Ecol., 13(3): 432–440.
Merritt, D.J. & Dixon, K.W. 2011. Restoration seed
banks – a matter of scale. Science, 332(6028):
424–425.
Mijnsbrugge, K., Bischoff, A. & Smith, B. 2010. A
question of origin: where and how to collect seed
for ecological restoration. Basic Appl. Ecol., 11(4):
300–311.
Prober, S.M. & Smith, F.P. 2009. Enhancing biodi-
versity persistence in intensively used agricultural
landscapes: a synthesis of 30 years of research in the
Western Australian wheatbelt. Agr. Ecosyst. Environ.,
132(3–4): 173–191.
Rodrigues, R.R., Lima, R.A.F, Gandolfi, S. & Nave,
A.G. 2009. On the restoration of high diversity for-
ests: 30 years of experience in the Brazilian Atlantic
Forest. Biol. Conserv., 142(6): 1242–1251.
Rokich, D.P. & Dixon, K.W. 2007. Recent advances in
restoration ecology, with a focus on the Banksia
woodland and the smoke germination tool. Aust. J.
Bot., 55(3): 375–389.
Sacande, M., Joker, D., Dulloo, M.E. & Thompsen,
K.A., eds. 2004. Comparative storage biology of
tropical tree seeds. International Plant Genetic
Resources Institute, Rome.
Shaw, N., Lambert, S.M., DeBolt, A.M. & Pellant, M.
2005. Increasing native forb seed supplies for the
Great Basin. In, R.K. Dumroese, L.E. Riley & T.D.
Landis, tech. coords. National proceedings: Forest
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Stevenson, M.J., Bullock, J.M. & Ward, L.K. 1995.
Re-creating semi-natural communities: effect of sow-
ing rate on establishment of calcareous grassland.
Restor. Ecol., 3(4): 279–289.
Tischew, S., Youtie, B., Shaw, N. & Kirmer, A. 2011.
Farming for restoration: building bridges for native
seeds. Ecol. Restor., 29(3): 219–222.
Török, P., Vida, E., Deák, B., Lengyel, S. &
Tóthmérész, B. 2011. Grassland restoration on
former croplands in Europe: an assessment of appli-
cability of techniques and costs. Biodivers. Conserv.,
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Turner, S.R., Pearce, B., Rokich, D.P., Dunn, R.R.,
Merritt, D.J., Majer, J.D. & Dixon, K.W. 2006.
Influence of polymer seed coatings, soil raking, and
time of sowing on seedling performance in post-
mining restoration. Restor. Ecol., 14(2): 267–277.
Williams, M.I., Schuman, G.E., Hild, A.L. & Vicklund,
L.E. 2002. Wyoming big sagebrush density: effects
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10(2): 385–391.

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A major constraint for afforestation and restora-
tion programmes around the world is the lack of
availability of large numbers of high-quality seeds
of indigenous species with suitable provenance
and accompanying data.
As part of an ongoing seed-longevity study,
Kew’s Millennium Seed Bank (MSB) recently iden-
tified a range of tree species listed on the World
Agroforestry Centre’s Tree Seed Supplier Directory
(TSSD) that were not present in the MSB’s collec-
tions and which would be suitable for the study.14
Seed availability
Kew targeted 30 of the largest public-sector and
commercial seed suppliers on the TSSD who be-
tween them should, according to the Directory,
have been able to supply 1624 species meeting
Kew’s requirements. However, of the 30 suppli-
ers listed, seven could not be contacted and one
was on the list twice. The remaining 22 were con-
tacted, but only seven responded, representing a
24 percent success rate for supplier responses.
Once contact had been made, Kew requested
a total of 633 species listed as available from
the seven suppliers. A minimum number of 2000
seeds were requested, and minimal accompany-
ing data on seed origin and storage conditions
were specified.
Eventually, six months after the process begun,
Kew was able to secure collections of 218 unique
species. This represents an overall seed-supply
success rate of 13 percent of the total number of
14 See http://www.worldagroforestrycentre.org/Sites-old/
TreeDBS/tssd/treessd.htm
species theoretically available and 34 percent of
the species advertised by the seven suppliers who
were successfully contacted.
Data quality
A subset of 572 species on the TSSD were checked
for current name status, and it was found that 48
percent of the names on the list are no longer
valid (i.e. they are synonyms). When the correct
names were compared with the MSB’s accession
list it was found that 25 percent of the collections
were already in the MSB.
The following provenance data accompanied
the collections received: wild/cultivated origin (48
percent of collections); date of collection (88 per-
cent); country of origin (100 percent); region of
origin (65 percent); precise locality (14 percent).
Seed quality
Seed-quality testing is currently taking place.
However, the collections were accompanied by
the following information on seed processing:
drying conditions (specified for only 15 percent of
collections); date of storage (89 percent); relative
humidity during storage (15 percent); and tem-
perature of storage (100 percent).
Conclusion
All of the above indicates the common difficulties
encountered in sourcing high-quality seed collec-
tions in reasonable numbers and with minimal
provenance data, even from reputable sources.
Insight 6
Seed availability: a case study
Paul P. Smith
Millennium Seed Bank, Royal Botanic Gardens, Kew, United Kingdom

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Seed banks have a major role to play in habitat
restoration, both as a source of material and in
solving research problems related to reintroduc-
ing species back into the landscape (Hardwick et
al., 2011; Smith et al., 2011).
Seed banks have an important advantage over
nurseries in that they can store a large amount of
genetic diversity in a very small space. For exam-
ple, a typical 30 m3 cold room in Kew’s Millennium
Seed Bank stores 20000 seed collections totalling
1 billion seeds. In addition, seeds kept under cool,
dry conditions are more secure than seedlings in a
nursery, the latter being more susceptible to pests
and diseases, extreme weather etc. From this per-
spective, it makes sense to store plant diversity
as seed right up to the time when it is needed.
Finally, from the restoration practitioner’s view-
point, direct seeding is far more cost-effective
than reintroducing seedlings or saplings (see
Section 8.4). However, for successful re-seeding,
research is required to optimize germination and
survival.
Seed-conservation research and expertise with
direct relevance to restoration programmes in-
cludes seed sampling, collection, handling and
developing appropriate storage methods (short,
medium and long term). Seed morphology can
also inform practitioners about natural disper-
sal mechanisms. However, perhaps the most im-
portant contribution that seed banks make is in
developing optimal germination protocols, tak-
ing into account the physical and physiological
dormancy mechanisms present in so many wild
species. Most seed banks routinely carry out ger-
mination testing to test for viability. However, for
wild species there are frequently challenges asso-
ciated with dormancy mechanisms that need to
be characterized, and appropriate pretreatments
or priming methodologies developed (Probert,
2000; Merritt et al., 2007).
Kew’s Millennium Seed Bank (MSB) is current-
ly the only global repository for wild species. It
stores seeds from 141 countries, and every col-
lection is tested for dormancy and germination.
Optimal germination protocols and information
on other traits, such as seed storage behaviour, are
freely available through Kew’s Seed Information
Database on line.15 The MSB’s Seed Information
Database currently contains information on more
than 11000 tree and shrub species.
For United Kingdom restoration practitioners,
Kew has gone a step further and produced a ger-
mination predictor tool that takes into account
where and when seeds are collected, and uses
this information to predict optimal germination
protocols.16 This approach takes variation in local
genotypes and climate into account.
Many national and regional seed banks fulfil
similar roles locally. Seed banks with a strong
restoration-ecology focus that provide both
material and methodologies include Kings Park
(Western Australia); Plant Bank (New South
15 http://data.kew.org/sid/
16 See http://www.kew.org/science-research-data/databases-
publications/uk-germination-tool-box/
Insight 7
The role of seed banks
in habitat restoration
Paul P. Smith
Millennium Seed Bank, Royal Botanic Gardens, Kew, United Kingdom

GENET IC CONS IDERATI ONS IN ECO SYSTEM RE STORATI ON USING N ATIVE TRE E SPECIE S
107
Wales, Australia); Chicago Botanic Garden’s Plant
Conservation Science Center (United States);
China’s Gene Bank of Wild Species in Kunming;
and Kirstenbosch Botanical Garden in South
Africa. In addition to these specialist institutes,
many forestry gene banks support afforestation
of native species. A recent survey of government
tree seed centres in 12 African countries (Kew,
unpublished), found that, collectively, these insti-
tutions supply 40 tonnes of seeds and 398 million
seedlings of 558 species each year. The major-
ity of seeds and seedlings supplied are of exotic
species. However, all of the forestry institutions
surveyed also supply indigenous tree seeds and
seedlings, albeit in smaller amounts than exotics.
In developed countries, capability related to the
propagation and use of indigenous species is far
more advanced. For example, each year Poland’s
State Forests supply 650 tonnes of seeds of native
tree species from 92 different seed zones, which
are used to produce around 850 million seed-
lings for introduction in to the landscape (Koziol,
2012).
In the private sector, the mining industry in par-
ticular is at the forefront of restoration efforts.
In large-scale restoration of complex habitats, a
combination of direct seeding and plug planting is
employed. For example, Alcoa’s Jarrah forest res-
toration programme in Western Australia (Koch
& Hobbs, 2007) and Rio Tinto’s Littoral forest res-
toration programme in Madagascar (Vincelette et
al., 2007) have established seed banks to support
restoration activities.
References
Hardwick, K.A., Fiedler, P., Lee, L.C., Pavlik, B., Hobbs,
R.J., Aronson, J., Bidartondo, M., Black, E., Coates,
D., Daws, M.I., Dixon, K., Elliott, S., Ewing, K.,
Gann, G., Gibbons, D., Gratzfeld, J., Hamilton,
M., Hardman, D., Harris, J., Holmes, P.M., Jones,
M., Mabberley, D., Mackenzie, A., Magdalena, C.,
Marrs, R., Milliken, W., Mills, A., Lughadha, E.N.,
Ramsay, M., Smith, P., Taylor, N., Trivedi, C., Way,
M., Whaley, O. & Hopper, S.D. 2011. The role of bo-
tanic gardens in the science and practice of ecological
restoration. Conserv. Biol., 25: 265–275.
Koch, J.M. & Hobbs, R.J. 2007. Synthesis: Is Alcoa
successfully restoring a Jarrah forest ecosystem after
bauxite mining in Western Australia? Restor. Ecol.,
15(Supplement S4): 137–144.
Koziol, C. 2012. Collection of seeds of forest trees,
shrubs and herbaceous plants – a comparison of
accepted standards. Presentation to the workshop
on Current technologies of forest seed treatment.
21–25 May 2012, Kostrzyca Forest Gene Bank,
Milkow, Poland.
Merritt, D.J., Turner, S.R., Clarke, S. & Dixon, K.W.
2007. Seed dormancy and germination stimulation
syndromes for Australian temperate species. Aust. J.
Bot., 55(3): 336–344. doi: 10.1071/BT06106
Probert, R.J. 2000. The role of temperature in the regula-
tion of seed dormancy and germination. In M. Fenner,
ed. Seeds: the ecology of regeneration in plant com-
munities, 2nd ed. Wallingford, UK, CABI Publishing.
Smith, P.P., Dickie, J., Linington, S., Propert, R. &
Way, M. 2011. Making the case for plant diversity.
Seed Sci. Res., 21: 1–4.
Vincelette, M., Rabenantoandro, J., Randrihasipara,
L., Randriatafika, F. & Ganzhorn, J.U. 2007.
Results from ten years of restoration experiments
in the southeastern littoral forests of Madagascar.
In J.U. Ganzhorn, S.M. Goodman & M. Vincelette,
eds. Biodiversity, ecology and conservation of littoral
ecosystems in southeastern Madagascar, Tolagnaro
(Fort Dauphin), pp. 337–354. SI/MAB Series #11.
Washington, DC, Smithsonian Institution.

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This chapter presents a broad-based overview of
how traditional ecological knowledge (TEK) and
traditional resource management (TRM) can in-
form ecological restoration and sustainable for-
est management, based on experiences from the
temperate forests of far western North America.
We do not generally associate forest restoration
with forest timber management; rather, we think of
restoring degraded wild ecosystems to some sem-
blance of their former healthy state. In this chapter
it is argued that silviculture, a form of agriculture,
can be enhanced by restoration and, reciprocally,
appropriate silviculture can assist in restoration and
maintenance of forest. The chapter concludes with
the presentation of some insights into the genetic
implications of silviculture, restoration and indig-
enous TRM and genetic relationships that can be
affected either negatively or positively by how we
manage both restoration and silviculture.
The relationships between restoration, silvicul-
ture and indigenous TRM are not well understood.
While Western managers and ecologists frequent-
ly express interest in local examples of TEK, e.g.
plant or animal indicators that could assist them
in their research, it will be necessary here to take
a more universal approach. This paper adopts
this broad, holistic perspective in order to clarify
these relationships and to bolster the argument
for the importance of TEK and TRM to restoration
and silviculture. To this end it is necessary to first
describe indigenous TEK/TRM and the key, nearly
universal, cultural practice of prescribed burning.
The extent and ecological importance of indige-
nous burning is still controversial, but it is founda-
tional to the argument for the use of an historical
indigenous-managed forest model with which to
guide restoration and enhance silviculture.
Indigenous cultural land-care practices or
TRM, in concert with natural processes, created
and maintained distinct cultural landscapes that
could be described as a kind of indigenous agro-
ecology or agriculture. These systems, which in-
cluded modifying vegetation by fire, were em-
ployed over millennia to enhance ecosystems in
order to produce food, medicine, cordage, bas-
ketry, cages and traps, ceremonial items, clothing,
games, musical instruments, tools and utensils,
weapons, fishing and hunting gear and structures
(Anderson, 2005). The forest was (and still is for
many indigenous peoples) the local supermarket,
pharmacy and hardware store. Indigenous agro-
ecology, like Western agriculture, influences the
local availability, abundance, composition and
distribution of plants (and, in the case of agro-
Chapter 10
Traditional ecological knowledge,
traditional resource management
and silviculture in ecocultural
restoration of temperate forests
Dennis Martinez
Chair, Indigenous Peoples’ Restoration Network (IPRN) of the Society for Ecological Restoration
International (SERI), Member, Indigenous Peoples’ Biocultural Climate Change Assessment
Initiative (IPCCA) Steering Committee, Co-Director, Takelma Intertribal Project (TIP)

THE STATE OF THE WORLD’S FOREST GENETIC RESOURCES – THEMATIC STUDY
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ecology, animals). It is roughly equivalent to
Western agriculture without the need for plough-
ing, fertilizing or irrigation, and without ecologi-
cally harmful side effects such as excessive nitrifi-
cation and dependence on fossil-fuel inputs. It is
not merely a kind of “proto-agriculture” repre-
senting a late phase in the evolution of what we
conventionally understand as “true agriculture.”
It had proven its worth as the most adapted kind
of management for the environments in which it
evolved. It is questionable whether it would have
had any need to evolve further.
TEK is a belief, knowledge and practice complex
(Berkes, 2008) passed orally from generation to
generation and informed by strong cultural mem-
ories and sensitivity to change. It encompasses a
wide variety of ecological knowledge, including
animal behaviour and social ecology, indicator
species, weather prediction, fire behaviour and
prescribed burning, gathering, fishing and hunt-
ing knowledge, relationships between insects,
birds, plants and animals, agroforestry, agro-
ecology, horticulture and memories of significant
weather and other ecological events. Much of this
knowledge is encoded in indigenous languages;
when a language is lost, so is valuable ecologi-
cal knowledge. Community knowledge specialists
guide and regulate resource use, while families
and clans exercise ownership management and
conservation responsibilities for their particular
places, thus avoiding the tragedy of the commons.
Reciprocity, sharing and restraint are informed and
maintained by a spiritual belief system with dire
consequences (shame and misfortune) for those
who are greedy and disrespectful toward the ani-
mals and plants that they regard as relatives in a
kincentric world. Kinship is the glue that holds it
all together. Traditional indigenous societies are
conservative to their core and highly risk averse.
To paraphrase what the International Indigenous
Commission (IIC) told the delegates at the 1992
Rio Earth Summit, indigenous production meth-
ods involve increasing biodiversity by constantly
creating new diverse habitats or niches – most
often with intentional use of fire that maintained
a fine-grained, patchy landscape mosaic – while
maintaining surplus biodiversity or overcapacity as
an untapped capital reserve. This brief summary of
the context in which TEK is rooted provides a good
segue to the main focus in this paper: TRM as just
one of many possible components of TEK.
This chapter focuses on cultural land-care prac-
tices that contributed historically to a particu-
lar forest structure and composition (Society for
Ecological Restoration International Science &
Policy Working Group, 2004) and how this unique
forest structure maintained by indigenous peoples
can inform the spatial arrangement of subsistence
and commercial timber and non-timber species. For
example, modified historical indigenous models
can be applied to even plantation forestry, modify-
ing the spatial structure, making it more diverse,
while sequestering carbon or providing timber and
non-timber products. Special attention will be paid
to forest genetics while integrating modified for-
est structure with native composition, i.e. how to
reconnect commercial and/or subsistence forests
with ecosystem function and resiliency in a time of
rapid environmental change.
What we conventionally call “novel,” “natu-
ral” or “pristine” landscapes are often, in part,
degraded cultural or agro-ecological landscapes
(some indigenous peoples call these landscapes
their “garden”). Here is where the line between
ecological restoration and restoration of cultural
landscapes becomes blurred, requiring a different
restoration term – “ecocultural” or “biocultural”
restoration.
Ecocultural restoration is the process of recov-
ering as much as possible of the key ecosystem
structure, composition, processes and function
that existed prior to European contact, along with
traditional, time-tested, ecologically appropriate
and sustainable indigenous cultural practices that
helped shape ecosystems and cultural landscapes
(Keenleyside et al., 2012). This is done while si-
multaneously building in resilience to future rap-
id climate disruptions and other environmental
changes (Box 10.1) in order to maintain ecologi-
cal integrity in a way that ensures the survival of
both indigenous ecosystems and cultures, includ-
ing culturally preferred species – a distinguishing

GENET IC CONS IDERATI ONS IN ECO SYSTEM RE STORATI ON USING N ATIVE TRE E SPECIE S
111
feature of ecocultural as opposed to ecological
restoration (Martinez, 2013). It should, however,
be noted that mostly non-cultural plant commu-
nities in which the cultural plants occur are also
valued as relatives deserving of protection, resto-
ration or both. Indigenous TEK and TRM and the
resulting historical forest structure and composi-
tion managed by indigenous peoples can inform
ecocultural restoration by supplying an initial
reference model or baseline and can provide a
way to bridge TEK/TRM and silviculture (Egan and
Howell, 2005).
Ecological/ecocultural restoration is not, as
is commonly believed, going back to some pre-
industrial snapshot in time. Nor does it mean
continuing with just the present degraded forest
(also a snapshot in time). Rather than turning the
clock back, we are resetting the evolutionary clock
and attempting to restart a trajectory bounded
by conceptually reconstructed historical ranges
of variability in the types, intensities, extents and
frequencies of natural disturbances or stressors,
with which the forest ecosystem is genetically
and ecologically familiar (Perry, 1994). When one
considers the length of time indigenous peoples
have been on the American continent (estimates
have been consistently rising over the past cen-
tury from a few thousand years to 30000 years or
more [Dobyns, 1966; Fiedel, 2000; Mann, 2005]),
indigenous stewardship surely has affected forest
genetics through selective harvesting and the use
of fire that influenced cultural plant and animal
abundance, characteristics and distribution (~300
plant species were typically utilized as well as
many non-useful species affected by larger hunt-
ing fires).Therefore the term “natural” should in-
clude indigenous caregivers as a keystone biotic
component of ecosystem dynamics. We hope, in
ecocultural restoration, to at least be able to cap-
ture key features of disturbance regimes, struc-
ture, composition, processes and function togeth-
er with longstanding cultural land-care practices
and important cultural species (Box 10.1).
The reconstructed reference model is only a
guide, but one that is anchored in real ecocul-
tural and historical time. In the process of setting
restoration goals, we will have to balance histori-
cal fidelity to the reference model with ecologi-
cal functionality, resilience and integrity given
changed environmental conditions (Higgs, 2003).
But the model will assist in giving us a sense of di-
rection by restoring an evolutionary trajectory that
has been seriously derailed. This historical baseline
is important for gauging environmental change
and the degree of degradation. It is what we are
striving to restore – even if we are not entirely suc-
cessful or the work completed. Indeed, restoration
will probably always require some periodic human
intervention, such as controlled burning.
Contact with Europeans and their diseases killed
up to 90 percent of the indigenous population in
many places and created the common misconcep-
tion of historically low populations. Indigenous
populations were relatively large before contact
with Europeans17 and required prodigious amounts
of material from plants that were burned the pre-
vious year, e.g. fire-induced epicormic and adven-
titious shrub or tree sprouts used in basketry, or
the burning of sometimes hundreds to thousands
of hectares to rejuvenate brush species (Ceanothus
spp., oak, plum, hazel, mountain mahogany etc.)
palatable to black-tailed deer (Odocoileus he-
mionus columbianus) and elk (Cervus canadensis)
(Lewis, 1973; Boyd, 1999; Stewart, 2002; Blackburn
and Anderson, 1993).18
17 Henry F. Dobyns, cited by Mann (2005), estimated the popula-
tion of the Americas in 1491 at 90–112 million, compared with an
earlier estimate by Mooney (1928) of 1.2 million for North America.
18 To give the reader an idea of the amount of burned plant ma-
terial required, consider the following: in California, 35000 stalks
of milkweed (Asclepias sp.) or Indian hemp (Apocynum cannabi-
num) were required for one deer net about 15 m long (Blackburn
and Anderson 1993) and 1200 sprouts of sourberry (Rhus trilobata)
were needed for a burden basket. Twenty-five basket weavers in a
typical California village of about 100 people might harvest about
250000 shoots in a single season. Lightening could not be relied
on to start the necessary fires because it strikes at random (i.e. it
could not be relied on to strike where it was needed on a regular
basis and was relatively rare in lower elevations and coastal areas)
and because fire started by lightening was often different from fires
started deliberately in terms of spatial selectivity, extent, frequency,
intensity and seasonality. Sixty percent of cultural items came from
plant material (Anderson 2005; Chester King in Blackburn and
Anderson 1993).

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Cultural landscapes in far western North
America were created and maintained by peri-
odic burning by indigenous peoples. This kept
forest succession in an arrested state, producing
a fine-grained, patchy vegetation mosaic (Lewis,
1973; Anderson, 2005). Fire had many ecological
and human benefits, including nutrient cycling,
better access to hunted animals, less groundwa-
ter lost through evapotranspiration, pest control,
stimulation of plant regrowth, improved wildlife
habitat, increased seed germination and seedling
survival and reduction of hazardous fuels. In the
wetter regions of the coastal Pacific northwest of
the United States and coastal western Canada,
patch burning for berries, habitat or baskets
(Turner, 2010), among other reasons, had less ef-
fect on forest succession. Decomposer arthropods
and fungi cycled nutrients, while forest gaps were
Indigenous peoples are rarely in a position to
assist migration of climate-vulnerable species
because their territories are mostly relatively
small and peoples are rooted in place, without the
opportunity to follow displaced species except to
higher elevations in some places. However, these
places are relatively small in extent. Some cooler
micro-sites in mountainous regions do occur at
lower elevations. Individuals from thermally stressed
species, such as the endangered keystone tree species,
whitebark pine (
Pinus albicaulis
), have become well
established on these sites (C. Millar., 2012, personal
communication). Important cultural plants will have
to be maintained on the reservation, rancheria or
reserve, reinforcing the importance of historical
reference models for indigenous peoples. Cultural
“resistance” will be necessary to maintain cultural
integrity through the adaptive suitability of the
germplasm of cultural species. The indigenous oral
tradition suggests that indigenous people have had
to adapt to environmental change many times before.
For example, oral tradition tells us that indigenous
peoples moved salmon spawn in wet moss when
rivers were blocked in Pacific Northwest North
America (Sproat 1868; Campbell and Butler, 2010;
Coast Salish Nuuchalnulth oral tradition). The last
time this happened was in 1913 at Hells Gate, when
a massive landslide blocked the Fraser River in British
Columbia, Canada, and the Salish people built a flue
around the slide to save sockeye salmon returning
to spawn. Eventually, some of these cultural species
may be displaced, but it is a question of buying time
to allow alternative species to become available from
adequate planning and not by default at the last hour.
Ecosystem-based adaptation is critically important;
global warming and climate weirdness are already
having an impact on indigenous peoples and the
vulnerable ecosystems they inhabit. The challenge
is to find sufficient genetic diversity in culturally
important species. We will have to seek out adapted
plants – isolated individuals, populations and
subspecies – in addition to the cross-breeding
currently being done. Forested landscapes with
considerable heterogeneity may provide a number of
possible micro-sites that could enhance forest refugial
capacity. A good course of action would be to collect
propagules from populations in extreme micro-sites
(exposed to extremes of weather etc.) and propagate
them in quantity in nurseries or cold frames for later
transplanting back to the place of their origin or to
similar micro-sites elsewhere. Assisted regeneration
could also include comparisons of growth and survival
of propagules from both extreme and non-extreme
sites in standardized greenhouse conditions to
analyse genotype differences or separate genotypic
from phenotypic characteristics.. Other adaptive
characteristics could also be explored, e.g. trees with
earlier or later flowering times than other individuals
of a population, drought tolerance or disease
resistance, or healthy conifers with particularly thick
rugose bark, exceptional sap flow, good sapwood-to-
heartwood ratios, or deep root systems for wildfire
and bark beetle resistance. This is an area for much-
needed genetic research.
Box 10.1.
Suitability of germplasm to site

GENET IC CONS IDERATI ONS IN ECO SYSTEM RE STORATI ON USING N ATIVE TRE E SPECIE S
113
created primarily by windfall trees and snow or
wind breakage. Stand-replacing wildfires oc-
curred rarely and only during long droughts. This
was not because the dominant tree species – Sitka
spruce (Picea sitchensis), western red cedar (Thuja
plicata), western hemlock (Tsuga heterophylla)
and Douglas-fir (Pseudotsuga menziesii) – were
fire-resistant (they are not) but because of the ex-
tremely moist environment.
Perhaps the best way to approach the extent
and ecological significance of burning by indig-
enous peoples is to think of numerous small to
medium-sized patch-burns occurring every one to
15 years or so and scattered across the landscape.
The cumulative ecological effects of these fre-
quent low to intermediate disturbances on eco-
system productivity and biodiversity were ampli-
fied by the frequency of these fire events.19 These
numerous, regularly burned patches exponen-
tially multiplied ecotones, maintaining high bio-
diversity and quality wildlife habitat (Anderson,
2005). However, it was not necessary to burn eve-
rywhere. Indigenous peoples were very aware of
the need for unburned wildlife cover and for pro-
tecting shade-adapted plants. Fuel-breaks were
made and ridge tops were kept open to check fire
spread into neighbouring watersheds, with back-
burns employed to protect vegetation and leave
thermal cover for deer and elk. Burns in previous
years acted as fuel-breaks for burns in the current
year. Some sacred places also escaped burning, as
well as selected brush-fields left to senescence be-
fore being harvested for fuelwood (Chester King
in Blackburn and Anderson, 1993).
Burning was highly selective. It was typically
performed in those vegetation types that pro-
duced significant amounts of cultural plants,
were prime wildlife habitat and high in species
richness: riparian zones, wet and dry prairies,
wetlands and marshes, pine and oak savannas
19 In addition to fire, these disturbance events included a num-
ber of horticultural techniques, including pruning and coppicing,
weeding, planting, seed sowing, tillage (women regularly dug in
numerous tracts to harvest a variety of geophytic corms called
“Indian potatoes”) and erosion control (Anderson 2005; Turner
2005).
and woodlands and mountain meadows. Burning
was also extensively utilized to create and main-
tain small to medium-sized gaps and larger mead-
ows in relatively resource-poor forest types such
as those dominated by coast redwood (Sequoia
sempervirens) and Douglas-fir, e.g. the “Bald
Hills” of coastal northern California (Lewis, 1973;
Bonnicksen et al., 1997), which have now lost over
30 percent of their former extent as a result of
invasion of Douglas-fir since circa 1910.
What is the relevance of the fire-maintained
forest to modern restoration and silviculture?
We can begin to answer this question by con-
sidering the structure of the firescape managed
by indigenous peoples. If one counts old-growth
conifer stumps on west, south and east slope as-
pects in much of the interior, one frequently finds
approximately 20 to 65 stumps per hectare in
clumps, compared with 10 to 80 times that num-
ber of trees in a mid-successional state at present
(Martinez, personal observation). With less com-
petition from younger trees killed by repeated
fires and with more sun, many old-growth coni-
fers were practically almost open-grown, with full
crowns extending close to the ground, structured
like a carrot (called “grouse ladders” or “wolf
trees” by loggers).
Fire set by indigenous peoples, and to a much
lesser degree fire started by lightning strikes, was
the main architect of forest structure and compo-
sition, favouring important fire-adapted species.
One should think of the pre-industrial forest as
a slowly changing assemblage of multi-aged spe-
cies, including a mix of dominant mature and
old-growth hardwoods and conifers, with fire
recycling all seral stages of vegetation develop-
ment at the landscape scale (Senos et al., 2005)
– a kind of relatively stable “steady-state shifting
mosaic” (Perry, 1994). It is important not to con-
flate post-harvest early successional vegetation,
mostly a diverse and unstable mix of (frequently
introduced) annuals and short-lived perennials or
shrubs, with more stable long-lived native peren-
nial bunchgrasses, forbs and shrubs that are pe-
riodically renewed by intentional burning. Think
of a slow turnover or shifting of early, mid- and

THE STATE OF THE WORLD’S FOREST GENETIC RESOURCES – THEMATIC STUDY
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114
late-successional species together at one time and
in one forested landscape.
What are we restoring? Our indigenous refer-
ence model suggests a variable forest structure
with well-spaced conifer and hardwood trees
and tree clumps mixed with patches and irregular
colonnades and corridors of more-closely spaced
trees and shrubs, with all age classes of most
historical species represented (Anderson, 2005;
Senos et al., 2005). The newly created openings
of varying sizes would be repopulated (either by
natural regeneration if propagules remain on site
or by replanting and/or reseeding/plugging; see
“Ecological anchors” in Box 10.2) with restored
bunchgrasses, forbs, and shrubs, approaching the
historical species-rich understorey and meadow
Methods and strategies based on traditional
ecological knowledge (TEK) and traditional
resource management (TRM)
Kipuka
strategy:
Kipuka
is a native Hawaiian term
used to describe a rock outcrop that lava spewing
from volcanoes goes around instead of covering.
Used in the context of ecocultural restoration,
it suggests repeated islands or groups of trees,
shrubs, ferns, forbs and grasses, i.e. the fine-grained
landscape created and maintained through judicious
use of fire. These patches and openings, including
meadows, range from a size that is equivalent to
the height of surrounding trees (patches) to as
large as several hectares (meadows) depending
on site conditions, elevation, forest type and
restoration objectives. Objectives are not limited
to trees. They include species-rich understories and
meadows. While some – but by no means most –
timber harvesters leave irregular islands to imitate
fire effects, kipukas are more about the actual
restoration of firescapes, not their imitation. Fire
cannot be imitated in most of its effects on soils
and vegetation. This is as much about restoring
composition as structure. For example, spot-burns
or pile-burns are often done in openings following
thinning. These small patches will sometimes
gradually fill with bunchgrasses and forbs seeded in
the ashes. These kipukas, together with the hit-and-
miss effects on understorey herbaceous vegetation
of low-intensity burning, contribute to forest floor
heterogeneity. These species-rich openings and
meadows can be part of irregular herbaceous
corridors that could be described as “flower trails”
that guide pollinators and seed-carriers across
the landscape (Anderson, 2005). Instead of one-
time irregular structural manipulations by timber
harvesters, kipukas are meant to be periodically
burned. Frequent interventions have the cumulative
effect of increasing species richness and diversity.
Variable density management or variable
green retention: This is a method developed
by forest ecologist Jerry Franklin. The objective is
to “release” future old-growth and commercial
trees from competition by smaller trees and brush,
resulting in repeating sunny openings alternating
with repeating areas of thick shady to partly shady
vegetation. (The proportion of shade to sun extent
will depend on forest type and site moisture regime.)
In a relatively homogenous stand, the seedlings,
saplings and poles with the fullest crowns and the
largest diameters are retained, as are deformed
trees, slow growing trees or trees on harsh sites.
In the first case, it is hoped that these will become
very large, healthy trees that will reproduce their
superior characteristics over time. In the latter case,
it is hoped that at least a few of the poorer trees
will possess genes for exceptional drought and heat
tolerance, disease resistance etc., and that these
genes will be reproduced and perhaps multiplied
in the forest over time. (For selection criteria for
herbaceous understorey plants, see “Ecological
anchors,” below.)
Box 10.2.
Insights on diversifying a gene pool and restoring biodiversity through ecocultural
restoration forestry and ecosystem-based adaptation to climate destabilization/
global warming in temperate far western North America

GENET IC CONS IDERATI ONS IN ECO SYSTEM RE STORATI ON USING N ATIVE TRE E SPECIE S
115
Redundancy: Just as good engineering requires
structural redundancy to ensure safe structures in
case a component fails, so environmental components
– such as vegetation spatial combinations, closed–
open and sun–shade contrasts, wildlife guilds,
prey–predators, pollinators–seed carriers, flowering
plant diversity, food webs, down wood and snags
and compositional diversity – are repeated across the
landscape. If one component fails, others of a similar
class can take up the slack. Redundancy or risk-
spreading is a principal goal of VDM.
Landscape heterogeneity: This concept is
more than just genetic, structural and compositional
diversity. Landscape heterogeneity means that
managers actually look for and map unique micro-sites
that are harsher and warmer and that could serve
as possible sources for individuals, populations and
subspecies that are better adapted to global warming.
Random sampling of a particular species is not as likely
to pick up adapted plants as doing a stratified and
focused field meander that may reveal populations
better adapted to harsh or hot sites. For example,
United States Forest Service researcher Connie Millar,
working in California’s Sierra Nevada mountains, has
noted a downward movement of some endangered
white bark pines, finding seedlings established
in cooler lower elevation sites. There are many of
these micro-sites, especially in areas like the Coast,
Cascade, Sierra Nevada or Klamath mountains that are
topographically diverse. Restorationists will have many
opportunities in projects to find unique heterogeneity
in micro-sites, such as tree windfall, slash piles, large
down wood, topographic depressions, mesic or very
dry places, stream banks, rocky outcrops etc. (This is
discussed in more depth in Box 10.3 in the context of
building-in resilience to change.)
Ecological anchors: This is a method developed
by Canadian forester Herb Hammond. An ecological
anchor is any environmental component that will
assist managers working in more homogeneous
stands (e.g. tightly and uniformly spaced plantations
or typical dense mid-successional forests) to
determine which trees to thin, and those working
in less dense forests with the need for increasing
understorey genetic and compositional diversity.
Examples include sun-loving herbaceous species
that are culturally preferred, ecologically significant
or endangered species that are being shaded out
by trees, or culturally and ecologically valuable
hardwoods still in the stand (e.g. oaks) that need
release from overtopping conifers. Trees that are
shading out these species would be thinned to allow
the understorey to recover.
Conversely, trees protecting important shade-
adapted species in the understorey would be retained.
This method puts as great an emphasis on ecology
as on merchantability in determining which trees
to leave, i.e. removal of a conifer to release oaks
or leaving a conifer as an anchor for shading even
though it does not necessarily possess good or the
best merchantable qualities.
Stepping-stone habitats for linking
conservation reserves with forest matrices and
providing connectivity: Conventional conservation
wisdom divides forested landscapes into two spheres:
reserves and the matrix surrounding the reserves that is
sacrificed to timber. In fact, the matrix probably already
has good-quality habitat that could be linked within the
matrix and to nearby reserves, providing connectivity.
Reserves alone generally do not possess sufficient
topographic and other kinds of diversity for wildlife
habitat and for climate refugees. Linking up to the
matrix could amplify good habitat and connectivity,
provide cooler micro-sites that could increase
The refugial capacity of the forest to harbour plant
and animal climate refugees and contain harsher or
warmer sites that could provide adapted propagules
for ecosystem-based climate adaptation. It may also
facilitate gene flow between reserves and matrix, and
between sources and refuges (Society for Ecological
Restoration International Science and Policy Working
Group, 2004).
Box 10.2. (continued)
Insights on diversifying a gene pool and restoring biodiversity through ecocultural
restoration forestry and ecosystem-based adaptation to climate destabilization/
global warming in temperate far western North America

THE STATE OF THE WORLD’S FOREST GENETIC RESOURCES – THEMATIC STUDY
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flora and providing opportunities for future non-
timber and cultural products (Martinez, 2008 [un-
published]). Forest ecologist Jerry Franklin calls
this variable density management (VDM) or vari-
able retention management (VRM) (Lindenmayer
and Franklin, 2002). Depending on forest type, el-
evation, slope aspect and site conditions, a certain
number of young and mature trees with good po-
tential old-growth characteristics will be retained
as future permanent old growth. This is in addi-
tion to trees with good potential for future com-
mercial grade timber. Periodic management inter-
ventions (thinnings/harvests and prescribed fire)
will be performed periodically over decades so
that site conditions will not be changed too rap-
idly during any one intervention (see “Ecological
Anchors” in Box 10.2).
Unlike standard silviculture with trees plant-
ed and/or thinned to regular and even grid-like
spacing (especially in plantations), and with only
one or two dominant, even-aged commercially
valuable conifer species, it should be clear by
now that the forest influenced by indigenous
peoples is decidedly diverse, irregular and une-
ven-aged and fire-tolerant except for extreme
weather-driven fire events (i.e. made up of nu-
merous small even-aged stands from previous
small fire events in an overall uneven-aged for-
ested landscape; catastrophic fire events that we
see today were extremely rare and always fol-
lowed long periods of drought), with a species-
rich understorey, the nature of which depends
on whether trees are retained or removed.
Cultural and other non-timber products mostly
come from diverse forest understories (and from
oak and pine woodlands, savannas, prairies and
wetlands). This is a forest that is managed for
both relative stability and productivity (both are
a function of forest diversity) and for creating
and maintaining a balance between forest use
and conservation/restoration. The indigenous
model (TEK/TRM) shows us that careful and eco-
logically informed use is a prerequisite for di-
versity and productivity. Indeed, forest use must
further conservation and restoration as far as
possible, while they in turn must sustain use and
forest products. For more detail on how to bal-
ance silviculture/non-timber and cultural prod-
ucts with ecocultural restoration, see Martinez,
2008 [unpublished].
Instead of following United States and
Canadian agency recommended bole-to-bole
(trunk) tree spacing guidelines, crown-to-crown
spacing (measured from outer foliage [crown]
limits of one tree to outer foliage limits of an-
other) provides for greater tree and tree group-
ing separation. As Canadian forester Herb
Hammond notes, this makes for “better ecologi-
cal choices for leaving trees than arbitrary stand
density or basal area choices,” including “habitat
requirements for various [animal] species and
maintenance of stand level diversity” (Hammond,
2009). The primary problem with relying on basal
area (the total space occupied by tree boles per
hectare) is that it is only a cumulative value and
tells us little about their spatial arrangement in a
particular place.
The sooner thinning occurs, the better chance
trees have of achieving their genetic growth po-
tential (usually by 30 to 40 years of age). Some
remaining unthinned trees with sub-merchanta-
ble characteristics should be retained in case they
have genes for exceptional drought and heat tol-
erance or disease resistance, or are good wildlife
trees. While promising optimum characteristics of
commercial and potential old-growth trees are
important, indigenous peoples’ holistic philoso-
phy values the whole forest more than individual
trees, i.e. not sacrificing biodiversity and wildlife
habitat for optimum timber production. Cultural
and commercial use must further conservation
and restoration for the whole forest. This is the
essence of TEK: reciprocity is required when using
plant and animal “relatives.” It is also important
genetically: we may be sacrificing forest adaptive
capacity to climate destabilization by eliminat-
ing too many non-commercial grade trees. For
more detail on timber harvesting rotations, see
Martinez, 2008 [unpublished].
Commercial harvesting is part of the VDM thin-
ning process over several decades of multiple en-
tries. Sustainable logging will continue but would

GENET IC CONS IDERATI ONS IN ECO SYSTEM RE STORATI ON USING N ATIVE TRE E SPECIE S
117
be limited to harvest rotation cycles of 120 to 160
years for the entire stand, with parts harvested
over shorter intervals within the stand. (Further
north, harvest rotations for boreal forest should
be considerably longer.) While timber volume may
be reduced, longer rotations will ensure sustain-
ability over the long term, while prescribed burn-
ing will reduce the cost of wildfire control and
timber losses by significantly reducing hazardous
build up of fuel. Sufficient numbers of trees must
be retained to replace those lost through harvest-
ing and natural mortality, using ratios ranging
from 3:1 to 5:1, depending on forest type and site
conditions (Martinez, 2008 [unpublished]).
Fire-hazard-reduction goals require the re-
moval of ladder fuels, i.e. small and intermediate
trees that can carry ground fires into canopies.
Trees of different ages and sizes need to be seg-
regated to break up contiguous fuels. The stand
structure will, for the most part, be even-aged
groupings in an overall uneven-aged forest. This
is in fact the historical forest that resulted from
frequent low to moderately severe fires. Each
discreet tree grouping dated from a different
small fire event. Succession arrested by indig-
enous practices in interior forests favoured ear-
lier successional conifer species such as pines and
mid-successional tree species such as Douglas-fir
– valuable commercial species today. Advantages
of earlier successional species for fire-hazard
reduction include lower crown bulk densities
(less biomass weight per cubic metre of foliage),
self-pruning that removes fire-vulnerable lower
branches and deeper feeder roots that can avoid
excessive soil heating. Ground fuels were regu-
larly consumed by burning by indigenous peo-
ples, with charred large down wood sometimes
lasting a very long time. Frequent low- to mod-
erate-intensity fires leave a long-lasting (3000 to
12000 years) legacy of charcoal that gradually
mixes into the top metre of soil and sequesters
carbon as well as providing cation-exchange
sites, increasing forest productivity (Deluca and
Aplet, 2008).
Regular fire fertilized the forest by cycling nu-
trients and when combined with reduced com-
petition from smaller trees and brush, allowed
the genetic potential for optimum growth to be
realized. Managed burning contributed, along
with lightning-ignitions, to the healthy old-
growth giants that were used to build our cities.
Prescribed fire directly assists silviculture and re-
duces or eliminates the need for broadleaf her-
bicides to control competing deciduous plants.
A healthy ecosystem supports healthy timber,
and ongoing sustainable timber harvesting and
fire-based silviculture in their turn contribute
to the maintenance of restoration by repeated
harvest thinnings in perpetuity, with prescribed
burning acting as the principal architect of forest
structure. Reconnecting timber harvesting with
ecosystems means, in part, reconnecting with in-
digenous fire regimes. Reconnecting with indig-
enous fire regimes means reconnecting with TEK
and TRM, acknowledging the environmental
legacy of indigenous peoples and its relevance
today, and recognizing the environmental con-
ditions that influenced the genetic structure
of many species over a very long time as they
co-evolved with indigenous fire practices and
other disturbances – human and otherwise. As
ethnobotanist Kat Anderson writes: “Landscapes
are not just assemblages of species; rather, they
are expressions of human evolution and species
behavior. The adaptation of plants and animals
that exist today are responses to past sequences
of environmental conditions” (Anderson, 2005).
Those past sequences were induced in large part
by indigenous burning practices.
Local and traditional ecological knowledge
based on qualitative observational approaches
and Western experimental and quantitative ap-
proaches are increasingly being seen as comple-
mentary. As climate disruption continues to af-
fect ecosystems and cultures at multiple spatial
and temporal scales, observational data on sites
that are not easily manipulated experimentally
are becoming critically important. Even research
sites that appear environmentally similar can be
different enough to compromise experimental
results. There is a real possibility of climate dis-
ruption exacerbating already degraded ecosys-

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118
tems, causing them to cross potentially irrevers-
ible thresholds or tipping points well before we
are aware of it happening (Herrick et al., 2010).
It is for these reasons that place-based indig-
enous peoples are in a privileged position to
maintain and monitor conservation and restora-
tion in their homelands, ground-truth Western
science’s more generalized experimental and
remote technological approaches and contrib-
ute to a more sustainable and biodiverse silvi-
culture.20
20 For example, in 1979, Western scientists using passive micro-
wave technology discovered that Arctic sea-ice was losing extent
and thinning, yet Inuit and Iñupiat peoples knew this in the early
1960s – approximately 15 years earlier.
Gene flow is facilitated by variable density
management, which is based on thinning that
recreates the clumpy nature of forest trees of variable
sizes under traditional resource management,
including enough open spaces between tree
groupings to allow “exchange of alleles among
individuals and populations” (Friederici, 2003). This
box focuses on the ponderosa pine forests of the
southwestern United States, but the same principles of
free gene flow also apply to many other overstocked
forest types in far western North America – all
influenced historically by indigenous burning regimes.
While some geneticists maintain that conifers do not
generally have a problem with gene flow, most forests
in far western North America are impenetrably dense,
with stocking rates as high as 7000 trees/ha or more.
This is very likely to result in different patterns of
gene flow compared with the more open and clumpy
nature of the historical forest under indigenous
management, and before effective fire-suppression
policies began in the early twentieth century.
These small clumps are “genetic neighbourhoods”.
Dendroecologist Joy Nystrom Mast (in Friederici,
2003) explains:
“A group or clump of half-siblings is often created
by a single older tree in the clump, but pollinated
by different trees. As a result…unlimited pollen
movement and hence gene flow among clumps
helps prevent detrimental levels of inbreeding.
Any highly inbred seedlings are subject to reduced
reproductive rates, growth, and survivorship, and are
usually outcompeted by outcrossed individuals…
thereby reducing future levels of inbreeding within
clumps.”
Older trees became established and competed
in environmental conditions different from their
offspring – more-open stand conditions for older
trees and more-crowded conditions for younger
trees. This difference affects allele diversity so that,
if conditions change and little gene flow occurs
between older and younger trees, future adaptive
capacity to changed conditions may be lost. Thinning
too heavily can lead to loss of low-frequency or rare
alleles. Thinning too lightly could prevent unimpeded
gene flow between clumps. Clump spacing should be
of a distance appropriate to forest type and density
so that gene flow is neither too hindered nor too free.
Younger trees should be maintained along with older
trees. Indeed, forest restoration prescriptions should
specify that representatives of all age-classes and all
native species be retained on site (unless they are
very invasive generalist native species that are already
abundant). This is an area for genetic research.
How much distance is required between clumps of
which forest type, so that gene flow can occur while
preventing loss of low-frequency or rare alleles? It
should be noted that more than one entry will be
necessary to approach pre-industrial forest structures.
Multiple entries over decades are usually necessary
in order to change forest environments at a rate that
trees and other species can adapt to appropriately in
the future. Therefore another question is: how much
should be thinned in one entry in what forest type?
Can genetic research help here?
Box 10.3.
Extent and nature of gene flow across fragmented agro-ecosystems

GENET IC CONS IDERATI ONS IN ECO SYSTEM RE STORATI ON USING N ATIVE TRE E SPECIE S
119
Knowledge gaps and possible fruitful genetic
research
Considering the importance of regular intentional
burning, one wonders how both burning and se-
lective harvesting of plants may have altered their
genetic structure, much as plant breeders do to-
day through selective cross-pollination. This was
not proto-agriculture; rather it was indigenous
agro-ecology that, like Western agriculture, influ-
enced the local abundance, availability, composi-
tion, distribution and characteristics of plant (and
animal) species. The cumulative effects of fre-
quent burning of small patches carried fire effects
to much of the forest. Frequent, low-intensity
burning needs to be studied in order to reveal its
effects on forest productivity and genotype selec-
tion of culturally favoured tree species.
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practice. In D. Apostle & M. Sinclair, eds. Restoring
the Pacific northwest: the art and science of eco-
logical restoration in Cascadia, pp. 393–426. Eds.).
Washington, DC, Island Press.
Society for Ecological Restoration International
Science & Policy Working Group. 2004.
SER international primer on ecological restora-
tion. Washington, DC, Society for Ecological
Restoration International (available at http://
www.ser.org/resources/resources-detail-view/
ser-international-primer-on-ecological-restoration).
Sproat, G.M. 1868. The Nootka: scenes and studies of
savage life. London, Smith, Elder and Co.
Stewart, O. 2002. Forgotten fires, edited by H.T. Lewis
& K.M. Anderson. Norman, OK, USA, University of
Oklahoma Press.
Turner, N. 2005. The Earth’s blanket: traditional teach-
ings for sustainable use. Vancouver, BC, Canada,
Douglas and McIntyre Ltd.
Turner, N. 2010. Plants of Haida Gwaii. Wenlaw, BC,
Canada, Sononis Press.

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Reforestation is often seen as a necessary part
of any rehabilitation process once land becomes
degraded. Depending on how it is done, refor-
estation can improve biodiversity conservation,
stabilize hill slopes and improve watershed pro-
tection. However, designing any reforestation
programme raises a variety of problems, particu-
larly when several landholders are involved. This
is because it is rarely possible to restore forest
cover over the entire area, raising questions such
as just how much reforestation should be done,
what kind of reforestation should be carried out
and where these new forests should be estab-
lished. In most cases these questions are resolved
through the actions of individual landholders
acting independently and without reference to,
or knowledge of, the planned actions of other
landholders. Unfortunately, such an individualis-
tic approach is unlikely to result in a satisfactory
outcome. This is because many ecological pro-
cesses, such as gene flow, operate at large land-
scape scales and the collective effect of many ad
hoc decisions is unlikely to be as effective in re-
storing these ecological processes and function-
ing as a more strategic set of interventions that
carefully target key localities and specify the type
of reforestation carried out at each site. A more
strategic intervention necessitates some degree
of coordination across the landscape mosaic. This
means that, in addition to how much, what type
and where to reforest, a fourth question must be
considered: how to organize reforestation on a
landscape scale.
11.1. How much reforestation?
There is no simple answer to the question about
how much reforestation is needed. It depends on
how much natural forest remains and on the at-
tributes of the biota that are present in the land-
scape and are vulnerable to extinction because of
past deforestation. It will also be influenced by
the land-use practices on the cleared land and
on the socioeconomic circumstances of the peo-
ple living in the area. Some private landholders
may be interested in reforestation on part of
their land but much will depend on the oppor-
tunity costs of doing so. A strong timber market
or a market for ecosystem services (e.g. carbon
sequestration) may increase the attractiveness of
reforestation but only if the landholders believe
they will benefit from it.
In the case of biodiversity conservation, nu-
merous studies have shown that deforestation
results in a loss of species proportional to the
deforested area. However, once forest cover in
the landscape falls below 20–30 percent, the
spatial patterns and size of the forest fragments
become more important in determining species
survival than proportion of forest cover per se
(Andren, 1994). Yet it is difficult to prescribe a
Chapter 11
Designing landscape mosaics
involving plantations
of native timber trees
David Lamb
Centre for Mined Land Rehabilitation, University of Queensland, Australia

THE STATE OF THE WORLD’S FOREST GENETIC RESOURCES – THEMATIC STUDY
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122
minimum threshold target of forest cover for
those undertaking reforestation. Different spe-
cies have different habitat requirements; some
will be affected by deforestation well before the
forest area falls below 30 percent while others
will persist even when the forest cover is lower.
Perhaps the best that can be said is that more
forest cover is better than less and that a land-
scape with a large area of forest will conserve
more species and more diversity within species
than one with less cover. It is usually difficult to
predict how many species will return once a cer-
tain amount of restoration takes place. It is also
usually not possible to specify whether a particu-
lar species will recolonize a particular site: much
depends on the type of reforestation carried out
and the quality of the habitats created.
11.2. What kind of reforestation?
The best type of reforestation for biodiversity con-
servation is that which is structurally complex and
involves many native plant species. Some form of
ecological restoration that eventually leads to the
re-establishment of the former forest ecosystem
(e.g. natural regeneration or multispecies plant-
ings) would obviously be ideal. However, most in-
dustrial tree plantations use simple monocultures
of exotic, fast-growing tree species because they
generate a rapid financial return. Many small-
holders also favour these simple monocultures
when they grow trees for commercial purposes,
although many farmers in the tropics also practise
various forms of agroforestry that can involve a
number of tree species.
However, it can be possible to have a much
greater number of species present across a land-
scape even when the number of species at a par-
ticular site is small. This is because site conditions
vary (necessitating the use of different species)
and because different landholders have different
goals or aspirations. Differences in site conditions
and goals can lead to a mosaic of tree monocul-
tures of different species and considerable land-
scape heterogeneity.
The wildlife species most likely to benefit from
such monocultural plantings are those best de-
scribed as habitat generalists. These are species
able to utilize a wide variety of habitat types and
they are rarely among those classified as endan-
gered or vulnerable. A wider variety of species,
including some with more specialized habitat
requirements, can colonize monocultural planta-
tions if these are grown on longer rotations and
are not too distant from natural forests that act as
sources of colonists. In these circumstances large
numbers of tree species may eventually colonize
the site (Keenan et al., 1997). Initially these plants
simply provide a structurally complex understorey
but over time the colonists can grow up and add
structural complexity to the canopy layers. This
increases the value of the plantation as a wildlife
habitat.
An alternative form of reforestation is to estab-
lish timber plantations containing several species.
These may not have as many species as would
be used in ecological restoration but, if carefully
designed, can provide goods such as timber and
non-timber forest products as well as habitats for
some wildlife (Lamb, 2011). Their value in con-
serving biodiversity is further enhanced if any har-
vesting operations are infrequent. Such plantings
are also likely to be more effective in stabilizing
hill slopes and providing watershed protection
than simple monocultures. Again, these may be
colonized over time by further species if managed
on long rotations and located near natural forest.
Different landholders are likely to have differ-
ing views on the merits of these various forms
of reforestation (i.e. monocultures, multispecies
plantations, ecological restoration) and, because
of this, many landscapes could end up having rep-
resentatives of all types of reforestation.
11.3. Where to undertake
reforestation?
There are several ways of addressing the ques-
tion of where reforestation efforts should be
concentrated. Farmers interested in reforestation

GENET IC CONS IDERATI ONS IN ECO SYSTEM RE STORATI ON USING N ATIVE TRE E SPECIE S
123
for commercial purposes may simply plant trees
at sites not suitable for agricultural crops. Areas
close to roads or timber markets may be espe-
cially attractive. Landholders more interested in
reforestation for biodiversity conservation have
two choices. One is to identify those areas where
reforestation will help conserve small popula-
tions of species that are vulnerable to extinction.
These might be isolated remnant patches of for-
est where the populations of some species are
declining because their habitat areas are limited.
Reforestation that enlarges these habitat areas
could allow the populations of such vulnerable
species to increase. A second approach is to in-
crease the connectivity between remnant for-
est patches to allow the linkage of populations
of species that are reproductively isolated from
each other. This might be done by creating corri-
dors between the patches of natural forest or by
establishing small patches of forest within an ag-
ricultural landscape that might act as “stepping
stones” and enable a species to move across that
landscape between areas of natural forest. This
would foster genetic interchange between the
several populations and effectively increase the
overall population size. As noted above, the type
of reforestation undertaken at a site will influ-
ence which species can use the newly reforested
areas. But even monocultures can be useful be-
cause they begin the process of creating a forest
environment.
11.4. How to plan and implement
restoration on a landscape
scale?
All reforestation involves trade-offs and this is es-
pecially the case when it is being done at a land-
scape scale. Some landowners may be quite happy
to reforest some parts of their land because the
opportunity costs of doing so are low or because
they are interested in the goods or ecosystems
services that reforestation can provide. Others
may be unwilling to undertake reforestation
because they perceive the (opportunity) costs of
doing so as being too high. Of course, individual
farmers are not the only stakeholders involved.
Other stakeholders include downstream water-
users, wildlife conservationists, sawmillers and
the broader community. Some of these are likely
to have views that are quite different to those
of local farmers, meaning it can be very difficult
to get agreement on a reforestation programme
that balances the wishes of individual landholders
with those of the broader community.
Much land-use planning has been based on
what might be referred to as a top-down ap-
proach. This often involves technical specialists
working for a government agency and follow-
ing certain prescriptions or guidelines. The ad-
vantage of this approach is that these planners
can take a broad overview and make a judgment
about what should be the best balance between
competing interests in a particular area. Some
sophisticated computer-based tools have been
developed to assist these planners, including
some that can be used to optimize conservation
benefits (Chetkiewicz, St. Clair and Boyce, 2006;
Millspaugh and Thompson, 2009; Thomson etal.,
2009). But this top-down, model-driven approach
has a number of weaknesses. These include the
fact that species differ in their habitat require-
ments and a reforestation programme that
suits one species may be unsuitable for another.
Likewise, the process must make arguable as-
sumptions about trade-offs between different
environmental benefits. Conservation of biodi-
versity is important but, for many stakeholders, so
too is watershed protection or the maintenance
of hydrological flows. Lastly, the process focuses
on where to intervene but not on how to induce
landholders to comply. It relies on compulsion
(which is politically costly), compensation (which
is financially expensive) or universal cooperation
(which is improbable).
There is an alternative. Experience from many
places suggests some kind of consultative plan-
ning process that incorporates both bottom-up
and top-down approaches may be better than
either approach alone. It may not lead to the
most efficient design but it is likely to generate

THE STATE OF THE WORLD’S FOREST GENETIC RESOURCES – THEMATIC STUDY
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124
an outcome that is more acceptable to stake-
holders and hence more likely to be maintained
over time (Reitbergen-McCracken, Maginnis and
Sarre, 2007). The main stages in such a process
are as follows.
First, develop a landscape-level view of the
problem. This involves gathering data about the
existing biophysical and socioeconomic landscape
mosaics, including the distribution of species,
land ownership patterns, the economic circum-
stances of landholders and the trends in land use.
Document the presence of rare, endangered or
vulnerable species, together with information on
threats to biodiversity conservation, such as inva-
sive species or wildfires.
Second, engage with the stakeholders. Identify
landholders and other stakeholders and obtain
their views concerning future land-use practices.
Third, identify reforestation possibilities. Based
on the preceding stages, develop a variety of re-
forestation scenarios that differ in the amount,
type and location of reforestation activities.
Evaluate the advantages and disadvantages of
each scenario to the community and to individual
stakeholders.
Fourth, decide on a reforestation plan. Consult
with stakeholders and decide on a reforestation
plan and timetable. This may involve using incen-
tives or compensation to obtain the agreement
of landholders occupying key locations (e.g. pay-
ment to landholders for the ecosystem services
that reforestation on their land provides). It may
also mean having to accept a suboptimal out-
come for the sake of getting an agreement (an
ideal restoration plan may have to progress in
stages over a period of some years).
Finally, implement the plan, monitor the out-
come and practise adaptive management. The
final stage in any landscape restoration plan is
to monitor it over time to ensure that the plan is
actually implemented and that it generates the
outcomes expected. Restoration can sometimes
lead to unanticipated results and it may be nec-
essary to intervene at a later date to ensure that
biodiversity is indeed being conserved and that
stakeholders remain supportive.
11.5. Will forest landscape
restoration succeed in
conserving all biodiversity?
From a conservation viewpoint, a landscape mo-
saic involving timber trees will have significant
advantages over a homogeneous agricultural
landscape. Even simple plantation monocultures
surrounding small remnant patches of natural
forest will help protect these from further dis-
turbances and provide additional habitats for at
least some of the species they contain. Corridors
and small, scattered patches of trees, including
monocultures, are likely to assist some species to
move across an otherwise hostile environment.
Landscape restoration also initiates a process of
positive feedback in which wildlife, able to move
across the landscape, assist in dispersing the seed
of many plant species.
However, these types of landscape reforesta-
tion may not be enough to ensure the survival of
all species. In the case of plants, those with large
seeds are less likely to be able to be dispersed
across landscapes either because they have no nat-
ural dispersal agent or because that agent is ab-
sent or present only in small numbers in degraded
forests. The only way such species can be reintro-
duced to the landscape is to deliberately include
them in revegetation programmes. The animal
species of most concern are the habitat specialists,
especially those occupying upper trophic levels
and needing large home ranges. Partially forested
agricultural mosaics are unlikely to be sufficient
for such species and protected areas containing
large areas of natural forest are likely to be the
only way such species will be conserved.
11.6. Conclusion
It is difficult to develop reforestation designs that
enhance the capacity of agricultural landscapes to
conserve biodiversity. The problem is partly con-
cerned with ecological issues but is largely to do
with obtaining a consensus among stakeholders
about the amount, type and location of any tree-

GENET IC CONS IDERATI ONS IN ECO SYSTEM RE STORATI ON USING N ATIVE TRE E SPECIE S
125
planting. Apart from full ecological restoration,
the best outcome would be extensive areas of
multispecies plantations involving native species
managed on long rotations. Such plantations are
likely to provide a valuable complement to areas
of natural forest that form part of a protected-
area network.
References
Andren, H. 1994. Effects of habitat fragmentation on
birds and mammals in landscapes with different
proportions of suitable habitat – a review. Oikos, 71:
355–366.
Chetkiewicz, C., St. Clair, C.C. & Boyce, M.S. 2006.
Corridors for conservation: integrating pattern and
process. Annu. Rev. Ecol. Evol. Syst., 37: 317–342.
Keenan, R., Lamb, D., Woldring, O., Irvine, T. &
Jensen, R. 1997. Restoration of plant diversity be-
neath tropical tree plantations in northern Australia.
Forest Ecol. Manag. 99: 117–132.
Lamb, D. 2011. Regreening the bare hills: tropical forest
restoration in the Asia–Pacific region. Dordrecht, The
Netherlands, Springer.
Millspaugh, J. & Thompson, F.R. 2009. Models for
planning wildlife conservation in large landscapes.
Burlington, MA, USA, Academic Press.
Reitbergen-McCracken, J., Maginnis, S. & Sarre, A.
2007. The forest landscape restoration handbook.
London, Earthscan.
Thomson, J., Moilanen, A.J., Vesk, P.A., Bennett, A.F.
& MacNally, R. 2009. Where and when to revege-
tate: a quantitative method for scheduling landscape
reconstruction. Ecol. Appl., 19: 817–828.

THE STATE OF THE WORLD’S FOREST GENETIC RESOURCES – THEMATIC STUDY
PART 2
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When the Doi Suthep-Pui National Park near Chi-
ang Mai in northern Thailand was established in
1981 it contained a large population of Hmong
people who had been living in the area for many
years. These people initially practised shifting
cultivation but, over time, had changed to more
sedentary forms of agriculture. The villagers have
neither Thai citizenship nor legal land tenure. Be-
cause of this they have had an acrimonious rela-
tionship with the park managers, who see them
as illegal occupants destroying the conservation
values of the park.
In order to resolve these differences and to
plan a reforestation programme that would
cover some of the deforested lands, members of
the University of Chiang Mai organized and fa-
cilitated a meeting between park managers and
the Hmong villagers (Elliott et al., 2012). Both
the villagers and park managers had full knowl-
edge of the park and of the particular areas over
which there was some disagreement. Where the
opinions differed was in what should be done
about the disagreements. On the first day, the
facilitators met with National Park staff to de-
termine their view of the problems and to seek
ideas about a way forward. On the second day,
the facilitators met with representatives of the
village to seek their views. On the third day, the
two groups were brought together. The Head of
the National Park described what he saw as the
problem and how the villager’s livelihoods might
be met in future. A representative of the villag-
ers then gave their perspective on the problems
they faced and on a way forward. Guided by the
facilitators, discussions then took place on how
these views could be reconciled. This included
having the participants acknowledge (i) that for-
est conservation was something that both groups
supported, (ii) that some cleared areas should
be reforested to protect water supplies, and (iii)
that villagers could continue to practise agricul-
ture on some of the land currently being used but
that their future economic opportunities lay with
tourism and employment outside the Park.
Having achieved this common understand-
ing, some prospective locations for reforestation
within the Park were identified. This was done us-
ing maps derived from satellite imagery and GPS
mapping prepared prior to the meeting. These
showed the extent of the agricultural cropland,
including orchard areas and annual cropping ar-
eas. Prospective reforestation areas were then
identified on a laptop brought to the meeting.
On the final day of the meeting a visit was made
to the field where the alternative reforestation
options were discussed. These discussions cov-
ered the extent of reforestation, the location
of the areas to be reforested and the types of
Insight 8
Identifying and agreeing on
reforestation options among
stakeholders in Doi Suthep-Pui
National Park, northern Thailand
David Lamb
Centre for Mined Land Rehabilitation, University of Queensland, Australia

GENET IC CONS IDERATI ONS IN ECO SYSTEM RE STORATI ON USING N ATIVE TRE E SPECIE S
127
Figure I8-1.
Location of the Doi Suthep-Pui National Park in northern Thailand
reforestation to be undertaken at each area. A
final reforestation plan was then negotiated. This
involved a programme of ecological restoration
using native tree species based on techniques de-
veloped for the area over a number of years by
Elliott et al. (2006).
Two factors in particular appeared to help make
the process successful. One was that the facilita-
tors were well known to both parties and had
worked in the area for many years. Second, there
were detailed maps showing exactly what each
group had proposed. It was important that these
could developed in time to be taken into the field
on the last day of the meeting, where they gave
participants confidence that they under stood the
trade-offs being made.
References
Elliott, S., Blakesley, D., Maxwell, J.F., Doust, S. &
Suwannaratana, S. 2006. How to plant a forest:
the principles and practice of restoring tropical
forests. Chiang Mai, Thailand, Chiang Mai University.
Elliott, S., Kuaraksa, C., Tunjai, P., Toktang, T.,
Boonsai, K., Sangkum, S., Suwanaratanna, S. &
Blakesley, D. 2012. Integrating scientific research
with community needs to restore a forest landscape
in northern Thailand: a case study of Ban Mae Sa
Mai. In J. Stanturf, P. Madsen & D. Lamb, eds. A
goal-oriented approach to forest landscape restora-
tion, pp. 149–162. Dordrecht, The Netherlands,
Springer.

Part 3
METHODS

GENET IC CONS IDERATI ONS IN ECO SYSTEM RE STORATI ON USING N ATIVE TRE E SPECIE S
131
Many restoration approaches and methods focus-
ing on native species have been developed and
fine-tuned over the years, reflecting the diversity
of species and ecosystems, degradation factors,
stages and socioeconomic contexts. In Part 3,
some of the scientists who have developed these
approaches or have been most active in promot-
ing them describe some of the most widely ap-
plied and studied methods and their principles. In
many cases these descriptions are complemented
by case studies. The general methods are divided
into those focusing on ecological restoration
(Chapter 12) and those that also include produc-
tion objectives for timber or non-timber products
(Chapter 13), although the distinction between
these two categories is not always clear and
many of the methods yield systems that produce
multiple benefits. Approaches used for restor-
ing specific habitats and degradation conditions,
such as mangroves, dry lands and previous mine
sites, are presented separately, as these usually
require specific attention on restoring not only
vegetation but also soil properties and hydrology
(Chapter 14). Finally, three approaches for restor-
ing genetic diversity of particular threatened tree
species are described (Chapter 15).
Figure 3.0.
Geographical overview of the main sites applying the methods presented in the study

GENET IC CONS IDERATI ONS IN ECO SYSTEM RE STORATI ON USING N ATIVE TRE E SPECIE S
133
12.1. Miyawaki method
Akira Miyawaki
Japanese Center for International
Studies in Ecology, Institute for Global
Environmental Strategies, Japan
The Miyawaki method (Miyawaki, 1993, 2004)
was developed by integrating two concepts, the
first based on the study of potential natural veg-
etation and the second derived from observation
of Japanese sacred forests (Chinju-no-mori) re-
newed for centuries by monks, who planted seed-
lings of many species simultaneously.
The approach consists of planting seedlings
of the maximum possible number of tree spe-
cies that characterize the potential natural veg-
etation, from pioneer species to late-successional
ones. From the day they are planted the seedlings
change the ecology of the site, and the species
and individual trees undergo natural selection
through competition, resulting in the creation of
a diversified natural forest.
The restoration process can be divided in four
phases (Figure 12.1):
1. Definition of the potential natural vegetation:
The potential natural vegetation is described
by studying relict vegetation. Field survey data
are subjected to descriptive phytosociologi-
cal analyses, leading to the identification and
mapping of potential vegetation units.
2. Intervention planning: This phase identifies the
species required and determines the amount of
planting stock needed to establish the forest.
Surrounding areas are identified where the
propagation material for the production of
planting stock can be found.
3. Execution plan: This is divided into two stages:
a) Preparation of the material and the site.
The area to be restored is prepared by
adding topsoil from surrounding native
forests, straw and, where possible,
components of the understorey vegetation
of the neighbouring woods. Before
planting, the planting stock is acclimatized
for one to four weeks in the surrounding
areas, either under the shelter of existing
vegetation or under an artificial sheltering
system.
b) Plant seedlings that have extensive root
systems randomly at high density (3–5 indi-
viduals per square metre) and mulch with
straw or other organic materials.
4. After-planting operations: Weed and irrigate
once or twice, if necessary, during the first two
years.
The result is a diversified forest that is left to
grow naturally after the first two years.
The most innovative element of the Miyawaki
method is the application of the concept of “con-
temporary succession.” This assumes that the
Chapter 12
Ecological restoration approaches

THE STATE OF THE WORLD’S FOREST GENETIC RESOURCES – THEMATIC STUDY
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134
native species normally associated with different
successional stages, when planted simultaneously,
generate an “assisted succession” (human-
supported succession concept) that allows the
development in a few decades of the relatively
stable late-successional stage.
Planted simultaneously, all the species become
part of a rapid succession. After the first phase
of rapid growth, there is a natural selection of
species (and individuals) best suited to micro-
sites, and the plantation will evolve into a late-
successional stage without the need for further
action, through what can be described as a new
succession theory (Figure 12.2).
Climate, soil and topography interact to cre-
ate a certain type of climax forest. Human
Figure 12.1.
The four phases of the Miyawaki method
Source: Miyawaki (1999).

GENET IC CONS IDERATI ONS IN ECO SYSTEM RE STORATI ON USING N ATIVE TRE E SPECIE S
135
intervention alters the plant cover. At this point
two alternatives can be envisaged:
1. Classical succession: let nature take its course;
wait two to ten years for the annual herb com-
munity to be replaced by perennial grasses, an-
other 10 to 15 years for a community of shrubs
to develop, 15 to 50 years for the heliophilous
tree species to develop and finally 200 to 300
years or more for the late-successional species
to establish.
2. New succession: simultaneous planting of
seedlings (2–5/m2) of species belonging to the
potential natural vegetation. This can give
rise to a semi-natural environment very simi-
lar to a young late-successional forest within
40–50years.
This is explained by positive interactions among
the diverse species planted. Immediately after
planting the microclimate changes, becoming
more favourable for the young plants. Rather
Figure 12.2.
Comparison between Miyawaki’s new succession theory and classical succession theory
Source: Miyawaki (1999).

THE STATE OF THE WORLD’S FOREST GENETIC RESOURCES – THEMATIC STUDY
PART 3
136
than suffering from competition from neighbour-
ing plants, during the first months after planting
the seedlings benefit fromthe positive effects (e.g.
lower soil temperature during the day, windbreak
effect or mitigation of extreme heat). Beneficial
micro-site effects improve water availability and
soil stabilization. Over time, natural selection will
lead to the survival of the best-adapted individuals.
The Miyawaki method has been used in over
1700 sites around the world, on extensive areas
as well as to establish windbreaks along roads and
railways (Miyawaki, 1998). Since 1971, over 40
million native trees have been planted using this
method. In 2012 Dr Miyawaki launched the Green
Tide Embankment project, which is using the
Miyawaki method to establish a green embank-
ment all along the Japanese coast damaged by the
2011 tsunami to protect it against future events.
“On March 11, 2011, Eastern Japan suffered major
damage from the Great East Japan Earthquake and
tsunami that followed. We conducted surveys on the
disaster areas. The surveys proved that monoculture
forests of fast-growing intolerant exotic tree species
such as
Pinus thunbergii
(black pine) and
Pinus
densiflora
(red pine) were almost destroyed and some
were carried landward and extended damage by
colliding with people, houses and cars. But forests of
main and companion trees from the local potential
natural vegetation stood firmly and exerted an
influence on reducing the power of the tsunami. Main
tree species of the forests are
Persia thunbergii
and
evergreen
Quercus
(oaks), and companion tree species
are also evergreen broadleaved trees including
Camellia japonica
,
Neolitsea sericea
and
Euonymus
japonicus
.
“After the earthquake and tsunami, huge heaps of
debris remained dispersed in the disaster areas. Debris
is not industrial waste, but natural resources from
the earth. After removing poisonous and inorganic
objects, it should be used effectively. From our results
in reforestation in the Brazilian Amazon, I suggest to
the central and local governments, corporations and
non-profit organizations that we should build mounds
on the coastline of the disaster areas by mixing soil
and debris, and plant indigenous tree species on them
to form quasi-natural forests. Roots of plants also
breathe under the ground. Mounds built of soil and
debris have hollows and contain much air. Therefore
trees can grow well. The forests on the mounds
will function as a breakwater and protect lives and
properties of local people from future tsunamis. I
would like to build the Green Tide Embankment,
300 km long from the north to the south.
“The native forest system will last for 9000 years
until the next glacial age, though there is alternation
of individuals.
“Mature trees, which have grown large enough,
can be cut selectively and utilized for furniture,
architectural materials and other purposes. Forests
coexist with local economies. After selective cutting,
a successor replaces the harvested tree and the forest
ecosystem will be maintained.
“Everywhere in the world, forests consisting of
indigenous trees save lives and property of local
people. Ecological reforestation based on the potential
natural vegetation is indispensable in our safe living
environments and regional economy. Let’s extend the
reforestation movement by planting indigenous trees
proactively, from tropical rainforest regions to other
areas of the world.”
Description of the “The Green Tide Embankment” project from
a keynote speech given by A. Miyawaki at the International
Symposium on Rehabilitation of Tropical Rainforest Ecosystems
at the Universiti Putra Malaysia, Malaysia in October 2011.
Box 12.1.
The Green Tide Embankment

GENET IC CONS IDERATI ONS IN ECO SYSTEM RE STORATI ON USING N ATIVE TRE E SPECIE S
137
References
Miyawaki, A. 1993. Restoration of native forests from
Japan to Malaysia. In H. Lieth & M. Lohmann, eds.
Restoration of tropical forest ecosystem. Dordrecht,
The Netherlands, Kluwer Academic Publishers.
Miyawaki, A. 1998. Restoration of urban green environ-
ments based on the theory of vegetation ecology.
Ecol. Eng., 11: 157–165.
Miyawaki, A. 1999. Creative ecology: restoration of
native forests by native trees. Plant Biotech., 16(1):
15–25.
Miyawaki, A. 2004. Restoration of living environment
based on vegetation ecology: theory and practice.
Ecol. Res., 19: 83–90.
12.1.1.
Tropical rainforest rehabilitation
project in Malaysia using the
Miyawaki Method
Nik Muhamad Majid
Institute of Tropical Forestry and Forest
Products (INTROP), Universiti Putra
Malaysia, Malaysia
The Joint Research Project for the Rehabilitation
of Tropical Rainforest Ecosystems was launched
by Mitsubishi Corporation in 1991, with sup-
port from Universiti Putra Malaysia (UPM) and
Yokohama National University (YNU), Japan.
The project adopted the Miyawaki method (see
above).
The goals of the project include developing
techniques for the rehabilitation of degraded ar-
eas and conducting research to assess the health
of rehabilitated forests.
Site information
The project comprises two sites, one in Sarawak
and the other in Selangor, Malaysia.
The project was initiated in July 1991 on
a 47.5 ha site on the Universiti Putra Malay-
sia Bintulu Campus, Sarawak (113°03’41.67”E;
3°12’32.28’’N). The site previously had been badly
degraded by shifting cultivation activities. The
rehabilitation project in Bintulu was executed
in four phases, and currently the site has sever-
al different-aged forests, the oldest being over
20 years old. These forests give researchers the
opportunity to study various ecological param-
eters at different stages of forest growth.
In 2008, following the success of the Bintulu
project, a new agreement was signed between
UPM and Mitsubishi Corporation to establish a
new model forest by planting indigenous tree
species in an urban setting, using 27 ha of degrad-
ed land located inside the UPM Serdang Campus
(101°43’32.27’’E; 2°59’45.16’’N). The establishment
of this model planted tropical forest was initiated
at a tree-planting ceremony on 26 November
2008 at UPM’s Arboretum, which lies between
the Kuala Lumpur–Seremban Highway, the Kuala
Lumpur–Putrajaya Highway and the railway be-
tween Kuala Lumpur International Airport and
the city. Formerly pastureland, this area was de-
graded by the construction of the railway track
and six-lane highways.
Restoration activities
Malaysia is considered to be one of the world’s
leading mega-diverse biodiversity hotspots, with
tropical rainforest covering an area of more than
7.6 million hectares, or about 70 percent of the
country’s total land area. The country is richly en-
dowed with diverse flora and fauna that have the
potential to be developed and utilized in various
natural products and services. Forests are still the
main source of income for the country. However,

THE STATE OF THE WORLD’S FOREST GENETIC RESOURCES – THEMATIC STUDY
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138
harvesting activities have caused serious degrada-
tion to the forest ecosystems.
A comprehensive research approach was ini-
tiated at the two sites to determine the extent
of damage as well as the effectiveness of the
reforestation and rehabilitation programmes.
The area selected for planting on the Bintulu
site was a coastal forest that included heath and
lowland dipterocarp forests. Tree-canopy species
were selected from the natural vegetation of a
similar area to ensure the suitability of the tree
species to the environment, based on assumption
that indigenous species are well adapted to local
conditions. Seeds and wildings were collected in
Similajau National Park, Likau Forest Reserve, and
the Experimental Forest and Arboretum at the
UPM Bintulu Campus. Planting techniques were
mound planting, open-area planting and partial-
shade planting.
As of 2011, roughly 350 000 seedlings from
126 tree species have been planted in four dif-
ferent areas (Table 12.1). The species planted can
be classified into three groups: light-demanding,
shade-tolerant and slow-growing species. The
light-demanding species include Shorea ovata,
S.mecistopteryx, Artocarpus integer, Pentaspodon
motleyi and Whiteodendron moultianum. The
shade-tolerant species include Shorea macrophyl-
la, S. gibbosa, S. materialis, Hopea beccaariana,
Cotylelobium burckii, Calophyllum ferrugenium,
Parashorea parvifolia and Durio caranatus. The
slow-growing species include Diospyros sp., Hopea
kerangasensis, Palaquium gutta and Vatica sp. In
addition, over 100 research plots have been estab-
lished in the rehabilitated area and the growth of
the planted seedlings is regularly monitored.
A total of 19 500 seedlings, including rare
and endemic species, have been planted on the
TABLE12.1.
Number of species per family planted at the Bintulu and Serdang restoration sites, Malaysia
Family No. of species Family No. of species
Bintulu Serdang Bintulu Serdang
Anarcadiaceae 7 5 Lecythidaceae 1 2
Annonaceae 1 1 Leguminosae 4 6
Apocynaceae 4 1 Melastomataceae 1 –
Araucariaceae – 1 Meliaceae 1 2
Bombacaceae 3 2 Moraceae 5 5
Burseraceae 3 2 Myristicaceae 1 3
Celastraceae – 2 Myrtaceae 9 5
Combretaceae – 1 Olacaceae 1 1
Compositae 1 Oxalidaceae 1 1
Dipterocarpaceae 57 Rubiaceae – 2
Ebenaceae 2 Sapindaceae 4 6
Euphobiaceae 4 Sapotaceae 4 4
Fabaceae – Simaroubaceae 1 –
Guttiferae 5 Sterculiaceae – 3
Icacinaceae 1 Thymelaeaceae 1 2
Irvingiaceae – Tiliaceae – 2
Lauraceae 3 TOTAL 126 117

GENET IC CONS IDERATI ONS IN ECO SYSTEM RE STORATI ON USING N ATIVE TRE E SPECIE S
139
UPM Serdang Arboretum site using open-area
planting. Tree species selection (Table 12.1) was
based on vegetation studies in several forests in
Peninsular Malaysia:
• Ayer Hitam Forest Reserve, a high
conservation value forest of Dipterocarpus
crinitus and Hopea nervosa, representing
the lowland dipterocarp forest of Selangor.
• Semangkok Forest Reserve, representing
lowland dipterocarp forest with Shorea
leprosula vegetation association.
• Pasoh Forest Reserve, representing the
lowland dipterocarp forest of Negeri
Sembilan.
• Leban Condong Forest Reserve, representing
the heath forest of Pahang.
• Rompin forest, representing the swamp
forest of Pahang.
• Segari Melintang Forest Reserve,
representing the Shorea lumutensis
vegetation association since this species is an
endemic in this forest reserve.
• Mersing forest, representing the Shorea
peltata vegetation association since this
species is endemic in this type of forest.
Project outputs
Project outputs to date include the following:
• Forest restoration: A mixed, virgin tropical
rain forest has been recreated through
human innovation (Figures12.3 and 12.4).
The rehabilitated forest has attracted
wildlife and many other plant species, and
has improved soil fertility, the hydrological
cycle and the microclimatic environment.
• Research: Many scientific papers have been
published or presented at national and
international conferences. This project
contributed to UPM being ranked sixth
among 95 universities in the world in the
Green Metric World University Ranking
2010 for promoting sustainability through
environmental conservation and green
technology.
• Public awareness: Over the past two
decades, at least 10 000 people have
participated in planting ceremonies at
the Bintulu project site, and another
2000 people have been involved in the
Serdang project over the past four years
(Figure 12.5). The events were widely
covered by both local and international
media, including the National Geographic
Channel.
• Human capital development: The project has
been the subject of six Ph.D. dissertations,
seven M.Sc. theses and more than 20 B.Sc.
theses.
• Linkages: The Acid Deposition Monitoring
Network in East Asia (EANET), based in
Niigata, Japan, has started a research project
at the Bintulu site as one of its monitoring
stations in the Asia–Pacific region to
evaluate the effects of air pollution on
forest ecosystems.
• Two international symposia were organized
in 1991 and 2011 to discuss recent research
findings and current issues related to forest
rehabilitation and promote international
collaboration among scientists, academics,
policy-makers and forest industry
stakeholders.
Figure 12.3.
Bintulu site before planting (1991)

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12.1.2. Adapting the Miyawaki
method in Mediterranean
forest reforestation practices
Bartolomeo Schirone and Federico
Vessella
AgroForestry, Nature and Energy
Department (DAFNE), University of
Tuscia, Viterbo, Italy
Natural forests in north Sardinia, Italy, have been
degraded over centuries by human activities, such
as livestock husbandry and wood exploitation.
Since 1905, periodic attempts have been made to
reforest the region using traditional techniques,
mainly planting maritime pine (Pinus pinaster
Aiton.), Aleppo pine (Pinus halepensis Mill.), At-
las cedar (Cedrus atlantica (Endl.) Carrière), cork
oak (Quercus suber L.), downy oak (Quercus pu-
bescens Willd.) and sweet chestnut (Castanea sa-
tiva Mill.). The trees were planted at low densities
(300–2200 plants/ha) along contour lines after
forming terraces by subsoiling, or across the slope
in pits. Low planting density has traditionally
been considered appropriate in arid and semi-
arid environments to avoid competition for water
resources between plants. However, there is little
evidence that competitive processes outweigh co-
operative processes, such as mutual shading, that
can enhance seedling survival.
Experimental design
In May 1997, two experimental forest restora-
tion plots were planted in Pattada (Province of
Sassari, North Sardinia) to test the effectiveness
of the Miyawaki method for reforestation. The
Figure 12.4.
Bintulu site (2014)
Figure 12.5.
Planting ceremony at Serdang Site (2014)

GENET IC CONS IDERATI ONS IN ECO SYSTEM RE STORATI ON USING N ATIVE TRE E SPECIE S
141
Miyawaki method involves planting both pioneer
and late-successional species to a target density,
often up to 10 000 or more seedlings/ha, and
has been successful in reducing the time taken
to achieve complete environmental restoration
(see section 12.1). This is the first time that the
Miyawaki method has been tested in Mediterra-
nean Europe.
The trial was conducted by the University of
Tuscia with logistical and monitoring support
from the Regional Forest Directorate of Sardinia
and political support from the Municipality of
Pattada.
Both experimental plots were degraded and
abandoned sites on which several reforestation
projects had failed. A survey of the natural plant
communities in neighbouring areas was conduct-
ed and the climate was characterized to evaluate
the possible natural vegetation for the study sites
and to select appropriate species for the reforest-
ation. Seeds were collected from nearby natural
forest stands and germinated in four greenhous-
es owned by the Regional Forest Directorate of
Sardinia. Seedlings were grown in plastic bags for
one year before being planted out in the field.
Slight modifications were made to the
Miyawaki method. For instance, no new topsoil
was added to the restoration sites, but soil was
tilled to improve soil water storage over the win-
ter and reduce water stress during the summer.
Several mulching materials were used (sawmill
residues, dry and green materials), and no weed-
ing was done after planting. Local climatic con-
ditions were analysed using climate diagrams de-
veloped by Walter and Lieth (1967) to determine
optimal planting time.
Site A (40°37’32’’N, 09°11’08’’E, 760 m above
sea level) covered 4500 m2. Plot preparation con-
sisted of clearing and tilling a series of 3.5-m-wide
strips. Pot-grown tree seedlings were then plant-
ed at a density of approximately 8600 plants/ha.
Site B (40°36’54’’N, 09°10’04’’E, 882 m above sea
level) covered an area of 1000 m2. In contrast
to site A the entire plot was cleared and tilled.
Planting density was approximately 21000 seed-
lings/ha.
The plots were planted with both early-
successional species and late-successional species
to improve resilience of the plant community.
On Site A, 1723 seedlings were planted, belong-
ing to 22 indigenous tree and shrub species:
25 percent pioneers (e.g. Arbutus unedo, Pinus
pinaster, Spartium junceum and Myrtus commu-
nis); 10 percent mid-successionals (e.g. Celtis aus-
tralis, Ligustrum vulgare and Pyrus communis);
and 65 percent late-successionals (e.g. Quercus
suber, Quercus ilex, Acer monspessulanum, Taxus
baccata and Malus domestica). On Site B, 2139
seedlings belonging to 23 autochthonous spe-
cies were planted: 17 percent pioneers (Arbutus
unedo, Juniperus oxicedrus, Pinus pinaster and
Myrtus communis); 14 percent mid-successionals
(Celtis australis, Fraxinus orsnus, Phyllirea lati-
folia and Thymus vulgaris); and 69 percent late-
successionals (Quercus suber, Q. ilex, Ilex aquifo-
lium and Taxus baccata).
Results
The plots were surveyed in 1998, 1999 and 2009.
By 2009 (i.e. 12 years after planting) early-
successional tree species were well established,
with stable populations, and the plots had a
high level of plant biodiversity. Mean mortal-
ity rates for all species were 61 percent in Site
A (672 plants survived) and 84 percent in SiteB
(336plants survived; Table 12.2). The difference
in mortality rate between the sites was mainly
the result of poor drainage in Site B. The forest
species that are most prevalent in local natural
forest (i.e. maritime pine and the oak group)
survived well in both sites, thus maintaining
the possibility of achieving intermediate and
late-successional vegetation stages. In addition,
several indigenous species that had not been
planted were found on the sites (e.g. Erica ar-
borea and Prunus spinosa). The survey results
suggest that cooperative processes (e.g. mutual
shading) facilitated the establishment of some
species, in particular the mid- to late-successional
ones. The high planting densities adopted in the
sites reduced, for instance, the impact of acorn
predators, thus encouraging oak regeneration

THE STATE OF THE WORLD’S FOREST GENETIC RESOURCES – THEMATIC STUDY
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142
(i.e. the main late-successional forest species in
Mediterranean environments) and favoured root
anastomosis processes (connection of normally
separated roots), which seems to influence the
stability of the ecosystem and reforestation suc-
cess (Kramer and Kozlowski, 1979).
While the experiment consisted of only two
small field trials, comparison of the results, in
TABLE 12.2.
Survival of planted seedlings in two plots restored using the Miyawaki method. The seedlings were
planted in 1997 (year 1) and evaluated in 2009 (year 13). Dashes indicate the species was not planted.
Species
Site A Site B
No. of seedlings Survival (%) No. of seedlings Survival (%)
Year 1 Year 13 Year 1 Year 13
Pioneer species
Arbutus unedo
L. 50 41 82 11 0 0
Juniperus oxicedrus
L. – – – 45 30 66.6
Myrtus communis
L. 19 1 5.3 95 4 4.2
Pinus pinaster
Aiton. 273 208 76.2 155 80 51.6
Rosmarinus officinalis
L. 23 15 65.2 23 0 0
Salvia officinalis
L. 500400
Spartium junceum
L. 74 29 39.2 21 0 0
Middle-successional
Celtis australis
L. 22 3 13.6 37 0 0
Fraxinus ornus
L. 8 1 12.5 9 0 0
Ligustrum vulgare
L. 126 29 23.0 13 4 30.8
Phyllirea angustifolia
L. 1 1 100.0 – – –
Phyllirea latifolia
L. – – – 203 0 0
Pyrus communis
L. 19 10 52.6 22 10 45.4
Thymus vulgaris
L. – – – 24 0 0
Late-successional
Acer monspessulanum
L. 21 2 9.5 30 0 0
Castanea sativa
Mill. 42 1 2.4 – – –
Ilex aquifolium
L. 112 23 20.5 125 0 0
Laurus nobilis
L. 22 3 13.6 19 0 0
Malus domestica
Borkh. 21 7 33.3 19 0 00
Quercus ilex
L. 300 159 53.0 394 96 24.4
Quercus pubescens
Willd. 268 116 43.3 93 8 8.6
Quercus suber
L. 11 7 63.6 621 96 15.4
Sorbus torminalis
(L.) Crantz 18 4 22.2 24 8 33.3
Taxus baccata
L. 251 9 3.6 126 0 0
Viburnum tinus
L. 58 3 5.2 26 0 0

GENET IC CONS IDERATI ONS IN ECO SYSTEM RE STORATI ON USING N ATIVE TRE E SPECIE S
143
terms of species densities and choices, plant bio-
diversity and ecosystem composition, with those
of other reforestation practices traditionally ap-
plied in the same ecological context indicates
some interesting differences in the growth perfor-
mance of the species under the Miyawaki method.
Traditional reforestation methods resulted in sim-
pler vegetation structures (Table 12.3). Moreover,
when traditional methods are used, growth per-
formance of secondary species (measured by plant
density and mean height) is severely reduced by
the highly competitive shrub species (Erica arbo-
rea and Arbutus unedo) that occur spontaneously
and in large numbers. In contrast, trees on the
Miyawaki plots developed more rapidly. This was
particularly the case for early-successional species.
TABLE 12.3.
Height of 12-year-old trees in four plots reforested using different methods. Sites A and B were
established using the Miyawaki method, Site C was reforested using traditional pit planting
(785 plants/ha) and Site D employed contour planting on terraces (1048 plants/ha). Dashes indicate
species not planted, and zeros indicate planted species that did not survive in 2009. Data are given
for species for which at least two individuals survived on each plot.
Species Height (cm)
Site A Site B Site C Site D
Pioneer species
Arbutus unedo
L. 32.7 ± 4.1 0 500.0 ± 35.8 110 ± 20.6
Cedrus atlantica
Endl. – – – 162 ± 54.6
Erica arborea
L. – – 115.0 ± 12.7 130 ± 18.6
Juniperus oxicedrus
L. – 36.2 ± 18.5 – –
Myrtus communis
L. 10.0 10.0 ± 1.4 – –
Pinus pinaster
Aiton. 433.2 ± 143.6 325.5 ± 38.6 376.4 ± 73.0 425.7 ± 25.1
Rosmarinus officinalis
L. 89.3 ± 33.9 0 – 80.0 ± 14.9
Spartium junceum
L. 110.7 ± 62.2 0 – –
Mid-successional
Celtis australis
L. 26.7 ± 28.9 – – –
Ligustrum vulgare
L. 32.8 ± 52.6 30 ± 8.16 – –
Pyrus communis
L. 71.0 ± 65.1 60 ± 61.2 – –
Sorbus torminalis
(L.) Crantz 35.0 ± 50.0 40 ± 12.9 – –
Late-successional
Acer monspessulanum
L. 40.0 ± 14.1 0 – –
Ilex aquifolium
L. 45.2 ± 30.6 0 – –
Laurus nobilis
L. 30.0 ± 17.3 0 – –
Malus domestica
Borkh. 100.0 ± 45.5 0 – –
Quercus ilex
L. 34.2 ± 32.1 40.8 ± 36.2 69.4 ± 23.2 146.2 ± 38.1
Quercus pubescens
Willd. 23.6 ± 27.5 10 ± 5.3 – –
Quercus suber
L. 174.3 ± 49.6 77.5 ± 51.9 – –
Taxus baccata
L. 33.3 ± 38.0 0 – –

THE STATE OF THE WORLD’S FOREST GENETIC RESOURCES – THEMATIC STUDY
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144
Conclusions
Overall, this example of restoration using the
Miyawaki method can be considered quite suc-
cessful, even if some further improvements are
required. For instance, early-successional species
may have been planted in excessive numbers,
thus competing with the intermediate- and late-
successional species. Optimal planting density will
have to be tested. An economic analysis should be
performed to compare the costs of reforestation,
including post-planting silvicultural practices, be-
tween traditional reforestation methods and the
Miyawaki method. Planting costs when using the
Miyawaki method are relatively high because of
the high planting density and the associated la-
bour requirements (even with non-specialized la-
bour). On the other hand, the Miyawaki method
requires no post-planting care such as weeding or
thinning.
Even if costs of the Miyawaki method were are
higher than those of the traditional reforesta-
tion techniques, the quality of forest achieved in
a relative short time (i.e. after 12 years), would
make it worth considering for use in protected ar-
eas and natural parks where traditional plantings
are not easily accepted because of their aesthetic
and ecological impacts. In traditional plantings,
trees are placed in regular and fixed schemes,
creating an easily recognizable artificial land-
scape, especially when using exotic species. The
Miyawaki method, on the other hand, restores
forest that is better integrated in the surrounding
landscape because of its use of local species and a
randomized planting scheme that evolves mainly
according to the ecological and competitive pro-
cesses among the species.
References
Kramer, P.J. & Kozlowski, T.T. 1979. Physiology of
woody plants. Orlando, FL, USA, Academic Press.
Walter, H. & Lieth, H. 1967. Klimadiagram-Weltatlas.
Jena, Germany, VEB Gustav Fischer Verlag.
12.2. Framework species method
Riina Jalonen1 and Stephen Elliott2
1 Bioversity International, Malaysia
2 Forest Restoration Research Unit,
Chiang Mai University, Thailand
The framework species method can be a par-
ticularly effective approach for restoring forest
ecosystems where fragments of intact forest
remain within about 10 km of restoration sites.
In this method, selected indigenous tree species
are planted on the restoration site to promote
natural recruitment and succession. The planted
trees shade out weeds to “recapture” the site and
re-establish forest structure. They also reinstate
ecological processes such as litter accumulation
and nutrient cycling. These so-called “framework
species” are selected for their ability to provide
resources (e.g. nectar, fruit and nesting sites) at an
early age. These resources attract seed-dispersing
animals and thus facilitate dispersal of seeds of
non-planted tree species (i.e. recruit species) into
the site from forest remaining in the surrounding
landscape. Improved site conditions (i.e. weed-
free, humus-rich forest floor) favour germination
of naturally dispersed seed and establishment of
tree seedlings (FORRU, 2006; Figure 12.6).
Typically, 20–30 tree species are planted on
each restoration site. Good framework spe-
cies grow fast at seedling stage, rapidly develop
large and dense crowns that shade out weeds,
bear fruit at young age to attract seed-dispersing
animals, and survive well in field conditions, in-
cluding after fire where relevant (FORRU, 2008).

GENET IC CONS IDERATI ONS IN ECO SYSTEM RE STORATI ON USING N ATIVE TRE E SPECIE S
145
Framework species should preferably include
both early- and late-successional forest tree spe-
cies to accelerate natural succession and facilitate
the recovery of a complex forest structure. Many
late-successional tree species can be planted since
they perform well in the open, sunny conditions
of deforested areas. Under normal circumstances
they fail to colonize such areas because of lack of
seed dispersal (FORRU, 2006).
Seeds or wildings of framework trees are usually
collected from nearby forest. Ideally, they should
be collected from as many parent trees as possible
Figure 12.6.
Process of site restoration using the framework species method. Note the positive feedback loops
that reinforce and facilitate ecosystem recovery.
Source: Redrawn from FORRU (2006).

THE STATE OF THE WORLD’S FOREST GENETIC RESOURCES – THEMATIC STUDY
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146
to ensure that a wide range of genetic diversity is
captured. However, because the purpose of the
framework species is to quickly recapture the site,
phenotypically superior parent trees and propa-
gated seedlings should be selected (Blakesley,
Hardwick and Elliott, 2002). After propagation in
a nursery, seedlings are planted on the restora-
tion site at a typical average spacing of 1.8 × 1.8 m
(approximately 3100 trees/ha). The seedlings must
be tended for at least two years after planting by
regular weeding. Other management practices,
including fertilizer application, protection from
wildlife and care of naturally recruiting seedlings,
are also recommended (FORRU, 2006).
Studies have demonstrated the effectiveness of
the framework species method. In a trial plot in
Doi Suthep-Pui National Park, northern Thailand,
73 non-planted tree species had established eight
to nine years after planting framework tree spe-
cies (Sinhaseni, 2008). With 57 planted frame-
work tree species, the total tree species on the
site amounted to 130, equivalent to 85 percent
of the total tree flora expected in an intact forest
in similar area under the same conditions. Most
of the tree species recorded had germinated from
seeds dispersed from nearby forest by birds (par-
ticularly bulbuls), fruit bats and civets. The spe-
cies richness of the bird community also increased
from about 30 species before planting to 88 af-
ter six years, representing about 54 percent of
bird species recorded using the same methods in
nearby intact forest (Toktang, 2005). The species
richness of mycorrhizal fungi and lichens has also
been reported to increase dramatically in the re-
stored plots, often exceeding that of natural for-
est (Nandakwang et al., 2008; Phongchiewboon,
2008).
Little information is available on most tropical
tree species about how they meet the preferred
characteristics of framework species. So far, lists
of good framework species have been published
only for the wet tropics of Queensland, Australia,
where the method originates (Goosem and
Tucker, 1995), and for the seasonally dry tropics in
northern Thailand, where the method is actively
studied and promoted by Chiang Mai University’s
Forest Restoration Unit (FORRU) (Elliott et al.,
2003). FORRU has published detailed guidelines
for studying and identifying framework species
(FORRU, 2008; Box 12.2), and is carrying out re-
search to identify framework species in Cambodia
and other neighbouring countries.
Because the framework species method relies
on natural recruitment of seedlings, its applica-
bility depends on seed dispersers and dispersal
distances of native tree species in the area. In
tropical Asia, seed dispersal by mammals, large
birds and bats is known to occur over distances of
up to 10km, while shorter distances from a few
hundred metres to a few kilometres are probably
more common (Corlett, 2009). Remnant forests
must, therefore, be present within a few kilome-
tres from the restoration site. The nearer the for-
est, the faster will be the recovery of species rich-
ness (FORRU, 2006). Although even scattered trees
can act as seed sources, their seed may be largely
inbred and thus of low quality for restoration
(Blakesley, Hardwick and Elliott, 2002; Chapter2).
Subsequent enrichment planting is recommended
if biodiversity recovery is not evident four to five
years after tree planting (FORRU, 2006).
References
Blakesley, D., Hardwick, K. & Elliott, S. 2002. Research
needs for restoring tropical forests in southeast
Asia for wildlife conservation: framework species
selection and seed propagation. New Forest., 24:
165–174.
Corlett, R.T. 2009. Seed dispersal distances and plant
migration potential in tropical east Asia. Biotropica,
41: 592–598.
Elliott, S., Navakitbumrung, P., Kuarak, C., Zangkum,
S., Anusarnsunthorn, V. & Blakesley, D. 2003.
Selecting framework tree species for restoring
seasonally dry tropical forests in northern Thailand
based on field performance. Forest Ecol. Manag.,
184: 177–191.

GENET IC CONS IDERATI ONS IN ECO SYSTEM RE STORATI ON USING N ATIVE TRE E SPECIE S
147
FORRU (Forest Restoration Research Unit). 2006.
How to plant a forest: the principles and practice
of restoring tropical forests. Chiang Mai, Thailand,
Biology Department, Science Faculty, Chiang Mai
University.
FORRU (Forest Restoration Research Unit). 2008.
Research for restoring tropical forest ecosystems:
a practical guide. Chiang Mai, Thailand, Biology
Department, Science Faculty, Chiang Mai University.
Goosem, S.P. & Tucker, N.I.J. 1995. Repairing the rain-
forest – theory and practice of rainforest re-estab-
lishment in North Queensland’s wet tropics. Cairns,
Australia, Wet Tropics Management Authority.
Nandakwang, P., Elliott, S., Youpensuk, S., Dell, B.,
Teaumroon, N. & Lumyong, S. 2008. Arbuscular
mycorrhizal status of indigenous tree species used
to restore seasonally dry tropical forest in northern
Thailand. Res. J. Microbiol., 3(2): 51–61.
Phongchiewboon, A. 2008. Recovery of lichen diversity
during forest restoration in northern Thailand.
Graduate School, Chiang Mai University. (M.Sc.
thesis)
Sinhaseni, K. 2008. Natural establishment of tree seed-
lings in forest restoration trials at Ban Mae Sa Mai,
Chiang Mai province. Graduate School, Chiang Mai
University, Thailand. (M.Sc. thesis) (available at http://
archive.lib.cmu.ac.th/full/T/2008/biol0308ks_tpg.
pdf).
Toktang, T. 2005. The effects of forest restoration on
the species diversity and composition of a bird com-
munity in Doi Suthep-Pui National Park, Thailand,
from 2002–2003. Chiang Mai University, Thailand.
(M.Sc. thesis)
Relatively few common fruit-eating animals are
responsible for most seed dispersal between intact
forest and restoration sites in northern Thailand.
These include small to medium-sized birds, especially
bulbuls, fruit bats (e.g.
Cynopterus
spp.) and certain
medium-sized mammals, including civets, common
wild pig, common barking deer and hog badger. These
animals are equally at home both in forest and in
deforested areas.
Tree species that are most likely to attract seed-
dispersing animals to restoration sites produce
small to medium-sized fruits within three years
after planting. Such species indigenous in northern
Thailand include
Callicarpa arborea, Castanopsis
tribuloides, Eugenia grata, Ficus abellii, F. hispida,
F. semicordata, F. subincisa, Glochidion kerrii, Heynea
trijuga, Macaranga denticulata, Machilus kurzii,
Prunus cerasoides
and
Rhus rhetsoides
. Some species
also produce flowers with large quantities of nectar
(e.g.
Erythrina subumbrans
). In general, local fig
species (
Ficus
spp.) are good candidates for framework
species because of their fruiting patterns and high
survival even under unfavourable site conditions.
Some tree species can provide nesting sites
for birds within five years after planting, further
enhancing seed dispersal to the site. Such species in
northern Thailand include
Alseodaphne andersonii,
Balakata baccata, Bischofia javanica, Cinnamomum
iners, Duabanga grandiflora, Erythrina subumbrans,
Eugenia albiflora, Ficus glaberima, F. semicordata,
F. subincisa, Helicia nilagirica, Hovenia dulcis, Phoebe
lanceolata, Prunus cerasoides, Pterospermum
grandiflorum, Quercus semiserrata, Rhus rhetsoides
and
Spondias axillaris
.
Source: Forest Restoration Research Unit, 2006.
How to plant
a forest: the principles and practice of restoring tropical forests
.
Chiang Mai, Thailand, Biology Department, Science Faculty,
Chiang Mai University.
Box 12.2.
Examples of tree species for attracting seed-dispersing animals in Thailand

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12.3. Assisted natural
regeneration
Evert Thomas
Bioversity International, Regional office
for the Americas, Cali, Colombia
Approaches to ecological restoration of forest
ecosystems depend strongly on the initial state of
forest or land degradation, as well as the desired
outcomes, time frame and financial constraints
(Chazdon, 2008). In sites with low to intermediate
levels of degradation where soils are generally
intact (typically degraded [Imperata] grassland or
shrub vegetation), natural regeneration of forest
species is often sufficient to trigger the conversion
to more productive forests with relatively little
human intervention. This is what assisted natural
regeneration (ANR) is all about: accelerate, rather
than replace, natural successional processes by re-
moving or reducing barriers to natural forest re-
generation (Shono, Cadaweng and Durst, 2007).
The method was originally proposed by Dalmaico
(1986) and since then has gained considerable
popularity around the world (FAO, 2003, 2012).
One of the attractive characteristics of ANR is its
cost-effectiveness compared with conventional
reforestation methods, most of which have sub-
stantial costs associated with propagating, rais-
ing and planting seedlings (FAO, 2003; Shono,
Cadaweng and Durst, 2007). As ANR involves less
site preparation and nursery establishment, costs
can often be as low as half to one tenth of those
of conventional reforestation practice (FAC and
DANIDA, 2005; FAO, 2012). Furthermore, ANR is
very compatible with traditional systems of natu-
ral resource management, and easily understood
by field staff (FAO, 2003). However, the method
is generally labour-intensive, requiring nearly
constant maintenance of selected forest areas for
five to seven years to ensure establishment of de-
sirable tree species (FAO, 2012). Hence, in order
to obtain successful results it is crucial to involve
local communities.
ANR aims at enhancing the establishment of
secondary forests by protecting and nurturing
the mother trees and their offspring already pre-
sent in the area. This can be achieved by remov-
ing or reducing barriers to regeneration, such as
soil degradation, competition with weedy spe-
cies, and recurring disturbances (e.g. fire, graz-
ing and wood harvesting) (Shono, Cadaweng and
Durst, 2007). Particular care is given to liberating
naturally regenerating seedlings or saplings from
competition with undergrowth by weeding a cir-
cular area around them and to protecting them
from fire and grazing (e.g. through active estab-
lishment of fuel breaks and fences, respectively).
Where two or more seedlings or saplings are close
to each other, the smaller, less healthy or less de-
sirable one is removed and, where appropriate,
transplanted to empty places in the restoration
site (FAC and DANIDA, 2005). In some cases, fer-
tilizer may be applied to promote the growth of
existing seedlings or saplings.
ANR is most applicable in areas with remain-
ing trees or patches of natural forest within a
wider degraded landscape, as these trees provide
propagation material or attract dispersal agents
(birds, bats, mammals, etc). Five hundred to 800
wildings/ha is generally adequate to ensure es-
tablishment of a stable second-growth forest
and eventual restoration of a dense forest cover
(Shono, Cadaweng and Durst, 2007; FAO, 2012).
Wildlife is an essential component in the restora-
tion approach for its role in seed dispersal, and
should therefore be protected (FAC and DANIDA,
2005). Precisely because of ANR’s reliance on nat-
ural processes, it is especially effective in restoring
and enhancing biological diversity and ecological
processes (FAO, 2003). ANR is not to be recom-
mended for ecological restoration in seriously
degraded landscapes as it is likely that remain-
ing isolated trees do not produce viable seeds
or vigorous seedlings (FAC and DANIDA, 2005).
Depending on the desired outcome, quantity and
quality of natural regeneration (e.g. fewer than
500–800 naturally occurring wildlings per ha;
FAO, 2012), time constraints and/or available fi-
nancial resources, stands may need to be enriched

GENET IC CONS IDERATI ONS IN ECO SYSTEM RE STORATI ON USING N ATIVE TRE E SPECIE S
149
with a variety of species, such as fast-growing,
light-demanding species that create shade in the
understorey and a habitat for late-successional
species, orchard trees or commercial tree species.
Thus, it is important to choose a wide variety of
native species matched to different microclimatic
conditions in the restoration area, including spe-
cies that provide fruit for birds, bats and other
animals that spread seed. Once nurse trees and
existing woody species start casting appropriate
shade the stand can be enriched with shade-
tolerant (high-value) species (FAC and DANIDA,
2005). Hence, it is clear that ANR techniques are
flexible and allow for the integration of various
objectives, such as timber production, biodiversity
recovery and cultivation of crops, fruit trees and
non-timber forest products in restored forests
(Shono, Cadaweng and Durst, 2007).
References
Chazdon, R.L. 2008. Beyond deforestation: restoring
forests and ecosystem services on degraded lands.
Science, 320: 1458–1460.
Dalmacio, M. 1987. Assisted natural regeneration: a
strategy for cheap, fast and effective regeneration
of denuded forest lands. Tacloban City, Philippines,
Philippines Department of Environment and Natural
Resources, Region 8.
FAC (Forestry Administration, Cambodia) & DANIDA.
2005. Guidelines for site selection and tree planting
in Cambodia. Phnom Penh, Forestry Administration
(available at http://www.treeseedfa.org/guidelines_
site_eng.htm).
FAO (Food and Agriculture Organization of the
United Nations). 2003. Advancing assisted natural
regeneration (ANR) in Asia and the Pacific, compiled
and edited by P.C. Dugan, P.B. Durst, D.J. Ganz &
P.J. McKenzie. Bangkok, FAO Regional Office for
Asia and the Pacific (available at ftp://ftp.fao.org/
docrep/fao/004/ad466e/ad466e00.pdf).
FAO (Food and Agriculture Organization of the
United Nations). 2012. Assisted natural regenera-
tion of forests [web page] (available at http://www.
fao.org/forestry/anr/en/).
Shono, K., Cadaweng, E.A. & Durst, P.B. 2007.
Application of assisted natural regeneration to
restore degraded tropical forestlands. Restor. Ecol.,
15: 620–626.
For more information on assisted natural regeneration,
see: http://www.fao.org/forestry/anr/en/
12.3.1. Assisted natural regeneration
in China
Jiang Sannai21
Nearly 40 percent of China’s surface area is seri-
ously eroded, and more than one quarter of the
land is covered with desert soils. Since the 1990s,
the frequency of sand storms has increased, es-
pecially in northern China. Assisted natural re-
generation (ANR) has played an important role
in the country’s effort to counter expanding en-
vironmental degradation. In China, ANR can be
divided into two main categories: special ANR
and general ANR. Special ANR is practised on
cutover land with measures such as soil prepa-
ration conducted to improve site conditions for
forest establishment. General ANR refers to more
comprehensive regeneration and afforestation
activities accompanied by artificial sowing, tend-
ing and other treatments. It is conducted on bar-
ren hills, wasteland, barren desert lands, cutover
21 Based on Sannai (2003), published with permission.

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lands, riverbanks with important ecological sta-
tus, sandy regions damaged by wind and such
like. The objective is to establish vegetative cover
to protect the land. In regions where some natu-
ral sowing occurs, the area is closed to most forms
of exploitation for a number of years (depending
on local conditions). Use of the land is restricted
or prohibited during this period to facilitate for-
est establishment from natural seed fall. Where
natural sowing and natural regeneration are
unlikely to occur without human assistance, the
closed areas will be sown with tree and/or grass
seeds from the air.
Closed areas are subject to both administrative
and management measures. Three types of clo-
sure are distinguished in China:
1. Full-closure is adopted for a period of three
to five years or eight to ten years (depending
on local conditions) in regions such as remote
mountains, upper reaches of rivers, water
catchments of reservoirs, sites characterized by
severe soil erosion, desert soil areas subject to
wind damage and other regions where natural
regeneration is difficult.
2. Semi-closure is practised in areas where some
target tree species are growing well and where
the percentage of forest cover is relatively
high. Under semi-closure, strict protection is
prescribed to protect the saplings and seedlings
of target tree species. However, controlled cut-
ting of fuelwood and grass may be allowed.
3. Full-closure and semi-closure are combined
in regions where farmers are very poor and
fuelwood is scarce. Full closure periods alter-
nate with semi-closure periods. There are no
fixed standards; the lengths of full or semi-
closure period vary depending on the progress
achieved in restoring vegetative cover.
Between 2001 and 2003, over 30 million hec-
tares of forest was established through the clo-
sure system. Aerial sowing of tree or grass seeds
was implemented in 931 counties of 26 provinces
(autonomous regions and municipalities directly
under the central government). Approximately
8.68 million hectares of forests were established
through aerial seeding combined with closure.
This accounted for 25 percent of the total artifi-
cial forests of China in 2003.
ANR has played a very important role in ex-
panding forest resources, controlling soil erosion,
retarding the process of desertification, improv-
ing the ecological environment and improving
the living conditions of farmers.
References
Sannai, J. 2003. Assisted natural regeneration in China.
In FAO. Advancing assisted natural regeneration
(ANR) in Asia and the Pacific, compiled and edited
by P.C. Dugan, P.B. Durst, D.J. Ganz & P.J. McKenzie.
Bangkok, Thailand, FAO Regional Office for Asia and
the Pacific. pp. 29–31 (available at ftp://ftp.fao.org/
docrep/fao/004/ad466e/ad466e00.pdf).

GENET IC CONS IDERATI ONS IN ECO SYSTEM RE STORATI ON USING N ATIVE TRE E SPECIE S
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12.4. Post-fire passive restoration
of Andean Araucaria–
Nothofagus forests
Mauro E. González
Forest Ecology Laboratory, Faculty of
Forest Science and Natural Resources,
Universidad Austral de Chile and
Center for Climate and Resilience
Research.
In 2002 large and highly intense fires caused by
lightning strikes affected vast areas of Andean
Araucaria–Nothofagus forests (ANF) within sev-
eral national parks and forest reserves in south-
central Chile. The two worst-affected protected
areas were Tolhuaca National Park and Malleco
National Reserve, with over half of their combined
total area burned (14 536 ha). Considering the sci-
entific and cultural importance of Araucaria arau-
cana (Molina) K. Koch, these events prompted the
rapid development of plans to restore and evalu-
ate the recovery of Araucaria forests in the two
areas. Here, we present a restoration effort that
used a passive approach.
Fire has been an intrinsic ecological process
influencing the ANF, although its role has only
recently been more fully understood (Burns,
1993; González, Veblen and Sibold, 2005, 2010;
Quezada, 2008; Mundo, 2011). Fire regimes in the
ANF are dominated by mixed-severity fires (low-
severity surface fires and crown fires) that result
in a range of fire effects and responses (González,
Veblen and Sibold, 2005, 2010). Thus, fires typical-
ly result in a mosaic of vegetation burnt to vary-
ing degrees of severity (Peñaloza, 2007).
Fires, both naturally occurring and human-
induced, have influenced the ANF over the past
several millennia (Heusser, 1994). Since ancient
times, native tribes of the Araucarian region (e.g.
Pehuenche and Mapuche people) have used the
area for activities such as hunting, grazing and
the collection of Araucaria seeds (Aagesen, 1998;
Tacón, 1999; Bengoa, 2000). Fire was typically
used by native tribes for hunting guanaco (Lama
guanicoe; Veblen and Lorenz, 1988), and possibly
to clear undergrowth vegetation to facilitate the
collection of Araucaria seeds. With the introduc-
tion of domestic livestock (by the late 1500s), fires
were also used to clear travel corridors and ma-
nipulate forage. After the arrival of Euro-Chilean
settlers to the Araucarian region (after 1882)
human-induced fires increased dramatically. Fire
was used as the main tool to clear forests for ag-
riculture and cattle grazing and also to improve
pasture quality in high-altitude wood land val-
leys covered usually with open wood lands of
Araucaria–Nothofagus antarctica (G. Forster)
Oerst. In sum, humans have had a significant
impact on the historical fire regime of these eco-
systems (Gonzaléz, 2005; González, Veblen and
Sibold, 2005; Quezada, 2008).
Post-fire early secondary development as
passive restoration of Araucaria–Nothofagus
forests
Although most details of post-fire Araucaria–
Nothofagus forest recovery still are not com-
pletely understood, secondary succession is rec-
ognized as an important process in ecological
restoration. Research on early secondary succes-
sion provides basic information on key ecologi-
cal processes and species’ responses to enhance
forest restoration activities (i.e. methods and
procedures) and ecosystem integrity. Passive res-
toration – the recovery of forest by natural regen-
eration after fire – has been used as an initial step
before active restoration is implemented. Where
the natural forest response achieves the desired
processes, functions, structure and composition,
restoration may rely mostly on natural recovery.
The present case study is intended to illustrate

THE STATE OF THE WORLD’S FOREST GENETIC RESOURCES – THEMATIC STUDY
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early post-fire ANF recovery in terms of recruit-
ment of tree seedlings and understorey species
under the effect of two different fire severities,
moderate and high.
Study area and sampling design
This study was carried out in subalpine temperate
old-growth Araucaria araucana–Nothofagus pu-
milio (P. et E.) Krasser forests (c. 1200 m above sea
level) within the Tolhuaca National Park (38º12’S
and 71º45’W; Figure 12.7). This area – especially
high-altitude open woodland intermixed with
grassland – was historically used for summer cat-
tle ranching both by Native Americans (c. 1700–
1900) and later (1900–1960) by early settlers,
who established nearby the protected areas. At
lower elevations (less than 900 m above sea level)
Nothofagus forests were selectively harvested be-
tween 1940 and 1960.
During the summer the Tolhuaca National Park
is visited by many people for camping, fishing and
trekking. These recreational activities, especially
trekking, have been negatively affected by the
2002 fire because of the danger from fallen dead
trees.
After the 2002 fire, we established six perma-
nent plots of 1000 m2 to evaluate vegetation re-
covery in areas affected by mid- and high-severity
fire. Moderate fire severity killed 40 percent to
60 percent of trees and consumed most of the
undergrowth. High fire severity killed more than
95 percent of trees and the undergrowth was
completely consumed. In each plot we meas-
ured the diameter at breast height (dbh) of all
live and dead trees with a dbh greater than
5cm. Vegetation response was evaluated in 30
subplots of 1 m2 systematically laid out in each
of the six plots, where we counted the number
of saplings (less than 5cm dbh and greater than
2 m height) and seedlings of tree species (less
than 2m height), and estimated the abundance-
dominance of undergrowth species using Braun-
Figure 12.7.
Location of Tolhuaca National Park in the province of Malleco, Araucanía region, Chile. Circle
indicates the study site, classified based on severity of fire.

GENET IC CONS IDERATI ONS IN ECO SYSTEM RE STORATI ON USING N ATIVE TRE E SPECIE S
153
Blanquet cover-class values. Importance value
for each species was determined by calculating
the sum of the relative frequency and relative
cover.
Tree seedling regeneration under different
fire severities
Fire severity significantly influenced tree seedling
recruitment in the ANF (Figure 12.8). Recruit-
ment of Nothofagus species was greatest follow-
ing fire of moderate severity, which left remnant
trees alive. Seedlings of N. dombeyi (Mirb.) Oerst.
and N. pumilio originated from wind-dispersed
seeds. Seedlings of N. nervosa (Phil.) Krasser
originated from basal resprouts and also from
wind-dispersed seeds. By contrast, recruitment
of medium-sized pioneer and opportunistic trees
such as Embothrium coccineum J.R. et G. Forster
and Lomatia hirsuta Diels ex Macbr. (Proteaceae
family) was greater following fire of high se-
verity. Recruitment of these species originated
through resprouting of basal buds (when indi-
viduals were present in the former, more open
stands) and from wind-dispersed seeds. Seedlings
of Araucaria araucana established either from
gravity-dispersed seeds – seeds protected inside
the cones of female trees (González et al., 2006)
– or resprouts from basal buds of burned juvenile
trees (Figure 12.9). Seedlings of N. pumilio were
unable to establish following high-severity fire.
This fire-sensitive species (González, Veblen and
Sibold, 2005) is an important component of the
original forest stand. Given that it is dependent
on seeds for recruitment and its seeds have only a
limited range of dispersion, remnant trees are key
for its successful re-establishment.
Bamboo (Chusquea culeou Desv.) colonized
the sites more rapidly than any other understo-
rey species (Figure 12.10). This species reached
importance values (IV) of 32 percent follow-
ing high-severity fire and 62 percent following
moderate-severity fire. Moderate-severity fire
lightly burned some patches of bamboo, causing
minor damage to the rhizome system and allow-
ing a rapid response. Other understorey species
reached higher IVs following high-severity fire.
Figure 12.8.
Post-fire recruitment of seedlings in Araucaria–Nothofagus forests burned by fires of medium and
high severity

THE STATE OF THE WORLD’S FOREST GENETIC RESOURCES – THEMATIC STUDY
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154
These included Muelhenbeckia hastulata (J.E.
Sm.) Johnst., Alstroemeria aurea R. Graham,
Dioscorea brachybotrya Poepp., Ribes magel-
lanicum Poir and Gaulteria phillyreifolia (Pers.)
Sleumer. The cosmopolitan species Senecio vul-
garis L. (Asteraceae) colonized the site via wind-
dispersed seeds, forming relatively dense patches,
especially in sites with a lower cover of Chusquea
culeou affected by high-severity fire. Other in-
vasive weeds (e.g. Hypochoeris radicata L. and
Cirsium vulgare (Savi) Ten.) were brought to the
sites by cattle which occasionally grazed the area.
Figure 12.10.
Araucaria–Nothofagus stand affected by a high severity fire. Note the dense resprouting of bamboo
culms (Chusquea culeou)
Figure 12.9.
a) Seedlings of Araucaria araucana from seeds that survived high temperature inside the cones
b) basal resprout of a relatively juvenile individual of Araucaria araucana
a) b)

GENET IC CONS IDERATI ONS IN ECO SYSTEM RE STORATI ON USING N ATIVE TRE E SPECIE S
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Main conclusions
Fires burn forest stands with different severities,
providing various opportunities for species re-
cruitment. Fire severity influences the amount of
organic material destroyed and hence the amount
and types of biological material that remain after
fire. The number of live trees and reproductive
structures remaining below ground has an impor-
tant influence on post-fire regeneration of woody
species. The recruitment of the obligate seeders
Nothofagus dombeyi and N. pumilio was gener-
ally low following high-severity fire. The almost
complete eradication of the adult population of
both fire-sensitive species contributed to low seed
availability post-fire and hence restricted oppor-
tunities for seedling establishment in the post-fire
environment. In contrast, Araucaria araucana and
N. nervosa, which are capable of resprouting af-
ter fire, were able to establish following both me-
dium- and high-severity fires. Seedlings of Arau-
caria established under remnant female trees.
Even though the impact of the presence of do-
mestic livestock has not been evaluated, observa-
tions indicate that livestock can have an important
influence on the process of forest recovery. In the
early stages of recovery, trampling, grazing and
browsing have significant detrimental effects on
tree seedling survival and growth, especially for
the most palatable species (i.e. all Nothofagus spe-
cies). Moreover, the combined effect of severe fires
and cattle would favour the dominance of shrub
species (e.g. Chusquea culeou; Raffaelle et al.,
2011). The high tree mortality and burning of the
undergrowth seem to promote weed growth and
facilitate (or attract) the presence of cattle.
These preliminary findings indicate some gen-
eral considerations and recommendations to
enhance a passive restoration approach. First, it
is important to recognize that fire severity influ-
ences stand composition and structure following
the fire, which may result in different successional
pathways in the forest community. That could
be the case especially for severely burned stands
where Chusquea culeou, an undergrowth species,
can outcompete the relatively poor tree seedling
recruitments because of its strong ability to rapidly
resprout from its rhizome system, which covers
the site at high density. Second, it is important
to evaluate the responses of the dominant tree
species, especially for fire-sensitive species, to the
new (abiotic and biotic) conditions following fires
of different severities. Although N.dombeyi and
N. pumilio typically establish with high density
after stand-replacing fires (Mera, 2009; González,
Veblen and Sibold, 2010), a very severe fire can
hamper or delay the successful establishment of
the main canopy species. Under this scenario, ac-
tive restoration could be implemented by supple-
mentary (enrichment) planting of tree seedlings.
Third, the post-fire environment of strong light,
bare soil and lack of groundcover competition
provides a temporary opportunity for abundant
recruitment of weed species. Therefore, monitor-
ing and controlling exotic weeds and domestic
cattle is an important measure to favour suc-
cessful passive restoration of the burned forests.
Passive restoration together with a little active
assistance could be an effective way to restore
ecosystem function, integrity (community compo-
sition and structure) and sustainability (resistance
to disturbance and resilience).
Acknowledgements
This research has received funding from the
Seventh Framework Programme of the European
Union (FP7/2007- 2013) under Project No. 243888
and CONICYT/FONDAP/15110009, Chile.
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the Andean cordillera of south-central Chile. Austral
Ecol., 35: 597–615.
Heusser, C. 1994. Paleoindians and fire during the late
Quaternary in southern South America. Rev. Chil.
Hist. Nat., 67: 455–442.
Mera, R. 2009. Etapas de desarrollo de rodales mixtos
postfuego de Araucaria araucana (Mol.) Koch y
Nothofagus dombeyi (Mirb) Blume, en el Parque
Nacional Villarrica. Universidad Austral de Chile,
Valdivia, Chile. (Tesis de Ingeniero Forestal)
Mundo, I.A. 2011. Historia de incendios en bosques de
Araucaria araucana (Molina) K. Koch de Argentina
a través de un análisis dendroecológico. Universidad
Nacional de La Plata, La Plata, Argentina. (Tesis
Doctoral)
Peñaloza, R. 2007. Zonificación de la severidad de
incendio natural y su distribución topográfica cuan-
titativa en el Parque Nacional Tolhuaca, IX región.
Universidad Austral de Chile, Valdivia, Chile. (Tesis de
Ingeniero Forestal)
Quezada, J. 2008. Historia de incendios en bosques de
Araucaria araucana (Mol.) Koch. del Parque Nacional
Villarrica, a partir de anillos de crecimiento y regis-
tros orales. Universidad Austral de Chile, Valdivia,
Chile. (Tesis de Ingeniero Forestal)
Raffaele, E., Veblen, T.T., Blackhall, M. & Tercero-
Bucardo, N. 2011. Synergistic influences of
introduced herbivores and fire on vegetation change
in northern Patagonia, Argentina. J. Veg. Sci., 22:
59–71.
Tacón, A. 1999. Recolección de piñon y conservación de
la Araucaria (Araucaria araucana (Mol.) K. Koch.):
un estudio de casos en la Comuna de Quinquén.
Universidad Austral de Chile, Valdivia, Chile.
(Magíster en Desarrollo Rural)
Veblen, T.T. & Lorenz, D.C. 1988. Recent vegetation
changes along the forest/steppe ecotone of northern
Patagonia. Ann. Assoc. Am. Geogr., 78: 93–111.

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12.5. Carrifran Wildwood:
using palaeoecological
knowledge for restoration
of original vegetation
Philip Ashmole
Borders Forest Trust, Monteviot
Nurseries, Ancrum, Jedburgh, United
Kingdom
Carrifran valley in the Southern Uplands of Scot-
land was denuded when Borders Forest Trust
(BFT) purchased it by public subscription in 2000
and commenced restoration of a “wildwood.”
The vision of the Wildwood Group (a somewhat
devolved element within BFT) was that by remov-
ing negative anthropogenic factors and initiat-
ing woodland development by planting, it might
be possible to restore a broadleaved forest and
moorland ecosystem similar to that which existed
in the 650 ha valley about 6000 years ago. This
was a time when the primary forest of Scotland
probably reached its greatest extent and diversity,
following immigration of all major tree types (Tip-
ping, 1994). The subsequent loss of natural forest
was caused primarily by human activity. Climate
and soils also changed to some extent, but the
large altitudinal range of the site, coupled with
the great variety in aspect, slope and substrate,
led to an expectation that most of the original
species of fungi, plants and animals would still
encounter suitable conditions somewhere on the
site, thus enabling restoration of nearly original/
natural woodland (Peterken, 1996, 1998).
Having secured Carrifran, the Wildwood Group
organized a discussion meeting to provide a basis
for restoration of broadleaved native woodland
on this site and elsewhere in southern Scotland
(Newton, 1998; Newton and Ashmole, 1998).
Previously, attention of Scottish environmental-
ists had focused mainly on the Highlands and
especially on native pinewoods, where different
considerations might apply.
Choice of appropriate woody species for
planting on site was of immediate concern.
Palaeoecology can supply a record, although inev-
itably incomplete, of the taxa that have occupied
a site at various times in the past. At Carrifran a
core through the peat at 620 m in Rotten Bottom
provided palynological data extending back to
the early Holocene; this was supplemented by
data from two other sites within 5 km of Carrifran
(Tipping, 1998). This information was used as one
basis for a list of tree and shrub species considered
native to Carrifran valley (Newton and Ashmole,
1999).
The uncertainties associated with pollen analy-
sis make it desirable to supplement the palaeo-
ecological record with other types of informa-
tion. For instance, existing ancient woodlands can
demonstrate the suitability of a region for those
tree and shrub species that occur within them. In
the Southern Uplands, however, surviving ancient
woodlands are rare, isolated, small and often lin-
ear (Badenoch, 1994) and it has been argued that
their use as a template for ecological restoration
of a denuded site might lead to establishment of
a woodland that was a degenerate and species-
impoverished reflection of the past (Tipping,
1998).
This danger can be countered to some extent
by use of the national vegetation classification
(NVC), which is based on information from a wider
range of British sites (Rodwell, 1991). This makes
it possible to use existing open-ground vegeta-
tion as a predictor of the appropriate composi-
tion for native woodland to be established on the
site (Rodwell and Patterson, 1994; Averis, 1998).
Caution is required because the NVC framework
is based on existing British vegetation types, but
most woodlands and other natural habitats have
been subject to a variety of human influences

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over many centuries. These include selective man-
agement for useful species and lack of protection
from grazing and browsing by domesticated her-
bivores (Smout, MacDonald and Watson, 2005).
NVC analysis was used at Carrifran, however, to
decide on the most appropriate composition for
woodland in the various parts of the site, which
differ in altitude, aspect, slope, soil and moisture.
Additional insight was obtained by the use of
ecological site classification (ESC) analysis devel-
oped by the Forestry Commission, which is based
on assessment of three principal factors: climate,
soil moisture and soil nutrient regime (Pyatt and
Suárez, 1997). At Carrifran this analysis was par-
ticularly influential in emphasizing the role of
juniper (Juniperus communis L.) on the high pla-
teau around the rim of the valley.
Choice of appropriate species is a crucial first
step, but must be linked with a strategy for ob-
taining appropriate seeds or cuttings to establish
on site. Genetic advice was clear: the aim should
be creation of a dynamic and expanding wood-
land resource with the capacity to evolve in the
future and respond to environmental change
(Ennos, 1998). To achieve this aim, planting stock
should be sourced from relict ancient woods near
to the restoration site and with conditions match-
ing it as closely as possible. However, populations
in isolated and small woodland remnants may
have low genetic variability and differ from one
another in genetic composition because of his-
torically low population sizes. In order to maxi-
mize genetic diversity in the new population it
is necessary to collect propagation material from
numerous individuals and from several different
suitable sites.
Some of the natural Scottish tree populations
listed by Wilson, Malcolm and Rook (2000) seem-
ed appropriate as sources of seed for Carrifran,
but other batches of seed were obtained from
woodland fragments that appeared ancient, even
though there was no documentary evidence of
their status. For some species there were particu-
lar problems. For instance, in sessile oak (Quercus
petraea (Mattuschka) Liebl.), the desired oak spe-
cies for Carrifran, hybridization between native
Figure 12.11.
Carrifran valley

GENET IC CONS IDERATI ONS IN ECO SYSTEM RE STORATI ON USING N ATIVE TRE E SPECIE S
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trees and planted pedunculate oak (Quercus
robur L.) made most local populations suspect. In
the early years of planting some acorns were ob-
tained from Cumbria, but doubts about the status
of the woods there led to a switch to Galloway as
the main source. However, a few remote woods
in Cumbria are probably ancient and may contain
trees adapted to high altitudes, so a special effort
was made to collect seed from them for planting
in the high parts of Carrifran.
Aspen (Populus tremula L.) was also a problem-
atic species, since it is now represented in south-
ern Scotland only by widely scattered, small and
often clonal stands. Yet it was an early colonizer
of Scotland and may have been a significant com-
ponent of pristine native woodland on a wide
variety of soil types and from sea level almost to
treeline (Quelch, 2002). By collecting root cut-
tings for propagation from about 20 different
stands and planting the progeny in many parts
of Carrifran, it was hoped that a representation
of aspen similar to that in the natural woodland
would eventually be achieved.
Now that many trees are more than a decade
old it has become obvious that the rate of tree
growth decreases markedly between the floor
of the valley, at around 250 m above sea level,
and the upper limit of the main planting at
450–550m above sea level, which may be around
the timberline (Ashmole, 2006). In recent years
the Wildwood Group has paid special attention to
planting above this level, in an attempt to restore
treeline woodland and montane scrub, habitats
that have been almost entirely lost from Scotland
(Hester, 1995; Gilbert, Horsfield and Thompson,
1997; Ashmole, 2006; Chalmers and Ashmole,
2007).
Field observations and data on accumulated
temperature and relative windiness indicated
that the land above about 750 m above sea level
would not support woodland or scrub and that
there would be a natural transition to montane
moss-heath, which would extend to the wind-
swept summit of White Coomb (821 m), the
fourth highest peak in the south of Scotland
(Hale, Quine and Suárez, 1998; Adair, 2005).
However, a high hanging valley at Carrifran pro-
vided an opportunity to attempt establishment
of montane scrub between 600 m and 750 m
above sea level. Some 12500 shrubs were plant-
ed between 2007 and 2012, and although mor-
tality is significant and growth very slow, scrub
vegetation is becoming established. Emphasis
has been on juniper, sourced from the highest
available natural populations in the area, and
on downy willow (Salix lapponum L.), which in
Britain has relict populations in only three locali-
ties south of the Scottish Highlands, one of them
only 1 km from Carrifran.
Thirteen years after the start of the restoration
work at Carrifran, over half a million trees have
been planted and about 300 ha of native wood-
land are well established in the lower half of the
valley. Ground vegetation is changing rapidly
and woodland animal species are colonizing the
newly created habitats (Ashmole and Ashmole,
2009). In years to come, as active management of
Carrifran is reduced and natural processes come
into play, it is hoped that this catchment in the
heart of the Southern Uplands can provide an ex-
emplar of a functioning ecosystem similar to that
which would have been present in the absence
of destructive human intervention during the sec-
ond half of the Holocene.
References
Adair, S. 2005. Carrifran montane scrub restoration.
Report to Scottish Natural Heritage.
Ashmole, P. 2006. The lost mountain woodland of
Scotland and its restoration. Scot. For., 60(1): 9–22.
Ashmole, M. & Ashmole, P. 2009. The Carrifran
Wildwood story: ecological restoration from the
grass roots. Jedburgh, UK, Borders Forest Trust.
Averis, A.B.G. 1998. A Scottish guide identifying appro-
priate new native woodland NVC types based on an
open ground survey. Woodnote No. 18. Perth, UK,
Tayside Native Woodlands.

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Badenoch, C. 1994. Woodland origins and the loss of
native woodland in the Tweed valley. In P. Ashmole,
ed. Restoring Borders woodland, pp. 11–26.
Peebles, UK, Peeblesshire Environment Concern.
Chalmers, H. & Ashmole, P. 2007. Restoring the natural
treeline at Carrifran. Scrubbers’ Bull., 6: 5–10.
Ennos, R.A. 1998. Genetic constraints on native wood-
land restoration. In A.C. Newton & P. Ashmole, eds.
Native woodland restoration in southern Scotland:
principles and practice, pp. 27–33. Jedburgh, UK,
Borders Forest Trust.
Gilbert, D., Horsfield, D. & Thompson, D.B.A., eds.
1997. The ecology and restoration of montane and
subalpine scrub habitats in Scotland. Scottish Natural
Heritage Review No. 83. Edinburgh, UK, Scottish
Natural Heritage.
Hale, S.E., Quine, C.P. & Suárez, J.C. 1998. Climatic
conditions associated with treelines of Scots Pine
and Birch in Highland Scotland. Scot. For., 52(2):
70–76.
Hester, A.J. 1995. Scrub in the Scottish Uplands. Scottish
Natural Heritage Review No. 24. Edinburgh, UK,
Scottish Natural Heritage.
Newton, A. 1998. Carrifran Wildwood Project: a new
restoration initiative in southern Scotland. Restor.
Manage. Notes, 16(2): 212–213.
Newton, A.C. & Ashmole, P., eds. 1998. Native wood-
land restoration in southern Scotland: principles and
practice. Jedburgh, UK, Borders Forest Trust.
Newton, A.C. & Ashmole, P., eds. 1999. Carrifran
Wildwood Project environmental statement.
Jedburgh, UK, Borders Forest Trust.
Peterken, G. 1996. Natural woodland. Cambridge, UK,
Cambridge University Press.
Peterken, G. 1998. Woodland composition and
structure. In A.C. Newton & P. Ashmole, eds. Native
woodland restoration in southern Scotland: princi-
ples and practice, pp. 22–26. Jedburgh, UK, Borders
Forest Trust.
Pyatt, D.G. & Suárez, J.C. 1997. An ecological site
classification for forestry in Great Britain. Forestry
Commission Technical Paper 20. Edinburgh, UK,
Foresty Commission.
Quelch, P. 2002. The ecology and history of aspen
woodlands. In P. Cosgrove & A. Amphlett, eds. The
biodiversity and management of aspen woodlands,
pp. 8–11. Grantown-on-Spey, UK, The Cairngorms
Local Biodiversity Action Plan 2002.
Rodwell, J.S., ed. 1991. British plant communities.
Volume 1. Woodlands and scrub. Cambridge, UK,
Cambridge University Press.
Rodwell, J. & Patterson, G. 1994. Creating new native
woodlands. Forestry Commission Bulletin 112.
London, Her Majesty’s Stationery Office.
Smout, T.C., MacDonald, A.R. & Watson, F. 2005. A
history of the native woodlands of Scotland, 1500–
1920. Edinburgh, UK, Edinburgh University Press.
Tipping, R. 1994. The form and fate of Scotland’s wood-
lands. Proc. Soc. Antiqu. Scot., 124: 1–54.
Tipping, R. 1998. The application of palaeoecology to
native woodland restoration: Carrifran as a case-
study. In A.C. Newton & N.P. Ashmole, eds. Native
woodland restoration in southern Scotland: princi-
ples and practice, pp 9–21. Jedburgh, UK, Borders
Forest Trust.
Wilson, S.McG., Malcolm, D.C. & Rook, D.A. 2000.
Locations of populations of Scottish native trees.
Edinburgh, UK, The Scottish Forestry Trust.

GENET IC CONS IDERATI ONS IN ECO SYSTEM RE STORATI ON USING N ATIVE TRE E SPECIE S
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12.6. The Xingu Seed Network
and mechanized direct
seeding
Eduardo Malta Campos Filho, Rodrigo
G. P. Junqueira, Osvaldo L. de Sousa,
Luciano L. Eichholz, Cassiano C.
Marmet, José Nicola M. N. da Costa,
Bruna D. Ferreira, Heber Q. Alves and
André J. A. Villas-Bôas
Instituto Socioambiental, Xingu, Mato
Grosso, Brazil
The Xingu River flows from the tropical savan-
nah of central Mato Grosso north to the Ama-
zon. With a length of nearly 2000 km, the area
it drains boasts extensive water resources, biodi-
versity and human diversity. The 24 culturally dis-
tinct indigenous peoples of Xingu have conserved
most of the native vegetation in their territories
along the rivers, but settlers who arrived during
the last 40 years have deforested much of the
area on the headwaters of those rivers to create
fields for growing soybean and pastures for cattle
ranching. Although prohibited by the Brazilian
Forest Code, the deforestation of 300 000 hec-
tares of the riparian zone has jeopardized water
quality and regulation of water flow, as well as
harming the health of people who, for centuries,
have depended on the river for water, food and
other services.
Since a meeting in 2004, Instituto Socio-
ambiental (ISA; www.socioambiental.org.br) has
brought together the region’s stakeholders into a
campaign called “Y Ikatu Xingu” (Save the Good
Water of Xingu, in the language of the Kamaiura
Indians) with three principal components: (i) for-
est restoration; (ii) education and c ommunication;
and (iii) regional cooperation between non-gov-
ernmental organizations (NGOs), communities
and policy-makers.22
In 2006, ISA and its partners started education-
al programmes with teachers, students, extension
agents and officials, while farmers were offered
technical assistance, material and financial sup-
port (mostly seeds and fences) for the restoration
of riparian zones. The objectives of the forest
restoration work included protection of water
resources, fruit production, timber production,
carbon sequestration and legal compliance with
the Forest Code, and thus addressed the needs of
a wide range of farmers.
Each restoration project has been aligned with
farmers’ knowledge and ideas to ensure it does
not require major changes from farmers’ exist-
ing practices. Indigenous peoples stated that
trees must be planted by direct seeding, so that
roots develop deeper and the trees can survive
drought. As a result, direct seeding with common
agricultural machinery has been a much better ac-
cepted and effective option than planting seed-
lings. Farmers use the machines and knowledge
used for growing soybeans, maize or grasses or
for spreading fertilizers and limestone to plant
native trees.
Direct seeding also proved to cost less than
planting seedlings (approximately US$2000/ha,
compared with US$5000/ha) and to be more prac-
tical, since seeds are easier to carry and to plant.
To plant one hectare, approximately 60 kilos of
seeds of native trees (200 000 seeds) are mixed
with 100 000 seeds of annual and subperennial
legumes and sand, in a mixture called muvuca.
The legumes help to create a multilayer vegeta-
tion, reducing niches for invasive grasses. Their
root systems can contribute to soil aeration and
decompaction, enhancing water absorption.
Their ability to fix nitrogen and their intense leaf
fall contribute to enhancing nutrient cycling and
soil fertility. Their flowers and fruits attract fauna
22 http://www.yikatuxingu.org.br

THE STATE OF THE WORLD’S FOREST GENETIC RESOURCES – THEMATIC STUDY
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and can be sold. However, if they grow too dense-
ly, they can shade out the tree seedlings, slowing
tree growth. If this occurs, manual or chemical
weeding or thinning will be necessary.
Ninety-one of the native tree species planted
have germinated and survived droughts of up to
six months without irrigation. Tree populations
of between 2500 and 32250 trees/ha have estab-
lished on the reseeded areas. The oldest planted
area is six years old and has a mean density of
7250 trees/ha, greater than the 1666 trees/ha
conventionally used when planting seedlings.
Natural thinning seems to occur over time as a re-
sult of ant herbivory and other mortality factors.
The campaign restored 2565 ha of riparian for-
est at 238 sites. The demand for seeds of indige-
nous tree species rose dramatically and was met by
the creation of the Xingu Seed Network,23 formed
by 300 indigenous people, small landholders and
peasants. Between 2006 and 2012 the Network
produced and sold 71 tonnes of seeds of 214 in-
digenous species and earned almost US$400000
from the environment they have preserved.
23 http://www.sementesdoxingu.org.br
Figure 12.13.
Kaiabi and Yudja people teaching techniques
for seed gathering in the forest at São José do
Xingu, MT, Brazil
Figure 12.12.
Women in the Panará Indigenous Territory
gathering seeds from the Amazon forest,
Guarantã do Norte, MT, Brazil
Figure 12.14.
Preparing muvuca, a mixture of seeds of native
trees, fast-growing legumes and sand for direct
seeding

GENET IC CONS IDERATI ONS IN ECO SYSTEM RE STORATI ON USING N ATIVE TRE E SPECIE S
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Figure 12.17.
Same area shown in Figure 12.16 after three
years, with the trees forming the new canopy,
the last pigeon peas dying, and jack beans and
maize already out of the ecosystem, Canarana,
MT, Brazil
Figure 12.16.
Restoration plot five months after being planted
with muvuca containing pigeon peas, jack beans,
maize and tree seed, Canarana, MT, Brazil
Figure 12.15.
Mechanized direct seeding: using machines designed for spreading fertilizers (left) and sowing
soybeans (right), Mato Grosso State, Brazil

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Network annual meetings discuss ecological
knowledge about indigenous tree species and tech-
niques for collecting and cleaning seeds and set
prices for seed of each species. Seed gatherers are
organized in local groups, and each group is repre-
sented by one of its members. Groups make lists of
what species they can collect each year and in what
quantity. Based on these lists, farmers and NGOs
order what they want to buy. A microcredit fund
allows seed-gatherers to invest in their activities.
Seeds are stored in four storage facilities call ed
“seed houses,” which are equipped with air-con-
ditioning and a dehumidifier. Each seed lot comes
with information regarding who collected it,
where it was collected, type of vegetation, name
of the species, number of parent trees and date
of collection. Lots are assigned a control number
at the seed house and 100 seeds are taken out
for viability tests. Seed quality is checked at least
three times: when seeds are selected in the for-
est, during cleaning and drying and at the seed
houses. Seed technology is still a great challenge
for many species, but a lot has been developed by
the Network, filling information gaps.
Results and learning are disseminated through
field-day demonstrations, practical courses, lec-
tures, workshops, videos, television, magazines,
newspapers, interchange expeditions, school ac-
tivities and demonstration areas.
Focusing on restoration of riparian zones in
a drainage basin has proved successful from a
practical point of view because it addresses wa-
ter conservation and quality issues and thus can
get people engaged. From a wider forest conser-
vation point of view the approach also serves to
connect fragmented patches of forests across the
landscape and thus promotes gene flow and di-
versity at the landscape level.

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13.1. Analogue forestry as an
approach for restoration
and ecosystem production
Carlos Navarro1 and Orlidia Hechavarria
Kindelan2
1 Coordinating Committee, Latin
American Forest Genetic Resources
Network
2 Agroforestry Research Institute, Cuba
“Analogue forestry” is a restoration approach
that aims to develop production- or conserva-
tion-oriented forest systems in degraded forest
areas by drawing on knowledge and observa-
tions about local climax vegetation (Senanayake
and Jack, 1998). The approach is based on the
structure and composition of Sri Lankan and In-
donesian forests and home gardens, which are
small plots of highly productive land located near
houses in traditional rural communities (Senanay-
ake and Jack, 1998; Gamboa and Criollo, 2011).
These gardens maintain a wide diversity of trees,
shrubs and herbs in a manner similar to a forest,
and represent an important part of the tradi-
tional knowledge of farmers.
Forest gardens also serve as a safety net. In
Indonesia rice is the staple food for most people.
Other crops such as cassava (Manihot esculenta
Crantz), taro (Colocasia esculenta (L.) Schott) and
sweet potato (Ipomoea batatas (L.) Lam.) are
grown in forest gardens but rarely consumed by
humans. These crops, which are considered “food
of the poor,” are often used as feed for domes-
tic pigs. However, in times of difficulty, when the
rice harvest fails or when rice stocks are exhaust-
ed just before the next harvest, people use cas-
sava, taro and sweet potato and other crops from
their gardens as emergency foods (Brodbeck,
Hapla and Mitlöhner, 2003). The same is true for
fruit species in Costa Rica, where part of the pro-
duction in gardens is not collected when farm-
ers are busy tending their coffee or cocoa crops.
These fruits are, however, important in times of
crisis. These garden sites, locally called solares,
also provide medicinal plants, basic foods and
fruits, have an ornamental value and benefit the
environment.
Analogue forestry attempts to both increase
biodiversity and improve the well-being of local
communities by creating enhanced and diversi-
fied production systems, valuing people’s own
resources and promoting respect for local values
and traditions. It uses a wide range of crops and
hence reduces risks to farmers of being depend-
ent on a single product. The approach aims to
recreate ecosystems based on the structure and
ecological functions of the original vegetation,
facilitating the spread of many species from the
original forest. It is used to accelerate restoration
in highly degraded areas, especially when there
is little natural gene flow from the surrounding
areas. Analogue forestry also allows use of exotic
species that are similar in structure and function
to native species if the native species has disap-
peared due to fragmentation or habitat loss.
Chapter 13
Approaches including
production objectives

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Analogue forestry builds on 12 guiding prin-
ciples
1. Observe and record: observation and recording
of the structure and physiognomy of the origi-
nal forest and vegetation and the soil condi-
tions of the site to be restored.
2. Understand and evaluate: information on
vegetation, soils, wind directions, water flow,
hedges or artificial fences, etc. is collected both
in the natural forest and the area to be re-
stored and is analysed.
3. Know your land: gather all available informa-
tion and knowledge on the soil and biodiver-
sity conditions of the land.
4. Identify levels of yield: determine potential
crop yields in the target area; this should be
done for all possible products (e.g. cocoa, cof-
fee, vanilla, timber).
5. Map flow and reservoir systems (existing and
potential): prepare maps of water flows, water
tanks and others components of the hydrologi-
cal system.
6. Reduce ratio of external energy in production:
avoid using external inputs when the necessary
inputs can be sourced locally.
7. Be guided by the landscape and the needs of
the neighbours: look at the site as part of a
larger unit to ensure an integrated approach
to site restoration.
8. Follow ecological succession: if the system is
degraded, plant pioneer species to improve soil
conditions for other, more-demanding species.
9. Make use of ecological processes: when the
system has been damaged by erosion or over-
grazing by livestock, start with pioneer species
to improve soil conditions to allow the site to
support a climax ecosystem at a later stage.
10. Value biodiversity: combine as many climax
forest species as possible, although it is some-
times difficult to obtain germplasm of all the
species desired.
11. Respect maturity: mature ecosystems are often
more productive than early-state systems in
terms of biomass production and ecosystem
services, and are especially important for
photo synthesis and carbon and water cycles.
12. Respond creatively: create a system where spe-
cies associations and diversity of components
help to control pests and diseases in an ecosys-
tem approach.
Contribution of analogue forestry to forest
regeneration
Under natural regeneration a forest may take 40
to 60 years to achieve something approaching its
original state, with a return of 60–70 percent of
the original flora and fauna. Analogue forestry
helps to reduce this period by accelerating ecologi-
cal succession. It follows a natural pattern through-
out the restoration process. Starting with pioneer
species, and by promoting ecological succession,
analogue forestry modifies the structure of the
forest canopy and soil quality to allow the site to
support a system of climax vegetation similar to
natural forest of the area. Pioneer species facilitate
restoration by helping improve soil conditions for
more demanding and late-successional species.
In contrast with many other restoration tech-
niques, analogue forestry does not focus on
only woody species, because at least 90–95 per-
cent of the biological diversity of forest plants in
many ecosystems is in non-arboreal components
(shrubs, grasses, epiphytes and lianas; RIFA, 2005).
The idea is to create a system in which various
species, products and plant combinations that can
help controlling pests and diseases are considered
in an integrated manner.
The soil at some restoration sites is not able to
support a climax ecosystem, and needs to be mod-
ified. In newly formed soils (e.g. those that are
the product of volcanic eruptions or sedimentary
soil processes such as flooding), the prevailing
habitat conditions may impede the development
an ecosystem that is similar to the natural veg-
etation of the area. On such sites, the first step is
to study the surrounding pioneer vegetation and
natural forests and describe their physiognomy,
structure, species composition and interactions,
both in terms of density and in their vertical or
horizontal spatial arrangements. The next step is
to replicate this vegetation in the new areas to
assist natural regeneration.

GENET IC CONS IDERATI ONS IN ECO SYSTEM RE STORATI ON USING N ATIVE TRE E SPECIE S
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Genetic diversity conservation and analogue
forestry can be combined to produce a bet-
ter environment and more resilient ecosystems.
Analogue forestry can contribute to the conserva-
tion of genetic diversity by:
• providing space for diversity conservation;
• establishing spatial arrangements that
encourage gene flow and increase
connectivity between patches of forests;
• preserving ecological relationships among
species; and
• creating demonstration plots for
environmental education.
In turn, analogue forestry benefits from genetic
diversity, for example, through increased resist-
ance to pests and diseases, better local adapta-
tion, adaptation to climate change and diversity
of products.
Methodology
During the initial stage of a restoration project,
the natural forest surrounding the area to be re-
stored must be studied to determine the ecosys-
tem to be re-established on the degraded area.
The structure of the ecosystem is important to
maintain the ecological relationships between
pollinators, herbivores and nutrient cycling. To
begin with, the number of canopy layers or strata
should be determined, and woody plants in each
of these identified. The height, coverage, consist-
ency and leaf size of the prominent or dominant
species in each stratum should be determined.
Next, the different growth forms (e.g. climbers,
small palm species or herbs) are described, in-
cluding their average height or height range and
coverage. Information must be gathered on the
ecological roles and human uses of the species,
especially where the restored system will provide
staple or cash crops. Species are selected for the
restoration of the various vegetation layers based
on their growth height, their uses and the charac-
teristics of their seeds, among other factors.
One tool that differentiates analogue forestry
from other restoration techniques is the use of
a physiognomic formula. It allows visualizing a
model for the restoration process as a codified
description of the structure of the tree and non-
tree components of the vegetation found in the
area of interest. In the formula, each stratum is
described by a specific code, followed by a de-
scription of special growth categories. In highly
diverse landscapes, the physiognomic formulas
of the vegetation are complex, as they include
all strata and life forms of the forest vegetation.
Some of the criteria upon which such assessments
are based are soil quality, biodiversity and vegeta-
tion structure.
The application of this approach is described
in more detail in the following case studies from
Costa Rica and Cuba.
References
Brodbeck, F., Hapla, F. & Mitlöhner, R. 2003.
Traditional forest gardens as “safety net” for rural
households in Central Sulawesi, Indonesia. Paper
presented at The International Conference on Rural
Livelihoods, Forests and Biodiversity, 19–23 May
2003, Bonn, Germany (available at http://www.cifor.
org/publications/corporate/cd-roms/bonn-proc/pdfs/
papers/T1_FINAL_Brodbeck.pdf).
RIFA (Red Internacional de Forestería Análoga).
2005. Manual de forestería. 2da edición. Quito,
Ecuador, RIFA. 21 pp.
Senanayake, R. & Jack, J. 1998. Analog forestry: an
introduction. Clayton, Victoria, Australia, Monash
University.

THE STATE OF THE WORLD’S FOREST GENETIC RESOURCES – THEMATIC STUDY
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13.1.1. Restoring forest for food
and vanilla production under
Erythrina and Gliricidia trees in
Costa Rica using the analogue
forestry method
Carlos Navarro
Coordinating Committee, Latin
American Forest Genetic Resources
Network
An ecological assessment of the forest vegeta-
tion of Los Espaveles primary forests at the Cen-
tro Agronómico Tropical de Investigación y Ense-
ñanza (CATIE), Turrialba, Costa Rica, identified the
following strata, starting for the tallest trees:
• First (topmost) layer (V8): trees of more
than 35 m tall; evergreen broadleaf species;
sparse canopy cover of 6–25 percent of the
forest area.
• Second layer (V7): woody plants; evergreen
broadleaf species, height 20–35 m; patchy
canopy cover (25–50 percent of forest area).
• Third layer (V6): woody plants; evergreen
broadleaf species, height 10–20 m; sparse
canopy cover.
• Fourth layer (V5): woody plants; evergreen
broadleaf species, height 5–10 m; sparse cover.
• Fifth layer (V4): evergreen broadleaf species,
height 2–5 m; patchy cover (1–6 percent).
• Sixth layer (V3): evergreen broadleaf species,
height 0.5–2 m; sparse cover.
• Seventh layer (V2): evergreen broadleaf
species, height 0.1–0.5 m; patchy cover
(1–6 percent).
• Eighth (lowermost) layer (V1): seedlings of
evergreen broadleaf species, length <2.5 cm,
height <0.1 m; very low cover (less than
1 percent).
In addition, the following growth forms were
identified:
• Climbers in canopy layers above 35 m (T8);
almost absent
• Epiphytes in 10–20-m layer (E6); almost
absent
• Palm species in 5–10-m layer (P5): sparse
cover
• Bananas in 2–5-m layer (R4); sparse cover
• Ferns in 0.1–0.5-m layer (F2); sporadic cover
• Herbs in 0.1–0.5 m layer (H2); sporadic cover
• Lichens and mosses in <0.1 m layer (L1);
sporadic cover.
The corresponding formula of this forest type is:
V8r, V7p, V6r, V5r, V4e, V3r, V2e, V1a; T8a; E6a;
P5r; R4r; F2e; H2e; L1e
where the first letter (capital) indicates the
growth form (V=evergreen broadleaf), the num-
ber indicates the canopy layer and the follow-
ing lowercase letter indicates extent of cover
(r=rare [6–25 percent], p=patchy [25–50 per-
cent], e=sporadic [1–6 percent] and a=almost
absent [<1 percent]). For non-woody plants the
code also includes information on plant bio-
logy (E=epiphytes, P=palm trees, R=herbaceous
[banana], F=ferns, H=herbs and L=lichens and
mosses).
The physiognomic formula of the area to be
restored is then constructed to allow comparison
with the formula of the original forest followed
by restoration of the layers that are missing or
under represented:
• V6: evergreen woody plants, height
10–20 m; patchy cover
• V5: evergreen broadleaf plants with hard
leaves (d), height 5–10 m; sparse cover
• V4: evergreen plants with hard leaves (d),
height 2–5 m; scarce cover
• Special growth forms:
• Palm species, height 0.5 –2 m; patchy cover (P3)
• Epiphytes, height 5–20 m with soft leaves (s);
almost absent (E6)
• Graminoids, height <0.1 m with soft leaves
(s); continuous cover (c) (G1)

GENET IC CONS IDERATI ONS IN ECO SYSTEM RE STORATI ON USING N ATIVE TRE E SPECIE S
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The resulting physiognomic formula of the area
to be restored is
• V6esm, V5rdm, V4rdg; P3e; G1csn; E6-5sma
where m, g and n indicate leaf size.
As can be seen, the formula in the restoration area
has fewer elements. The differences between the
formulae for the natural vegetation and the area
to be restored guide the restoration activities.
In consequence, it is important to have detailed
knowledge of species that can be used in the area,
their potential ecological and economic uses and
their stratum in the analogue layer of the forest.
CATIE has developed a demonstration site in
Turrialba, Costa Rica, where plants for food produc-
tion and vanilla were planted in association with
Erythrina and Gliricidia trees. The goal of the pro-
ject was to develop a productive site with a diverse
group of species that: (i) have multiple uses; (ii) pro-
vide stability through the availability of products
year round; and (iii) allow farmers to respond more
flexibly to market fluctuations. An additional aim
of the project was to simultaneously conserve di-
versity, provide wildlife with food sources and sup-
port a diversified production system.
In addition to plants that established through
natural regeneration, the project planted other
native and exotic species with: (i)medicinal val-
ue; (ii) importance for household consumption;
(iii)commercialization potential; and (iv) poten-
tial to generate ecosystem services. Many of the
species used provide multiple services including
fruit, wood and erosion control. These include
Erythrina sp., Gliricidia sepium (Jacq.) Kunth ex
Walp., soursop (Annona sp.), Caimito (Pouteria
caimito Radlk.), Guaba (Inga sp.), guava (Psidium
sp.), Zapote (Pouteria sp.) and peach palm (Bactris
gasipaes Kunth). Species planted for wood pro-
duction, either for commercial timber or for use
on farm, included mahogany (Swietenia macro-
phylla King), cedar (Cedrela odorata L.), nance
(Byrsonima crassifolia (L.) Kunth) and two varieties
of naranjilla (Solanum quitoense Lam.), among
others. Medicinal plants included mint (Mentha
sp.), rue (Ruta sp.) and oregano (Origanum sp.);
several of the other species listed above also have
medicinal properties.
13.1.2. Restoration of ecosystems on
saline soils in Eastern Cuba
using the analogue forestry
method
Orlidia Hechavarria Kindelan
Agroforestry Research Institute, Cuba
At present, large areas of agricultural land in
Cuba are degraded by salinization as a conse-
quence of poor soil management associated with
sugar-cane production. This degradation could be
reversed through reforestation and conservation
measures. Accumulating humus neutralizes the
toxic effects of salinization and vegetation cover
helps to maintain moisture in the top soil, which
in turn impedes the concentration and crystalliza-
tion of salts.
Study area
Numerous new farms were established in the vi-
cinities of the communities of Cecilia, Sombrilla
and Paraguay in 2000 as part of the national
forestry farms plan which aimed at encouraging
farm families to live in and reforest degraded
wooded areas. The three communities are com-
pletely dependent on the local sugar mill for
income, and relationships between them and
the new farmer communities are not very good.
There is no original forest left in the area, and the
landscape is fragmented by cultivation of sugar
cane. The farmers live on their farms with their
children, their average age being between 33 and
35 years. The majority of them work on the farms
as forestry workers. Women mostly take care of
the household and participate in agricultural

THE STATE OF THE WORLD’S FOREST GENETIC RESOURCES – THEMATIC STUDY
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170
work. The study reported here took place near
the community of Paraguay.
The original vegetation of the study area was
mainly composed of xerophytic evergreen shrubs
and thorny leguminous trees. On the degraded
areas this vegetation has replaced by degraded
secondary plant communities (Figure 13.1).
Mean annual precipitation fluctuates around
600 mm, with rainfall being concentrated from
May to October. Mean annual temperature is ap-
proximately 26°C. Soil in the area is Fluvisol of
moderate depth (20–50 cm), fairly saline, with al-
most flat topography and little erosion (Sánchez
et al., 2008).
Methods
Approximately 75 soil samples were collected
at depths of 0–20, 20–40, 40–60, 60–80 and
80–100 cm to assess the relation between salinity
and the occurrence of indicator plants. At a depth
of 0–20 cm, the soil salinity ranged from very
low (ECe 1.08 dS/m) to very high (ECe 5.28 dS/m),
with most samples being highly saline. Deeper
soil layers were evaluated as excessively saline
(ECe 4.00–13.01 dS/m).The aquifer was also saline.
The drainage canal was clogged, and the soil con-
tained little organic matter. As a result of these
unfavourable soil conditions the survival of tree
species planted by local farmers was low.
In an effort to remedy this situation, the Cuban
Agroforestry Research Institute, in coordination
with the International Analog Forestry Network
and the Falls Brook Centre, Canada, commenced
a cooperative project to restore the vegetation
of the xerophytic corridor of Guantanamo Valley
using the analogue forestry approach. The pro-
ject commenced in 2008 with the elaboration
of maps of the farms and landscapes. This initial
exercise showed a total farmed area of 338ha.
Species already occurring on site and other wild
species capable of reducing the impact of deg-
radation were identified from surveys, biblio-
graphic information, interviews with community
members and experts and results from previous
work and selected for planting in the initial res-
toration phase. Once the ecosystem showed signs
of stabilization, other exotic and native species
were introduced.
Medium-growth tree species already present
on site before initiation of the project included
Casuarina equisetifolia Forst, wild tamarind
(Lysiloma latisiliquum (L.) Benth) and Caesalpinea
violaceae (Mill.) Standl. During the first phases of
the restoration activities the fast-growing exot-
ics Moringa oleifera L., Prosopis juliflora (Sw.) DC
andneem (Azadirachta indica L.) were planted to
improve soil conditions before introducing other
native species, since they are well-adapted to sa-
line soils and increase the soil organic- matter con-
tent through the production of leaf litter. Native
species planted included the medium-growth
species soplillo (Lysiloma latisiliquum), the slow-
growth species Guaiacum officinale L. and several
other smaller species such as Colubrina arbores-
cens (Mill.) Sarg. and sea grape (Coccoloba uvifera
L.). Exotic fruit species, including peach (Prunus
persica (L.) Batsch), mango (Mangifera indica L.),
and coconut (Cocos nucifera L.), were also includ-
ed to provide food for local farmers. All species
were incorporated gradually to create a vegeta-
tion structure similar to the original vegetation
but that also met the food requirements of local
communities.
Figure 13.1.
Degraded farmlands near the community of
Paraguay before restoration work had begun (2008)

GENET IC CONS IDERATI ONS IN ECO SYSTEM RE STORATI ON USING N ATIVE TRE E SPECIE S
171
Results and discussion
Thirteen species (seven native species, three fast-
growing exotic species and three exotic fruit
trees) were planted on the farms in May and June
2008, taking advantage of existing soil humidity.
Two kilograms of organic matter made on farm
were deposited in each planting hole and mixed
with soil. The species were planted as stakes aver-
aging 35–45 cm in height.
After planting, various agro-ecological tech-
niques were applied, including weeding and
mulching the soil with weeding residue to reduce
evaporation, maintain soil moisture and increase
the organic matter around the tree stalks. These
measures were taken to promote the develop-
ment of the root system of plants and their up-
take of nutrients.
Survival of planted trees was assessed 36
months after planting. All indigenous species
showed a moderate growth after 36 months
(Table 13.1; Figure 13.2). These species are char-
acterized by slow to medium growth rates (Bisse,
1988), particularly under the extremely saline soil
conditions of the restoration site. Under these
conditions it is of great importance to incorporate
ground cover species to protect the soil from inso-
lation and promote maintenance of soil moisture.
Swietenia mahagoni showed the greatest growth
TABLE 13.1.
Characteristics and survival of indigenous species 36 months after planting on the research plots
on the farms
Species Common name
in Cuba Family Growth
characteristics Functions at site Survival assessment
Mean
height
(m)
Survival
rate
(%)
Guaiacum officinale
Guayacan negro Zygophyllaceae Slow-growing
stabilizing species
Erosion control, animal
shelter, wood†
1.23 83
Swietenia mahagoni
Caoba antillana Meliaceae Stabilizing species* Erosion control, animal
shelter, wood*
3.03 60
Lysiloma latisiliquum
Soplillo Fabaceae/
Leguminosae
Colonizing species* Soil protection†2.50 57
Conocarpus erectus
Yana Combretaceae Slow-growth
species*
Soil improvement†2.40 70
Cordia alba
Uvita Boraginaceae Colonizing species* Soil improvement and
protection†
–‡–‡
Coccoloba uvifera
Uvacaleta Polygonaceae Stabilizing species* Erosion control, fodder† 1.90 17
*Perez and Velázquez (2008).
† Survey findings, 2010.
‡ Not evaluated because the species is a soil creeper.
Figure 13.2.
The same area shown in Figure 13.1 in 2014,
five years after planting

THE STATE OF THE WORLD’S FOREST GENETIC RESOURCES – THEMATIC STUDY
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172
(3.03m on average) while Colubrina ferruginosa
had the highest survival rate (92 percent of all
planted individuals) (Table 13.1). Conocarpus
erectus performed well in terms of both survival
rate and growth. It also acts as a nursery plant for
soplillo (Lysiloma latisiliquum), which in turn cre-
ates favourable conditions in the lower stratum
for the establishment of other species. The high
survival rates of soplillo (L. latisiliquum), yana
(C. erectus) and guayacan (G. officinale), all na-
tive species in the area, indicates that even under
severe heat and water stress, certain xerophytic
plant species are still able to survive.
Results obtained so far show that it is impor-
tant to carefully select the species that match the
environmental characteristics of the degraded
site. Appropriate soil preparation is essential to
ensure high survival rates, including very deep
plantation holes, application of organic matter
and irrigation during establishment. In the 1980s
specialists of the Cuban Agroforestry Research
Institute established comparable experiments
on saline soils with some species used in this
study, but without application of organic matter.
Rapidly declining survival rates were observed
over time (Figure13.3), which demonstrates the
severe environmental pressure plants are under in
all these extreme conditions.
Conclusions
The incorporation of a range of native and exotic
species increased the diversity of species on the
areas to be restored, and created conditions that
facilitate the establishment of other species that
meet the economic and social needs of local com-
munities. Three years after initiating the study,
the results are still preliminary, but show the
usefulness of analogue forestry techniques for
achieving gradual reforestation and restoration.
References
Bisse, J. 1981. Árboles de Cuba. C. Habana, Cuba,
Editorial Científico-técnica.
Gamboa, L. & Criollo, M.C. 2011. Forestería análoga y
su rol en la recuperación de ecosistemas y el cambio
climáticoía Cristina. LEISA, 27(2): 8–12.
Figure 13.3.
Survival rate (percent) of Lysiloma havanensis, Guaiacum officinale and Coccoloba uvifera in saline
soils without organic matter added, obtained from experiments conducted in the 1980s

GENET IC CONS IDERATI ONS IN ECO SYSTEM RE STORATI ON USING N ATIVE TRE E SPECIE S
173
Brodbeck, F., Hapla, F. & Mitlöhner, R. 2003.
Traditional forest gardens as “safety net” for rural
households in Central Sulawesi, Indonesia. Paper
presented at The International Conference on Rural
Livelihoods, Forests and Biodiversity, 19–23 May
2003, Bonn, Germany (available at http://www.cifor.
org/publications/corporate/cd-roms/bonn-proc/pdfs/
papers/T1_FINAL_Brodbeck.pdf).
Pérez, E. & Velázquez, F.A. 2008. Habilidades competi-
tivas y adaptativas de algunas especies de impor-
tancia forestal existentes en Camagüey. Rev. Forest.
Baracoa, 21(1): 113–123.
Sánchez, R. Milá., F., Planas., J. Sánchez., Cintra, I.M.
& Lugo, D. 2008. Informe de suelos realizado a
fincas forestales de Paraguay, Guantánamo, Cuba.
Guantánamo, Cuba, Centro provincial de suelos.
13.2. Post-establishment
enrichment of restoration
plots with timber and non-
timber species
David Lamb
Centre for Mined Land Rehabilitation,
University of Queensland, Australia
The term “enrichment planting” is used here to
describe the situation in which the species in an
existing natural forest or plantation are supple-
mented by adding additional species. The most
common reason for undertaking enrichment
planting is to increase the proportion of trees
that have a commercial value. The technique has
been used in logged-over natural forests where
natural regeneration has been insufficient and
the purpose has been to increase the density of
commercially attractive timber species (or, where
the species is already present, to increase the
density of these commercially attractive trees).
An outline of the silvicultural issues is given by
Baur (1964), Lamb (1969), Appanah and Weinland
(1993), and Dawkins and Philip (1998). Enrich-
ment planting has also been carried out in natural
forests in the tropics to create what some have
called agroforests (Michon, 2005). In these cases
enrichment with food, medicinal or resin-produc-
ing species has been done during the fallow stage
of shifting cultivation to improve the supply of
these products to the forest owner. Such forests
are found in many parts of the tropical world
and often cover large areas (Clarke and Thaman,
1993; Michon, 2005). Finally, enrichment plant-
ing has been used to modify the composition of
some planted forests although this has probably
not been as widely implemented as in the case of
natural forests.
Enriching planted forests
There are two situations in which the enrich-
ment of planted forests can be attractive. One
is where the objective is to increase the commer-
cial value of the forest, while the other is where
the intent is to increase the conservation value
of the forest.
The commercial value can be enhanced in sev-
eral ways. The first is where a species producing a
non-timber forest product (NTFP) can be grown
in the shade of existing plantation trees and pro-
duce a commercial crop in a shorter time than it
would take for the trees to reach a harvestable
size. This can make tree-growing attractive to
landholders because of the earlier-than-normal
cash flow. Examples of this are the growing of rat-
tans, medicinal plants or food crops in plantation
understories (Lamb, 2011).
A second form of commercial enrichment is
when a plantation monoculture is enriched with
timber trees rather than NTFP species. The ap-
proach has been used when the preferred timber
species cannot tolerate the present site condi-
tions and the initial plantation trees are used to
modify these conditions to facilitate the estab-
lishment of the preferred species. This situation
can arise when native species are being estab-
lished at degraded sites. This process might be
thought of as being less a case of enrichment and
more a case of using the initial plantation as a

THE STATE OF THE WORLD’S FOREST GENETIC RESOURCES – THEMATIC STUDY
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174
temporary nurse crop. However, the silvicultural
issues confronting the manager are similar. An ex-
ample of this approach is where the nurse crop
or facilitator species is established to shade out
weeds, improve the soil fertility or provide some
early shade for the target species (e.g. McNamara
et al., 2006). In other cases the nurse crop is used
to create an environment that reduces insect
damage to the target species (Keenan, Lamb and
Sexton, 1995). In all of these situations the nurse
trees in the original plantation are eventually re-
moved leaving the now-established commercially
attractive species to grow.
Enrichment of a planted forest to increase its
conservation value is beginning to be practised
in situations in which changes in environmental
attitudes (or in timber markets) mean that some
timber plantations are being taken out of the
production forest estate and being added to the
protection or conservation forest area (Knoke
et al., 2008; Lamb, 2011). Timber plantations
close to urban areas or in locations that are stra-
tegically important for biodiversity conservation
(e.g.adjacent to national parks) may be destined
for enrichment of this type. A summary of some
of the situations in which enrichment might be
practised is shown in Table 13.2.
Methods for undertaking enrichment
in natural forests
Some of the principles for undertaking enrich-
ment have been developed from silvicultural
research conducted when enriching natural
forests. In all cases the key task is to create
conditions on the forest floor in which the in-
troduced seedlings can establish and grow (or,
where species are added as seed, can germinate
and grow). This usually means manipulating the
canopy to improve the light environment. Those
enriching disturbed natural forests have found
that the best time to undertake enrichment
is immediately after logging (or, in the case of
shifting cultivation, after the cropping period
is mostly complete) when the canopy openings
are greatest. But even under these ideal condi-
tions it is usually necessary to remove compet-
ing vegetation around the planted seedlings at
least several times a year for one or two years to
allow the planted seedlings to thrive. If circum-
stances dictate that the process of enrichment
is delayed then it is best carried out by cutting
strips through the forest and planting seedlings
along these strips. Alternatively, it can be done
by poisoning or ring-barking a group of trees to
create large canopy gaps. In both cases the ob-
jective is to manipulate the forest canopy to al-
low sufficient light to reach to forest floor. Some
prescriptions are outlined in Box 13.1.
Some experimentation is usually needed. Large
canopy gaps may help newly planted seedlings
become established but they will also encourage
the growth of competing weed species. A balance
has to be struck between opening the canopy
enough to promote the growth of the introduced
seedlings but not so much that the costs of weed
control becomes excessive.
TABLE 13.2.
The types of species that might be used in enrichment and the types of forests to which these might
be added
Types of species used to enrich an existing forest Types of forest that could be enriched
NTFP species Monoculture or mixed-species timber plantation
Timber trees or NTFP species Monoculture plantation of nurse trees
Native species with high conservation value Monoculture plantation
Timber trees or NTFP species Disturbed (logging, shifting cultivation, etc.) natural forest
Native species with high conservation value Disturbed (logging, shifting cultivation, etc.) natural forest
NTFP = non-timber forest product

GENET IC CONS IDERATI ONS IN ECO SYSTEM RE STORATI ON USING N ATIVE TRE E SPECIE S
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Methods for undertaking enrichment in
planted forests
Similar constraints apply when enriching planted
forests. However, the methods used depend on
whether the objective is to add diversity to a
plantation or to increase the commercial value
of the plantation. Where the purpose is to add
diversity the emphasis is likely to be more on en-
suring seedling survival than on maximizing seed-
ling growth. This means that the primary task is
likely to be one of creating conditions allowing
the newly planted seedlings to simply survive
rather than to manipulate the canopy cover to
enhance their subsequent growth. The most com-
mon approach is to create canopy gaps and plant
seedlings of the desired species in these gaps. Just
how large the canopy openings must be will de-
pend on the height of the canopy trees, the solar
elevation (i.e. on latitude) and on the shade tol-
erance of the seedlings being planted. Dawkins
and Philip (1998) note that wildlife herbivores
can sometimes damage newly planted seedlings
in existing tropical forests, and Bergquist, Lof and
Orlander (2009) and Beguin, Pothier and Prévost
(2009) note similar results in forests in Sweden
and Canada, respectively. Control of herbivores
may be a necessary management activity in many
enrichment programmes. This has been done us-
ing fences although this method can be expen-
sive.
Not all enrichment need be carried out by plant-
ing seedlings and there may be scope for directly
sowing seeds on the forest floor of a plantation.
Direct sowing has the significant advantage that
it can be cheaper to undertake than raising seed-
lings in a nursery and planting these in the forest.
On the other hand, simply broadcasting seeds on
the floor of a plantation can be an inefficient use
of seeds, since many trials have shown the success
rates with direct seeding can be low (Hau, 1997;
Engel and Parrotta, 2001; Doust, Erskine and
Lamb, 2008). This means it may only be a suitable
technique to use with species that are easily avail-
able and cheap to collect.
When the species in planted forests are be-
ing supplemented for commercial reasons there
is usually more emphasis given to maximizing
growth as well as survival rates. Again, gaps can
• The species used must be capable of fast
growth, meaning that most will be light-
demanding.
• Seedlings should have well-established roots,
meaning that container-grown seedlings
are preferable to bare-rooted seedlings or
wildlings.
• Species should have a low crown ratio (ratio of
crown diameter to stem diameter).
• These species should be self-pruning and have
good form.
• Planting lines should be oriented in an east–
west direction and be separated by a distance
about the same as the crown diameter of the
species when mature (e.g. around 10–15 m).
• Seedlings should be planted more closely along
these lines (i.e. <10 m) to allow for deaths and
perhaps thinning.
• All overstorey competition should be removed
before planting to avoid damaging young
seedlings.
• Weeds along the planting line should be
removed at least three times in the first year in
a strip about 2 m wide.
• The technique will fail if seedlings are
susceptible to grazing by wildlife.
• The regrowth between the planting lines should
not be flammable.
Source: Based on Lamb (1969), Appanah and Weinland (1993),
and Dawkins and Philip (1998).
Box 13.1.
Principles for enrichment planting in disturbed natural forest

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be created to increase the amount of light on the
forest floor. These gaps must be sufficient to al-
low the growth of the planted seedlings but not
so great as to encourage weed growth. In many
cases gaps are created by removing every second
or third row of planted trees. The underplanted
seedlings used to enrich the plantation are then
planted along this line.
Care is needed to ensure that this degree of
canopy cover remains appropriate. For example,
seedlings may benefit from a certain amount of
shade when they are young but require rather
more light once they are established. In the ab-
sence of extra light their growth may decline and
stagnate. This means there can be a trade off
between the beneficial advantages of the cover
crop, such as providing early shade when the
seedlings are sensitive to full sunlight, and then
hastening shoot growth rates once seedlings have
passed beyond this stage (Keenan, Lamb and
Sexton, 1995; McNamara et al., 2006).
Case study: enrichment of monocultures
in western Europe
In recent years there has been increased interest
in adding additional species, mostly broadleaves,
to the simple monoculture forests found in many
parts of western Europe (Hansen and Spiecker,
2005; Harmer, Thompson and Humphrey, 2005).
Many of these monocultures involve exotic con-
fers such as Sitka or Norway spruce. Silvicultural-
ists have been motivated to add additional spe-
cies because enrichment can improve biodiversity
conservation and also increase resilience, which
means the stands are better protected against
biotic and abiotic changes. Enrichment can be
carried out by planting seedlings of the desired
species or by allowing natural regeneration of
these species to occur. In both cases better sur-
vival and growth is dependent on improving the
light environment for the new seedlings. The way
the light environment is manipulated depends
on the nature of the existing stand but involves
creating gaps in the existing forest canopy or re-
moving rows of trees. In some cases these have
been commercial fellings, but in others the trees
have been felled to waste. Clear felling is usually
avoided to prevent sites from being over-run with
weeds. Herbivores have been excluded by fenc-
ing or by culling. Targets for conversion are usu-
ally unstated but may involve reducing the exist-
ing dominants to less than 20 percent of the tree
density.
Case study: enrichment of logged-over
natural forest in Sabah, Malaysia
Large areas of natural forest in Sabah have been
degraded by intensive logging and fire. The in-
tensity of these disturbances limited the capacity
of the forests to recover naturally. This means it
will be some years before another timber harvest
will be possible. Enrichment has been carried out
to accelerate this recovery process (Yap, 2011).
Ideally this should have taken place immediately
after the disturbances but the extent of the area
needing treatment meant that many areas could
not be treated. Instead, these have been occu-
pied by natural regeneration of mainly pioneer
species, especially Macaranga sp. Enrichment in-
volved girdling or ring-barking these trees and
planting seedlings of more than 100 commer-
cially valuable native timber tree species in rows

GENET IC CONS IDERATI ONS IN ECO SYSTEM RE STORATI ON USING N ATIVE TRE E SPECIE S
177
or in gaps, depending on the presence and dis-
tribution of native timber trees. Where seedlings
are planted in rows, the rows have been about
10 m apart with the seedlings spaced 3–5 m apart
along the rows. When seedlings are planted in
gaps the aim is to have a cluster of three seedlings
in a 10 × 10 m plot. Some tending is carried out to
enable seedlings to become established and the
planting areas are surveyed after three months
to ensure survival is greater than 90 percent (re-
planting is done if survival is less than this). In the
first two years around 2–3 tendings are usually
done with less thereafter. Tending can continue
for up to ten years depending on need. Across
Sabah about 45 000 ha have been enriched using
similar methods.
Monitoring
What constitutes success? Anyone undertaking
enrichment should have a clear idea of what is
an acceptable outcome and when they should in-
tervene to ensure enrichment is succeeding. The
most obvious measure of success is that a high
proportion of the planted seedlings have sur-
vived. But equally important might be that these
seedlings are uniformly distributed across the for-
est area and not clustered in a single small area.
Rapid seedling growth may – or may not – be an
important additional factor. When enrichment is
for commercial purposes the growth rates of the
newly planted trees will have to be monitored so
that the canopy cover can be adjusted when nec-
essary. But for enrichment planting programmes
aimed at improving forest biodiversity, perhaps
the most important indicator of success is that
the species used to enrich the site have been able
to flower, fruit and regenerate themselves on
the forest floor. Once this occurs, a simple even-
aged plantation begins to be transformed, via
enrichment, into a self-sustaining uneven-aged
forest. Any monitoring programme needs to be
designed as a series of questions such that the
answers will give unequivocal guidance to the
manager (e.g. Have at least 80 percent seedlings
survived? Are these seedlings uniformly distrib-
uted across the site? Have the species used in the
enrichment programme flowered and produced
seed?).
Conclusion
Enrichment planting can be used to add species to
existing natural forests and plantations. Much of
what we know about the silvicultural techniques
needed to undertake enrichment comes from
work carried out in degraded natural forests. The
primary task is to create canopy openings suffi-
cient to allow light to reach the forest floor and
promote seedling growth. When enrichment is
being carried out to improve the conservation
value of the forest the amount of light needed
is that sufficient to allow the seedlings to survive.
But when enrichment is being carried out to gen-
erate a commercial benefit it may be necessary to
continue to monitor and manipulate this light en-
vironment to ensure rapid seedling growth.
References
Appanah, S. & Weinland, G. 1993. Planting quality tim-
ber trees in peninsular Malaysia: a review. Kepong,
Malaysia, Forest Research Institute.
Baur, G.N. 1964. The ecological basis of rainforest man-
agement. Sydney, Australia, Forestry Commission of
New South Wales.
Beguin, J., Pothier, D. & Prévost, M. 2009. Can the
impact of deer browsing on tree regeneration be
mitigated by shelterwood cutting and strip clearcut-
ting? Forest Ecol. Manag., 257: 38–45.
Bergquist, J., Lof, M. & Orlander, G. 2009. Effects of
roe deer browsing and site preparation on perfor-
mance of planted broadleaved and conifer seedlings
when using temporary fences. Scand. J. Forest Res.,
24: 308–317.
Clarke, W.C. & Thaman, R. 1993. Agroforestry in the
Pacific islands: systems for sustainability. Tokyo,
United Nations University Press.
Dawkins, H.C. & Philip, M.S. 1998. Tropical forest silvi-
culture and management: a history of success and
failure. Wallingford, UK, CAB International.

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Doust, S.J., Erskine, P.D. & Lamb, D. 2008. Restoring
rainforest species by direct seeding: tree seedling es-
tablishment and growth performance on degraded
land in the wet tropics of Australia. Forest Ecol.
Manag., 256: 1178–1188.
Engel, V.L. & Parrotta, J.A. 2001. An evaluation of
direct seeding for reforestation of degraded lands in
central Sao Paulo state, Brazil. Forest Ecol. Manag.,
152(1–3): 169–181.
Hansen, J. & Spiecker, H. 2005. Conversion of Norway
spruce (Picea abies [L.] Karst.) forests in Europe. In
J.A. Stanturf & P. Madsen, eds. Restoration of boreal
and temperate forests, pp. 339–347. Boca Raton, FL,
USA, CRC Press.
Harmer, R., Thompson, R. & Humphrey, J. 2005. Great
Britain – conifers to broadleaves. In J.A. Stanturf &
P. Madsen, eds. Restoration of boreal and temper-
ate forests, pp. 319–338. Boca Raton, FL, USA, CRC
Press.
Hau, C.H. 1997. Tree seed predation in degraded
hillsides in Hong Kong. Forest Ecol. Manag., 99:
215–221.
Keenan, R., Lamb, D. & Sexton, G. 1995. Experience
with mixed species rainforest tree plantations in
North Queensland. Commonw. Forest. Rev., 74:
315–321.
Knoke, T., Ammer, C., Stimm, B. & Mosandl, R. 2008.
Admixing broadleaved to coniferous tree species: a
review on yield, ecological stability and economics.
Eur. J. Forest Res., 127: 89–101.
Lamb, A.F.A. 1969. Artifical regeneration within the
humid tropical lowland forest. Commonw. Forest.
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Lamb, D. 2011. Regreening the bare hills: tropical forest
restoration in the Asia-Pacific region. Dordrecht, The
Netherlands, Springer.
McNamara, S., Tinh, D.V., Erskine, P.D., Lamb,
D., Yates, D. & Brown, S. 2006. Rehabilitating
degraded forest land in central Vietnam with mixed
native species plantings. Forest Ecol. Manag., 233:
358–365.
Michon, G. 2005. Domesticating forests: how farmers
manage forest resources. Bogor, Indonesia, Center
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scale forest rehabilitation project in Sabah, Malaysia.
Paper presented at the International Symposium on
Rainforest Rehabilitation and Restoration, 25–29 July
2011, Maliau Basin Study Center, Sabah, Malaysia.
13.3. Enrichment planting
using native species
(Dipterocarpaceae) with
local farmers in rubber
smallholdings in Sumatra,
Indonesia
Hesti L. Tata,1,2 Ratna Akiefnawati2 and
Meine van Noordwijk2
1 Forest Research and Development
Agency, Indonesia
2 World Agroforestry Centre (ICRAF),
South East Asia Regional Office,
Indonesia
Indonesia has the world’s third largest area of
tropical forest. An estimated 50 percent of the
country’s total land area still has forest cover
(FAO, 2005). Natural forests in the lowland of
Sumatra and Borneo are dominated by Diptero-
carpaceae, which is one of the most important
families for good-quality timber. Some species
provide non-timber forest products, such as dam-
mar resin, camphor and illepe nuts. The family

GENET IC CONS IDERATI ONS IN ECO SYSTEM RE STORATI ON USING N ATIVE TRE E SPECIE S
179
consists of 16 genera and is widely distributed
from Africa (Congo, Côte d’Ivoire, Ghana, Guinea,
Madagascar and the Seychelles), to Asia (the An-
daman Islands, India, Indonesia, Malaysia, Nepal,
Pakistan, Sri Lanka, Papua New Guinea and the
Philippines) and South America (Colombia, Ecua-
dor and Venezuela) (Ashton, 1982; Maury-Lechon
and Curtet, 2005). In Sumatra alone, 106 species
of Dipterocarpaceae have been recorded. Con-
struction timbers derived from Dipterocarpaceae
include red meranti, white meranti, yellow mer-
anti and bangkirai (Ashton, 1982).
The nature of forest in Indonesia is rapidly
changing, even if cover is being maintained.
Indonesia has become the global leader in car-
bon-dioxide emissions from land-use change as
a result of the rapid loss of forest biomass and
destruction of peatlands (Archard et al., 2002; de
Fries et al., 2002). The overall loss of forest cover
in Indonesia from 2003 to 2006 was 1.2 million ha/
year (MoFor, 2010). In the Bungo district of Jambi
province alone, forest cover decreased by 9964
ha/year in the period of 1988–1993, but by only
1211 ha/year in the period of 2002–2005. Between
1988 and 2005, almost 40 percent of the Bungo
area was converted to intensive agriculture, such
as rubber and oil palm plantations. Rubber trees
are planted in both monoculture and agroforest-
ry systems. Between 1973 and 2005, the area un-
der rubber agroforest in Bungo decreased from
15 to 11 percent, while the area under monocul-
ture plantations increased from 3 percent to over
40percent (Ekadinata and Vincent, 2011).
Although rubber monoculture systems using
clonally propagated rubber trees can produce
large amounts of latex, agroforests can provide
multiple environmental services while ensuring
farmer livelihoods (Tomich et al., 2002; Schroth
et al., 2004). Rubber agroforest consists of mix-
tures of rubber trees with other species that re-
generate naturally from seed banks or dispersal
agents. Some important species, such as fruit
trees, are deliberately planted (Joshi et al., 2002).
Rubber agroforests range in intensity from sec-
ondary forests with some rubber (e.g. 5–10 per-
cent of tree basal area) to vegetation dominated
by rubber with a complement of native forest
trees. So-called “complex agroforest” systems
are characterized by a substantial (but less than
50percent) proportion of rubber trees in the total
biomass and a high diversity of native forest trees
and understorey plants (Gouyon, de Foresta and
Levang, 1993).
To counterbalance the high rate of deforesta-
tion, the Government of Indonesia has initiated
tree-planting efforts during the last three dec-
ades. Tree plantings using exotic and fast-grow-
ing species, such as brown salwood (Acacia man-
gium Willd.) and Eucalyptus spp., would provide
resources for pulp and paper industries. Some for-
est rehabilitation is based on enrichment planting
with native tree species, such as Dipterocarpaceae
species (Nawir, Murniati and Rumboko, 2007).
Dipterocarp seedlings tend to be shade-tolerant,
so are suitable to be planted in an agroforestry
system with rubber trees. Planting dipterocarp
trees helps to meet the challenge posed by do-
mestic demand for timber, despite being con-
strained by rules and regulations on extracting
hardwood from farm-forests.
Several studies have been conducted on en-
richment planting in rubber plantations with
Dipterocarpaceae in various areas of Bungo and
Tebo districts, Jambi province (Anonymous, 2004;
Tata et al., 2010). These have shown that diptero-
carp species grow well in rubber plantings and do
not suffer from mycobionts and abiotic factors
such as soil and microclimate (particularly light
availability).
Here we report on the early growth of meranti
in rubber agroforests in three villages in Bungo
district and farmers’ participation in tree enrich-
ment planting in rubber smallholdings.
Activities of enrichment planting
Study site
Bungo district is located in western Jambi Prov-
ince, Sumatra, Indonesia. Bungo has the third
largest area of rubber agroforest in Indonesia.
The sites were selected based on degree of land

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intensification: (i) low intensification (with forest
and complex rubber agroforests dominating the
landscape) was represented by Lubuk Beringin
village; (ii) intermediate intensification (with
complex to simple rubber agroforests dominat-
ing) was represented by Tebing Tinggi village;
and (iii) high intensification (with simple rubber
agroforests, monoculture rubber and oil palm)
was represented by Danau village (Therville, Fein-
trenie and Levang, 2011) (Figure 13.4).
Farmer selection
One farmer was selected at each site based
on willingness to collaborate. The farmers are
aware that wood stocks are getting scare, ow-
ing to few remnant natural forests and lack of
wood supplies from plantations to meet local
demand. Staff from the World Agroforestry Cen-
tre (ICRAF) provided technical support to farmers
to familiarize them with the different character
and silvicultural requirements of dipterocarp spe-
cies compared with rubber trees. An earlier study
showed that very few rubber farmers (about 12.5
percent of respondents) had experience in plant-
ing forest-tree species (Tata and van Noordwijk,
2012). Farmers were responsible for regular seed-
ling maintenance in the plots.
Rubber agroforest development and
maintenance
Establishment of rubber agroforests begins with
slashing the forest cover and burning it during the
dry season. This method is relatively cheap and
commonly applied by farmers in the area, in part
because they believe that ash improves soil fertility
(Ketterings et al., 1999). The plots in the three sites
were established 10–12 years ago, at a planting
distance of 6 × 3 m or 6 × 5 m. Rubber trees were
being tapped by the time Shorea leprosula seed-
lings were being interplanted in the rubber plots.
Figure 13.4.
Sites where dipterocarp species were planted in rubber agroforests in Bungo district, Jambi

GENET IC CONS IDERATI ONS IN ECO SYSTEM RE STORATI ON USING N ATIVE TRE E SPECIE S
181
Tree species and planting
Shorea leprosula is a native species that grows in
the lowland forest of Sumatra. It is known locally
as meranti batu. S. leprosula wood is sold as red
meranti, and is used for light construction, fur-
niture and moulding. The species can be grown
in various soil types, from fertile to poor. It has a
long life cycle; wood can be harvested 20–25 years
after planting. Shorea leprosula grown on Ultisol
soil in central Kalimantan grows by about 3.2 cm/
year, and hence is classified as a fast-growing me-
ranti (Soekotjo, 2009).
Seedlings were bought from an uncertified
vendor in Sungai Duren village, Jambi. Wildings
were collected from the surrounding remnant
forest areas. Therefore the age and the origin of
mother trees of the wildings were unknown. The
seedlings were planted between rubber trees in
the rubber gardens at a spacing of 10× 7m. The
number of S. leprosula seedlings planted depend-
ed on the area of the rubber garden, ranging
from 48 to 70 trees per plot. All farmers actively
maintained the S. leprosula seedlings, weeding
an area around them, but applied no fertilizer.
Dipterocarps form symbioses with ectomycor-
rhizal fungi, but meranti do not need to be in-
oculated with the fungi to establish in tropical
forest (Lee, 2006; Tata et al., 2010). This proved
to be the case in this study; S. leprosula seedlings
were not manually inoculated with ectomycor-
rhizal fungi but most of the roots of seedlings
were naturally inoculated by unidentified ecto-
mycorrhizal fungi.
Monitoring and experiences
Survival of S. leprosula in rubber plots
Survival rate of S. leprosula six months after
planting ranged from 46.5 percent to 59.2 per-
cent in the three plots and remained the same at
12 months after planting (Figure 13.5). Survival
rate was lowest at the Tebing Tinggi site because
wild pig (Sus scrofa) attacked both rubber trees
and S. leprosula trees in the plots. Similar attacks
on Shorea seedlings in other plots in Bungo and
Tebo district were also reported by Tata et al.
(2010).
Figure 13.5.
Survival rate of S. leprosula in three sites in Bungo District, Indonesia

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Early growth of S. leprosula in rubber plots
Height growth was greatest in Dusan Danau
while growth in stem diameter was greatest in
Dusan Danau and Tebing Tinggi (Figure 13.6). The
poor growth of S. leprosula in Lubuk Beringin was
the result of poor maintenance, particularly lack
of weeding, by the farmer at that site.
Farmer participation
Many rubber farmers are reluctant to plant
forest-trees in rubber plots because they believe
that rubber and timber trees compete for soil
nutrients and light, which reduces production
of latex. The farmers who took part in the cur-
rent study were willing to do so because: (i) they
received free seedlings and technical assistance,
(ii) they were aware of the shortage of timber in
their areas, (iii) they could use the planted trees
as a means of saving and as collateral for credit,
(iv) they were innovators and (v) because they
received recognition from others (Tata and van
Noordwijk, 2012).
The participating farmers also planted other
trees, including Litsea sp., bitter bean (Parkia
speciosa Hassk.) and Archidendron jiringa (Jack)
I. C. Nielsen. Litsea sp., which produce light tim-
ber and usually regenerate naturally, while
P. speciosa and A. jiringa, which are grown for
their fruit, are usually planted but can regener-
ate naturally in rubber plots during the fallow
period. Although S. leprosula is a native species
and produces good timber, it is not commonly
planted by the farmers in rubber plots. Farmers
are mostly interested in planting exotic species,
such as teak (Tectona grandis L.f.) and big leaf
mahogany (Swietenia macrophylla King) because
of their market value. Market access is one of the
key reasons why a farmer will plant a commodity
tree. Farm forests of teak, albizia (Paraserianthes
falcataria (L.) Nielsen) and some other timber spe-
cies are already well established and supported
by the Government of Indonesia. In contrast, dip-
Figure 13.6.
Early growth of S. leprosula in three different plots in Bungo districts: Tebing Tinggi, Lubuk Beringin
and Dusun Danau: (a) diameter growth, (b) height growth
Note: vertical line shows standard error.

GENET IC CONS IDERATI ONS IN ECO SYSTEM RE STORATI ON USING N ATIVE TRE E SPECIE S
183
terocarps such as Shorea spp. can be harvested for
sawn wood only in forest-concession areas, that
is, industrial plantation forest (hutan tanaman
industri) and community plantation forest (hutan
tanaman rakyat), and not from farm forests (hu-
tan rakyat). This restriction on harvesting, trans-
porting and marketing timber of dipterocarps
from farm forests, such as rubber agroforest,
hampers forest restoration using native species.
Conclusions
Dipterocarp species native to Indonesia are rec-
ommended for use in ecosystem restoration in
Sumatra. Red meranti (S. leprosula) grows well in
the rubber agroforestry systems in Bungo district,
Jambi province. However, active participation
of farmers in restoration activities is essential to
achieve high survival rate and performance of the
planted seedlings. Changes to government regu-
lations are required to permit harvesting, trans-
port and marketing of red meranti from farm for-
ests, such as rubber agroforests.
Acknowledgements
The authors acknowledge Landscape Mozaics
Project funded by the Swiss Agency for Deve-
lopment and Cooperation (SDC). We thank two
anonymous reviewers for their critical reviews of
the manuscript.
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13.4. “Rainforestation”:
a paradigm shift in forest
restoration
Paciencia P. Milan
Visayas State University, Philippines
With less than 24 percent of its forest cover re-
maining, the Philippines is experiencing loss of
ecosystem services such as biodiversity mainte-
nance, carbon sequestration, watershed protec-
tion and forest products for local communities.
As a result, natural disasters such as flash floods,
landslides and even water shortages have in-
creased. Despite measures taken to curtail forest
destruction, the forests continue to decline. Most
reforestation efforts in the Philippines focus on
the development of forestry and agroforestry
systems using tree species selected for their fast
growth and easy germination. Species composi-
tion of the forest that covered the land prior to
logging is rarely taken into account (Margraf and
Milan, 1996).
The use of non-native species in forest resto-
ration impacts greatly on forest biodiversity in
the Philippines. Native tree species are being lost
from the landscape because their timber is still
sought, especially for construction. They continue
to be cut even if other types of timber are avail-
able. However, in spite of their popularity for
their high wood quality, native or local trees are
not propagated or used in reforestation. Hence,
mother trees have become rare, which in turn re-
duces the availability of propagation material. As
the rainforest tree species are depleted and mon-
oculture of exotic or introduced species in refor-

GENET IC CONS IDERATI ONS IN ECO SYSTEM RE STORATI ON USING N ATIVE TRE E SPECIE S
185
estation proliferates, the survival of local wildlife
species is at stake. Some of them play an impor-
tant role in pollination and seed dispersal (Hutter,
Goltenboth and Hanssler, 2003). The repeated
clearcutting of fast-growing exotics on reforesta-
tion sites rapidly exhausts soil nutrients and, to
some extent, water, making reforestation increas-
ingly difficult. Another drawback of monoculture
of exotic species such as Gmelina, mahogany and
Leucaena is their vulnerability to pests.
A paradigm shift in reforestation is needed
to achieve sustainability. As reforestation in the
Philippines can be described as a failure in terms of
restoring original vegetation, an innovative strategy
known as “rainforestation” has been developed
through a joint research project with the Philippine-
German Tropical Ecology Program, a bilateral project
between the German Organization for Technical
Cooperation (GTZ, now the Deutsche Gesellschaft
für Internationale Zusammenarbeit, GIZ) and the
Visayas State College of Agriculture, the Philippines
(now Visayas State University) in 1996.
The rainforestation concept
Only indigenous and native forest tree species
are used in rainforestation. This approach em-
phasizes improvement of the structural habitat
to support wildlife. It consists of three opera-
tional frames: habitat restoration, biodiversity
conservation and provision of ecological ser-
vices. Rainforestation is more consistent with
biodiversity conservation strategies such as pro-
tected-area management and natural regenera-
tion than conventional reforestation efforts.
The planting scheme in the restoration areas
involves planting sun-loving native tree species
at a close spacing of 2 × 2 m to shade out
weeds. Species used on limestone soils include
Terminalia microcarpa Decne. (known locally as
kalumpit), Calophyllum inophyllum L. (bitaog),
Vitex parviflora Juss. (molave), Melia dubia
Cav. (bagalunga), Dracontomelon dao (Blanco)
Merr. & Rolfe (dao), Calophyllum blancoi
Planch. & Triana (bitanghol), Vitex pinnata L.
(lingo-lingo) and other fast-growing pioneer
species. Shade-loving trees (Table 13.3) are
planted between rows when the pioneer trees
start to provide shade. Selection of pioneer
tree species depends on the soil type, whether
limestone or volcanic.
Once these tree species have established,
some specialized arthropods, birds and lizards
stage a comeback. Over time, biodiversity
increases with the number of native tree species
and the structural diversity it offers to wildlife
(Milan, 1996). In addition, the closed forest
system promotes nutrient cycling. Aside from
biodiversity improvement of rainforestation
farms, soil properties, biological activities
and microclimate improve noticeably. Soil
pH and structure improve, increasing water-
holding capacity. In calcareous soil (Punta)
and the demonstration site of the Visayas
State University, soil colour changed from
light sandy to dark brown or black, and soil
organic matter and nutrient content increased,
especially nitrogen, calcium and magnesium.
Soil microclimate improved becoming moist
and cooler, and soil arthropods and other fungi
proliferated (Asio et al., 1995).
TABLE 13.3.
Shade-loving local forest tree species of Leyte,
Philippines, recommended for rainforestation on
volcanic soils
Local Name Scientific Name
Palosapis
Anisoptera thurifera
(Blanco) Blume
Apitong
Dipterocarpus grandiflorus
Slooten
Hairy Apitong
Dipterocarpus philippinensis
Foxw.
Hagakhak
Dipterocarpus warburgii
Brandis
Dalingdingan
Hopea foxworthyi
Elmer
Yakal-kaliot
Hopea malibato
Foxw.
Almon
Shorea almon
Foxw.
Guijo
Shorea guiso
Blume
Yakal-malibato
Shorea malibato
Foxw.
Red lauan
Shorea negrosensis
Foxw.
Tangile
Shorea polysperma
Merr.
Kamagong
Diospyros philippensis
(Desr.) Gürke

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Over almost two decades rainforestation
believers have been raising planting materi-
als of native tree species in nurseries across the
Philippines. Most of those trained in nursery
man agement took up growing native tree seed-
lings as a livelihood endeavour.
In spite of advocacy for the use of indigenous
forest tree species, the use of native tree species
in reforestation programmes has received very lit-
tle support, mainly because of the following per-
ceived limitations:
• Native species, especially dipterocarps, grow
slowly.
• Dipterocarps fruit approximately only every
three to five years, depending on species
and locality.
• Too few seedlings can be produced in a
short time because nursery management of
native species is not well understood.
• Dipterocarp seedlings and other native
species require shade and cannot be used to
reforest open areas.
Rainforestation farming
Rainforestation farming is a systems perspective
in forest restoration that emanated from the
rainforestation concept. It not only preserves
forest biodiversity but simultaneously sustains
human food production (Milan and Margraf,
1994). In rainforestation farming, native or
indigenous tree species are used in combination
with agricultural crops and fruit trees (Figure 13.7).
The rainforestation farming system, when
appropriately understood and implemented, can
replace the destructive form of kaingin or slash-
and-burn farming, allowing upland farmers to
continuously crop even a small area (minimum
of 0.05 ha). Planting fast-growing native trees
in the first year and premium tree species in the
following years contributes to the conservation
of local genetic resources. By incorporating fruit
trees and other crops, rainforestation farming
can provide farmers with additional income.
Thus, rainforestation not only contributes to
saving forest ecosystems but also helps to address
Figure 13.7.
Example of the combination of native tree species and fruit trees in rainforestation farming
Source: adapted from Margraf and Milan (1996).

GENET IC CONS IDERATI ONS IN ECO SYSTEM RE STORATI ON USING N ATIVE TRE E SPECIE S
187
the needs of farmers for food, timber and other
forest products in a sustainable way (Goltenboth
and Hutter, 2004). As a result, it is acceptable to
resource-poor farmers and landowners alike.
Most assessments of the benefits of reforesta-
tion focus on the easily measurable economic
benefits and ignore non-monetary benefits, such
as ecosystem services, which are harder to quan-
tify. Promoting adoption of rainforestation rather
than use of traditional approaches to reforesta-
tion will depend on achieving a deeper under-
standing of the interplay between the potential
to improve farmer income and the ecological
function of the forest biodiversity.
Rainforestation offers the prospect of sus-
tainable development in the uplands and for-
est ecosystems of the Philippines. This was rec-
ognized by the Philippine Government in 2011,
when rainforestation was adopted as a strategy
in the National Greening Program implemented
by the Department of Environment and Natural
Resources (Executive order no. 26, 24 February
2011).
References
Asio, V.B., Jahn, R., Stahr, K. & Margraf, J. 1995. Soils
of the tropical forests of Leyte, Philippines. II: Impact
of different land uses on status of organic matter
and nutrient availability. In A. Schulte & D. Ruhiyat,
Eds. Soils of tropical forest ecosystems: characteris-
tics, ecology and management, pp. 37–44. Berlin,
Springer.
Goltenboth, F. & Hutter, C.P. 2004. New options
for land rehabilitation and landscape ecology in
Southeast Asia by “rainforestation farming.” J. Nat.
Conserv., 12: 181–189.
Hutter, C.P., Goltenboth, F. & Hanssler, M. 2003. Paths
to sustainable development. Stuttgart, Germany,
S. Hirzel Verlag.
Margraf, J. & Milan, P.P. 1996. Ecology of lowland ever-
green forests and its relevance for island rehabilita-
tion on Leyte, Philippines. In A. Schulte & D. Schöne,
eds. Dipterocarp forests ecosystems: towards
sustainable management, pp. 124–154. Singapore,
World Scientific.
Milan, P.P. 1996. Conserving biodiversity through water-
shed management. Paper presented at the Colegio
de San Juan de Letran, Calamba, on the occasion of
their Foundation Day, 11 March 1996.
Milan, P.P. & Margraf, J. 1994. Rainforestation farming:
an alternative to conventional concepts. Ann. Trop.
Res., 16(4): 17–27.

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13.5. The permanent polycyclic
plantations: narrowing the
gap between tree farming
and forest
Enrico Buresti Lattes,1 Paolo Mori2 and
Serena Ravagni3
1 Association for Tree Farming for
Economy and the Environment, Arezzo,
Italy
2 Compagnia delle Foreste, Arezzo, Italy
3 Agricultural Research Council (CRA),
Forestry Research Center, Arezzo, Italy
Many of the environmental benefits provided by
tree plantations (tree farming) are lost at the end
of a management cycle when the trees are felled.
Permanent polycyclic24 plantations, which com-
bine the advantages of plantations with some of
those of the forest, can help avoid this disadvan-
tage (Buresti Lattes and Mori, 2009).
Many living organisms, including animals, in-
sects and plants, are associated with each tree.
Thus, diversity increases and becomes more com-
plex as the number of trees and tree species and
the complexity of the vertical or horizontal struc-
ture of the forest increase. This aspect is very im-
portant not only in forested landscapes, but also
in areas such as intensively cultivated agricultural
lands and peri-urban areas, where the presence
of trees and shrubs can increase biological diver-
24 In this chapter polycyclic means the contemporary presence of
two or more wood production cycles of different lengths on one
plot of land.
sity. Trees also influence microclimate, regulate
water flows and reduce the effect of some pollut-
ants. Moreover, trees sequester CO2 from the at-
mosphere. On permanently forested sites, carbon
steadily accumulates in the soil and the amount
of carbon sequestered in the soil may exceed the
amount stored in the plant biomass (Petrella and
Piazzi, 2006).
Thus, trees are important not only for produc-
tive purposes but also because of their ecological
and landscape impacts. Therefore, tree farming,
especially polycyclic plantations, has an impor-
tant environmental role. Polycyclic plantation
tree farming covers a wide range of planning and
management approaches, from the Italian classi-
cal poplar cultivation (AA.VV., 1987) with large
farmer inputs and strong impact on the environ-
ment to silvicultural approaches with low farmer
inputs and little environmental impact. However,
all tree plantations generally have the same end
point: when the main trees reach the end of their
economic life, all trees in the plantation are felled
and the ecological and landscape benefits of the
plantation are lost.
Recently, researchers have started testing new
permanent polycyclic plantations in order to ex-
tend the ecological benefits derived from planta-
tions while maintaining profits to the farmer.
What are permanent polycyclic plantations?
Polycyclic plantations are defined as plantations,
generally mixed, where there are several groups
of main trees with different objectives and
lengths of productive cycles. Thus, for example,
a classical cloned poplar plantation is monocyclic
while a mixed plantation of poplar clones and
walnut (Juglans regia L.) is a polycyclic plantation.
Recently, it was considered necessary to distin-
guish polycyclic plantations from permanent poly-
cyclic plantations (Buresti Lattes and Mori, 2006,
2007a). In polycyclic plantations the species with
the longest production cycle are planted at a den-
sity that allows them to develop a closed canopy
at the end of their production cycle; these trees
are generally clear cut once they reach maturity.
Permanent polycyclic plantations differ from non-

GENET IC CONS IDERATI ONS IN ECO SYSTEM RE STORATI ON USING N ATIVE TRE E SPECIE S
189
permanent plantations in terms of tree spacing
and management strategy. Distances between
the main trees with a longer production cycle are
greater than in polycyclic plantations to prevent
development of a closed canopy (Buresti Lattes
and Mori, 2007b). This allows trees of the same
or different species to be planted between these
main trees on a different production cycle, and
the land remains continuously under tree cover.
As an example we can consider a mixed polycy-
clic plantation with poplar clones (but also native
species of Populus alba L. or Populus nigra L.) and
walnut. In a polycyclic plantation (e.g. Ravagni
and Buresti, 2003; Buresti Lattes and Mori, 2007b,
2009; Buresti Lattes et al., 2008a), when poplar is
felled walnut occupies all of the liberated space,
and the farmer must wait for the walnut to com-
plete its production cycle before establishing a
new cycle with poplar or other (native) species.
However, with permanent polycyclic plantations,
walnut trees are spaced such that at the end of
the production cycle their canopies do not occupy
all the available space but instead leave space for
subsequent tree populations. Thereby, after po-
plar felling the farmer may decide to start a new
production cycle by planting poplar, or another
(native) species according to the environmental
conditions and available space. While walnut con-
tinues to grow, the farmer may thus be able to
produce two or more cycles with other species in
the available space. When the walnut trees are fi-
nally cut, trees of other species remain in the plan-
tation to buffer the temporary and partial lack of
large trees. A new productive cycle can then be
started, replacing the harvested trees with walnut
or other native species according to the farmer’s
production aims.
The arrangements and benefits of polycyclic
plantations
Permanent polycyclic plantations require careful
planning and management that is adapted to the
needs of the species and their different produc-
tion cycles. The planner has to choose the spacing
between trees of the same and different produc-
tion cycles to allow for optimal performance of all
trees. The larger the number of production cycles
to be combined in the plantation, the greater the
complexity of the design.
The farmer has to understand the growth dy-
namics of the different trees and the timing of
each production cycle in order to conduct the
management interventions at the appropriate
times (e.g. pruning, felling and introduction of
the new production cycle). Technical advice is very
important during all these operations.
Planning and managing a polycyclic, mixed,
multi-objective permanent plantation (PMMP) is
certainly more difficult than, for example, plan-
ning and managing a normal mixed plantation.
However, PMMPs can provide a number of ad-
vantages for the farmer (Buresti Lattes, Mori and
Ravagni, 2001; Buresti Lattes and Ravagni, 2003;
Becquey and Vidal, 2008a, 2008b; Buresti Lattes
and Mori, 2010):
• With the right spacing, older trees will
influence the form of younger trees, making
pruning simpler.
• With overlapping production cycles, income
is more frequent and economic return can
be higher than from a simpler plantation.
• The plantation can be redesigned after trees
have been felled in each production cycle;
changes can be made to species, spacing and
production objectives and exploitation of
the available space improved.
Moreover, permanent plantations provide eco-
logical and other benefits for society that cannot
be achieved with traditional plantations and non-
permanent polycyclic plantations. These include:
• less change in the landscape over time;
• continuous carbon storage; and
• less habitat change for fauna that depend
on trees for refuge and food.
Case study: polycyclic permanent plantations
in Mantua, Italy
One of the first experimental PMMPs was estab-
lished in the province of Mantua, Italy, in 2006.
The plantation was established on a farm devoted
to poplar cultivation where the owner was inter-
ested in producing poplar with fewer external

THE STATE OF THE WORLD’S FOREST GENETIC RESOURCES – THEMATIC STUDY
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190
inputs and higher environmental value (Buresti
Lattes et al., 2008b).
The planting scheme proposed (Figure 13.8)
used main trees of three different tree species:
pedunculate oak (Quercus robur L.) and poplar
(several clones) to produce timber and hornbeam
(Carpinus betulus L.) to produce biomass for fuel
or wood panels. Oak and hornbeam are native
to Italy. The objective for the oak was to produce
logs of 40–45 cm in diameter and 4m in length;
for poplar the objective was to produce logs of
40 cm in diameter and 6 m in length. Hornbeam
was selected not only for the good quality of its
wood as biomass, but also for its capacity to grow
in partial shade and for its low competitiveness
towards oak located only 4.5m away.
A main tree provides at least one of the main
products for which the plantation was designed.
An accessory tree or shrub facilitates the
management of the plantation by the farmer but can
be substituted by cultural care.
Multifunctional tree farming refers to tree
cultivation designed to satisfy multiple functions
(e.g. timber production and reduction of pollutants
into waterways, or, in the case of common walnut,
timber and fruit).
Multi-objective tree farming refers to tree
cultivation designed to obtain more than one type of
wood product (e.g. timber and biomass).
Box 13.2.
Definitions of terms
Legend to Figures 13.8 to 13.18
Main plants
Pedunculate oak (Quercus robur L.)
Poplar clone
Hornbeam (Carpinus betulus L.)
Accessory plants
Black alder (Alnus glutinosa (L.) Gaertner)
Buckthorn (Rhamnus frangula L.) or
viburnum (Viburnum spp.)
Figure 13.8.
Year 0. The first plantation scheme of all species.
Figure 13.9.
Year 10. Poplars should have reached the
production target of 40 cm trunk diameter and
are felled.

GENET IC CONS IDERATI ONS IN ECO SYSTEM RE STORATI ON USING N ATIVE TRE E SPECIE S
191
Accessory trees and shrubs (all native) were
planted along the oak rows to improve shape and
reduce branching in young oaks and to mitigate
the isolation stress that might occur after harvest-
ing the poplar. The accessory trees and shrubs in-
cluded black alder (Alnus glutinosa (L.) Gaertner),
buckthorn (Rhamnus frangula L.) and laurustinus
viburnum (Viburnum tinus L.). Nitrogen fixation by
the symbiotic bacteria in black alder roots is also
expected to improve the productivity of the oak.
Figures 13.9 to 13.18 describe one possible evo-
lution and management strategy for the planta-
tion over 35 years. Production cycles are expected
to be 10 years for poplar, 35 years for oak and
10–15 years for hornbeam. The first cycle of
hornbeam is relatively long compared with that
of other native species that are suitable for the
production of woody biomass. This is because of
the relatively slow initial growth of hornbeam
and the need to protect the adjacent oak trees.
Sudden isolation of the oaks should be avoided
as it could cause stress reactions, and therefore
surrounding poplar and hornbeam must be felled
alternately. After the first ten years, one or other
tree species will be harvested every five years.
Felled trees are replaced with the same or differ-
ent species, adjusting the planting distance ac-
cording to species competition at each stage of
development. Over time, location of the species
on the site can rotate; where there were valuable
timber trees is it possible to plant trees for bio-
mass production and vice versa.
The scheme may change over time, depend-
ing on the development of one or more species
or changing needs and preferences of the owner.
This plantation scheme is just one possibility.
The choice of species and management strategies
will depend on environmental conditions, farmer
needs and applicable regulations in each area and
may change over time.
Figure 13.10.
Year 10. After felling the poplar trees, two new
poplar rows are planted. The new poplar rows
will be separated by 7 m and will be 10 m from
the oaks. The hornbeam trees should not be
cut at the same time as the poplars to avoid the
excessive and sudden isolation of the oaks, which
could cause stress reactions.
Figure 13.11.
Year 15. Hornbeam is felled. The first cycle of
hornbeam is relatively long compared with that
of other native species that are suitable for the
production of woody biomass. This long time is
due both to the relatively slow initial growth of
the hornbeam and to the need to extend the
oak protection for additional years.

THE STATE OF THE WORLD’S FOREST GENETIC RESOURCES – THEMATIC STUDY
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192
Figure 13.12.
Year 20. Poplars should have reached the
production target of 40 cm in diameter and are
felled.
Figure 13.13.
Year 20. Two new rows of poplar are planted.
This time the rows are staggered in order to
increase the spacing (trees will be 5 × 5.6 m
from each other). The minimum distance
from oaks (which are 20 years old and with
a well-developed crown) will be 11 m. This
distance should be enough to allow poplar to
grow without competition with oaks. The two
five-year-old hornbeam rows, which are 4.5 m
from the oaks, should have a positive effect,
protecting oaks from isolation stress.
Figure 13.14.
Year 25. The hornbeam suckers, which grow
faster than seedlings, should be ready for felling.
Figure 13.15.
Year 30. The third cycle of poplar should be
felled and the hornbeam between the two
poplar rows removed.

GENET IC CONS IDERATI ONS IN ECO SYSTEM RE STORATI ON USING N ATIVE TRE E SPECIE S
193
Figure 13.18.
Year 35. New rows of poplar and hornbeam are planted. Trees for timber production are now
planted where trees for biomass production were before, and vice versa.
Figure 13.16.
Year 30. A new row of oaks and accessory trees
is planted together with two new rows of
hornbeam 4.5 m from the oaks.The two new
hornbeam rows are 9 m from the 30-year-old
oaks.
Figure 13.17.
Year 35. The oaks should have reached the target
size and should be felled. The hornbeam that
is closest to the oak trees should also be felled,
but yield will be low. At this point, thanks to the
positive competition with the older oaks and to
the microclimate provided by the oaks, the five-
year-old hornbeam plants should be adequately
developed.

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References
AA.VV. 1987. Pioppicoltura. Rome, Ente Nazionale per la
Cellulosa e per la Carta.
Becquey, J. & Vidal, C. 2008a. Enseignements de deux
plantations mélangée de peuplier I214 et de noyer
hybride. Forêt Entreprise, 178: 31–36.
Becquey, J. & Vidal, C. 2008b. Le mélange peuplier-
noyer est il économiquement intéressant? Forêt
Entreprise, 178: 37–39.
Buresti Lattes, E. & Mori, P. 2006. Legname di pregio e
biomassa nella stessa piantagione. Foreste ed Alberi
Oggi, 127: 5–10.
Buresti Lattes, E. & Mori, P. 2007a. Progettare impianti
policiclici e termine e multi obiettivo. In Arboricoltura
da legno: schede per la progettazione e la con-
duzione della piantagione. Schede 4° and 10°.
Direzione Centrale Risorse Agricole Naturali Forestali
e Montane, Regione Friuli Venezia Giulia.
Buresti Lattes, E. & Mori, P. 2007b. Distanze minime
d’impianto: prime indicazioni per le piantagioni da
legno. Foreste ed Alberi Oggi, 137: 13–16.
Buresti Lattes, E. & Mori, P. 2009. Impianti policiclici
permanenti: l’arboricoltura si avvicina al bosco.
Foreste ed Alberi Oggi, 150: 5–8.
Buresti Lattes, E. & Mori, P. 2010. Les plantations
plycicliques permanetes: l’arboriculture se rapproche
à la forêt. Forêt Entreprise, 195: 61–64.
Buresti Lattes, E. & Ravagni, S. 2003. Piantagioni con
pioppo e noce comune: accrescimenti e sviluppo
dopo i primi anni. Foreste ed Alberi Oggi, 94: 19–24.
Buresti Lattes, E., Mori, P. & Ravagni, S. 2001.
Piantagioni miste con pioppo e noce comune.
Foreste ed Alberi Oggi, 71: 11–17.
Buresti Lattes, E., Mori, P., Pelleri, F. & Ravagni, S.
2008a. Des peupliers et des noyers en mélange avec
des plants accompagnateurs. Forêt Entreprise, 178:
26–30.
Buresti Lattes, E., Cavalli, R., Ravagni, S. & Zuccoli
Bergomi, L. 2008b – Impianti policiclici di arbo-
ricoltura da legno: due esempi di progettazione e
utilizzazione. Foreste ed Alberi Oggi, 139: 37–39.
Petrella, F. & Piazzi, M. 2006. Carbonio nei suoli se-
minaturali piemontesi. Foreste ed Alberi Oggi, 123:
29–34.
Ravagni, S. & Buresti, E. 2003. Piantagioni con pioppo
e noce comune: accrescimento e sviluppo dopo i
primi anni. Foreste ed Alberi Oggi, 94: 19–24.

GENET IC CONS IDERATI ONS IN ECO SYSTEM RE STORATI ON USING N ATIVE TRE E SPECIE S
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14.1. Mangrove forest
restoration and the
preservation of mangrove
biodiversity
Roy R. Lewis III
Lewis Environmental Services Inc., Salt
Springs, Florida, United States
Mangrove forest ecosystems covered 13.8 million
ha of tropical shorelines in 2000 (Giri et al., 2011),
down from 19.8 million ha in 1980 and 15.9 mil-
lion ha in 1990 (FAO, 2003). These losses represent
about 2 percent per year from 1980 to 1990 and
1 percent per year from 1990 to 2000. Therefore,
achieving no net loss of mangroves worldwide
would require the successful restoration of ap-
proximately 150 000 ha per year, unless all major
losses of mangroves ceased. Increasing the total
area of mangroves worldwide towards their origi-
nal extent would require an even larger effort.
An example of documented losses of man-
groves is the combined losses in Malaysia, the
Philippines, Thailand and Viet Nam of 7.4 million
ha (Spalding, 1997). These figures emphasize the
magnitude of the loss. The opportunities that ex-
ist to restore areas back to functional and biodi-
verse mangrove ecosystems are also significant,
including mosquito-control impoundments in
Florida (Brockmeyer et al., 1997) (several tens of
thousands of hectares) and abandoned shrimp
aquaculture ponds in Southeast Asia (Stevenson,
Lewis and Burbridge, 1999) (several hundreds of
thousands of hectares).
While great potential exists to reverse the loss
of mangrove forests worldwide, most attempts
to restore mangroves fail completely or fail to
achieve the stated goals (Erftemeijer and Lewis,
2000; Lewis, 2000, 2005, 2009). Previously docu-
mented attempts to restore mangroves (Field,
1996, 1999), where considered successful, have
largely concentrated on creation of plantations
of mangroves consisting of just a few species with
the objective of providing wood products (Kairo
et al., 2002) or collecting eroded soil and raising
intertidal areas to usable terrestrial agricultural
elevations (Saenger and Siddiqi, 1993).
Restoration of a biodiverse mangrove forest
Successful mangrove forest restoration requires
careful analyses of a number of factors before
attempting actual restoration. Lewis (2005, 2009)
notes that existing hydrology of a proposed res-
toration site needs to be characterized and com-
pared with that of a reference forest to estab-
lish what conditions preclude natural recovery
in damaged forests, or what conditions prevent
natural recolonization of supratidal and subtidal
Chapter 14
Habitat-specific
approaches

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flats that might be proposed for conversion to
mangrove forests. A six-step process called eco-
logical mangrove restoration has evolved from
earlier attempts to standardize successful ap-
proaches (Stevenson, Lewis and Burbridge, 1999),
and is now taught around the world (Lewis 2010).
This method emphasizes getting the hydrology
right first and then observing and documenting
natural recovery through volunteer mangrove
propagule recruitment (Figure 14.1) before large-
scale planting of mangroves is even considered.
As seen in Figure 14.1, mangrove propagules
can voluntarily recruit to a restored site and es-
tablish the natural biodiversity of mangrove spe-
cies. Planting is therefore not needed in most
cases. Situations where planting of mangroves is
needed are described as “propagule limited” sites
(see below).
Unfortunately, as noted in Stevenson, Lewis and
Burbridge (1999) and Samson and Rollon (2008),
massive attempts to plant mudflats where man-
groves have never existed have been the norm for
many decades and have almost uniformly failed.
Where they occasionally do work because local
topographic conditions are conducive to planting
of mangroves, the results are typically plantations
of a single species of mangrove. Various species
of Rhizophora are commonly used in plantings as
they have large propagules that are easily collect-
ed, grown and planted. This emphasis on single-
species plantings ignores the mix of species found
in most mangrove forests. Mangrove forests in the
New World typically contain four species of man-
grove, and a single forest in a location such as the
Philippines, Viet Nam and northern Australia may
contain up to 30 species (Duke, 1992). There are 69
species worldwide called mangroves (Duke, 1992).
Biodiversity is also threatened by the introduc-
tion of non-native species of mangroves for res-
toration. Chen et al. (2009) notes that Sonneratia
apetala Buch.-Ham. has been introduced to China
from Bangladesh and, surprisingly, used to con-
trol another introduced plant species, Spartina
alterniflora Loisel, “even though the invasiveness
of this exotic mangroves species was not fully un-
derstood” (Chen et al. 2009: 49).
An important goal of many restoration projects
is to provide habitats for fish and invertebrates
to restore local fisheries. Maximizing use of such
habitats usually means maximizing biodiversity
of the plant species, and therefore a monotypic
stand of mangroves in an area that normally sup-
ports 20 or more mangrove species is not a logi-
cal goal. Establishment of persistent tidal creeks
to assist with entry and exit of juvenile and adult
fish and invertebrates is also an essential restora-
tion objective. Lewis and Gilmore (2007) discuss
the use by fish of both natural and restored man-
grove forests and report specifically on monitoring
a successful 500 ha mangrove restoration project in
Hollywood, Florida, United States (see Figure 14.1),
where fish populations sampled in both reference
and restored sites were statistically indistinguish-
able within three to five years of restoration. They
emphasize three restoration and design goals to
ensure functional and naturally biodiverse ecologi-
cal restoration of mangrove forests:
1. Achieve plant cover similar to that in an ad-
jacent relatively undisturbed control area of
mangrove forest.
2. Establish a network of channels that mimic the
shape and form of a natural tidal creek system.
3. Establish a heterogeneous landscape similar to
that exhibited by local mangrove ecosystems.
Lewis (2005) introduced the term “propagule
limitation” to define a condition in which natural
recovery is slowed or halted because no natural
mangrove propagules are available to volunteer
at a damaged site. The absence of propagules
may be caused by a large-scale loss of adult trees
capable of producing propagules or by hydrologic
restrictions or blockages (e.g. dykes) that prevent
natural waterborne transport of mangrove prop-
agules to a restoration site. Since propagules are
produced at different times of the year by dif-
ferent species in different locations (Tomlinson,
1986), more than one site visit may be necessary
to correctly identify a propagule limited site. Lack
of propagules at a single time of year does not
necessarily define a propagule-limited site, and
therefore careful evaluation of this parameter
is important. If a damaged forest will recover on

GENET IC CONS IDERATI ONS IN ECO SYSTEM RE STORATI ON USING N ATIVE TRE E SPECIE S
197
its own within an acceptable time frame, any at-
tempt to introduce propagules, plant propagules
or plant nursery-grown mangroves is likely to be
a waste of time and money. Recovery is here de-
fined as the recolonization of a restoration site
and growth of plant materials on that site reach-
ing some predefined numerical target (e.g. per-
cent cover, total basal area). Priority should be
given to restoration sites that would indeed ben-
efit from human intervention at the least per unit
cost, given that time and money to devote to any
restoration project are always limited.
Figure 14.1.
Time sequence photographs of a portion of the 500 ha West Lake Park mangrove restoration
project utilizing non-native exotic plan removal, site excavation, tidal creek restoration and natural
recruitment of mangrove propagules. No planting of mangroves took place.

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These suggestions may seem obvious, but there
are very few documented examples of successful
mangrove forest restoration. More commonly,
well-intentioned, but often faulty, mangrove res-
toration efforts target areas on which mangroves
were not previously present, such as mudflats or
seagrass meadows seaward of natural mangroves
or damaged areas without a properly document-
ed history (Field, 1996; Erftemeijer and Lewis,
2000; Lewis, 2005). The result of unsound evalu-
ations of restoration opportunities has, unfortu-
nately, emphasized first establishing a mangrove
nursery and then planting mangroves at a casual-
ly selected site as the primary tool in restoration,
rather than first assessing the reasons for the loss
of mangroves in an area and working with the
natural recovery processes (Lewis, 2009).
Both Brockmeyer et al. (1997) and Stevenson,
Lewis and Burbridge (1999) present examples of
successful mangrove restoration following re-
establishment of historical tidal connections to ad-
jacent estuaries. This is termed “hydrologic resto-
ration” (see discussion in Turner and Lewis, 1987).
In the examples discussed, volunteer propagules
of mangrove and mangrove nurse-plants were
sufficient to allow for rapid establishment of plant
cover. No planting of mangroves was required.
Establishing success criteria
Once a site is finally chosen for restoration and
a design developed, quantifiable success criteria
should be established. Establishing such criteria is
important in order to actually measure progress
towards successful restoration. The first step in es-
tablishing numeric criteria for success is to prepare
a brief narrative goal or set an objective for the
project (Saenger, 2002). This will define the next
steps. For example, a goal may be to establish a
monotypic plantation of Rhizophora apiculata Bl.
to be harvested after 12 years as poles. It may be
an acceptable goal to local stakeholders in the
project, such as local villages and fishermen, and
harvest of wood products from locally managed
forests is a typical goal (see discussion of timber
production in the Matang Forest, Malaysia, in
Saenger [2002]: 231–234).
A second example of a goal might be to maxi-
mize biodiversity. In this case, the restoration site
might be left alone and not planted immediately
to allow for volunteer colonization of the largest
number of different species of mangroves from
propagules produced by trees adjacent to a res-
toration site.
The next step is to look at available information
on both plantation and natural recruitment indi-
ces of success. Saenger (2002: 256–270) discusses
in great detail what is to be expected in terms of
biomass and stem density, for example, from typi-
cal plantation projects. There has been much work
on plantation projects in which just a few species
of mangroves are managed, and thus there is a
wealth of data to examine. In contrast with this,
data on natural recruitment within a mixed forest
are generally not available. McKee and Faulkner
(2000) report on the results of sampling for den-
sity and basal area within two restored mangrove
forests in Florida, United States, and compared
these with two adjacent control areas. Their data
show that density and basal area of volunteer
mangroves in the restoration areas exceeded that
of planted mangroves. Proffitt and Devlin (2005)
report similar results from one of the same sites
sampled by McKee and Faulkner (2000) but that
they sampled in later years as the system ma-
tured. Lewis, Hodgson and Mauseth (2005) report
on the results of cover sampling over a period of
five years within a restored mangrove forest in
another location in Florida, United States. These
studies help define parameters that need to be
sampled and sampling methodologies, but pro-
vide limited data to apply to local situations in
other parts of the world.
Few studies exist on trends in biodiversity in re-
stored mangroves, and the range in age, species
and inundation class of restored sites makes gen-
eralizations difficult. However, the co- occurrence
of many animal species in both restored and
comparable natural forests suggest that coloniza-
tion of restoration sites by both mobile and non-
mobile fauna is a rapid process, and equivalent
populations of mangrove fauna in both natural
controls and restored mangrove sites can typically

GENET IC CONS IDERATI ONS IN ECO SYSTEM RE STORATI ON USING N ATIVE TRE E SPECIE S
199
be found within 5–10 years of restoration (Lewis
and Gilmore, 2007; Bosire et al., 2008).
Concluding remarks
Restoration of mangrove forest has not been
generally successful except where timber produc-
tion was the goal and monotypic stands were
established. Establishment of a biodiverse mixed-
species forest cover and restoration of functions
equivalent to those of an adjacent reference for-
est, have not typically been design criteria, and
most restoration projects with some general eco-
logical goals have not been successful (Erftemeijer
and Lewis, 2000; Lewis, 2005). The chosen restora-
tion sites for many of these projects have been
mudflats or seagrass beds lying seaward of the
outer edge of existing mangrove forests. These
sites are typically planted with nursery-grown
mangrove seedlings which do not survive because
of frequent inundation and waterlogging.
Although there are relatively few studies on
trends in biodiversity in restored mangroves, it
appears that colonization of restoration sites by
both mobile and non-mobile fauna is a rapid pro-
cess that may take 5–10 years to reach levels com-
parable to natural sites (Bosire et al., 2008). The
scientific basis for optimum design of restoration
projects to meet certain established criteria, such
as increased fish production or more use by wad-
ing seabirds, is, however, very minimal.
In future, mangrove restoration projects should
be more carefully designed to ensure successful
establishment of a biodiverse plant cover over
large areas at minimal cost. This can be achieved,
for example, by restoring hydrologic connec-
tions to impounded mangrove areas, as has been
done in Florida (Brockmeyer et al., 1997), Costa
Rica and the Philippines (Stevenson, Lewis and
Burbridge, 1999) using the basic principles of eco-
logical mangrove restoration (Lewis, 2010). Use of
non-native species of mangroves in management
and restoration projects should be avoided.
References
Bosire, J.O., Dahdough-Guebas, F., Walton, M.,
Crona, B.I., Lewis, R.R., Field, C., Kairo, J.G. &
Koedam, N. 2008. Functionality of restored man-
groves: a review. Aquat. Bot., 89: 251–259.
Brockmeyer, R.E. Jr., Rey, J.R., Virnstein, R.W.,
Gilmore, R.G. & Ernest, L. 1997. Rehabilitation of
impounded estuarine wetlands by hydrologic recon-
nection to the Indian River Lagoon, Florida (USA).
Wetl. Ecol. Manag., 4(2): 93–109.
Chen, L., Wang, W., Shang, Y. & Lin, G. 2009. Recent
progresses in mangrove conservation, restoration
and research in China. J. Plant Ecol., 2(2): 45–54.
Erftemeijer, P.L.A. & Lewis, R.R. 2000. Planting man-
groves on intertidal mudflats: habitat restoration or
habitat conversion? In Proceedings of the ECOTONE
VIII Seminar Enhancing Coastal Ecosystems
Restoration for the 21st Century, Ranong, Thailand,
23–28 May 1999, pp. 156–165. Bangkok, Royal
Forest Department of Thailand.
Duke, N.C. 1992. Mangrove floristics and biogeogra-
phy. In A.I. Robertson & D.M. Alongi, eds. Tropical
mangrove ecosystems, pp. 63–100. Coastal and
Estuarine Studies 41. Washington, DC, American
Geophysical Union.
FAO (Food and Agriculture Organization of the
United Nations). 2003. Status and trends in
mangrove area extent worldwide, by M.L. Wilkie &
S. Fortuna. Forest Resources Assessment Working
Paper 63. Rome.
Field, C.D., ed. 1996. Restoration of mangrove eco-
systems. Okinawa, Japan, International Society for
Mangrove Ecosystems.
Field, C.D. 1999. Rehabilitation of mangrove ecosystems:
an overview. Mar. Poll. Bull., 37(8–12): 383–392.
Giri, C., Ochieng, E., Tieszen, L.L., Zhu, Z., Singh, A.,
Loveland, T., Masek, J. & Duke, N. 2011. Status
and distribution of mangrove forests of the world
using earth observation satellite data. Global Ecol.
Biogeogr., 20(1): 154–159.

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Kairo, J.G. 2002. Regeneration status of mangrove
forests in Mida Creek, Kenya: a compromised or
secured future? Ambio, 31(7–8): 562–568.
Lewis, R.R. 2000. Ecologically based goal setting in man-
grove forest and tidal marsh restoration in Florida.
Ecol. Eng., 15(3–4): 191–198.
Lewis, R.R. 2005. Ecological engineering for successful
management and restoration of mangrove forests.
Ecol. Eng., 24(4 SI): 403–418.
Lewis, R.R. 2009. Methods and criteria for successful
mangrove forest restoration. In G.M.E. Perillo, E.
Wolanski, D.R. Cahoon & M.M. Brinson, eds. Coastal
wetlands: an integrated ecosystem approach, pp.
787–800. Amsterdam, The Netherlands, Elsevier.
Lewis, R.R. 2010. Mangrove field of dreams: if we
build it, will they come? SWS Research Brief No.
2009–0005. Madison, WI, USA, Society of Wetland
Scientists (available at www.sws.org/researchbrief/
brief_pdf_dl.mgi?pdf=Lewis_061209).
Lewis, R.R. & Gilmore, R.G. 2007. Important considéra-
tions to achieve successful mangrove forest restora-
tion with optimum fish habitat. Bull. Mar. Sci., 80(3):
823–837.
Lewis, R.R., Hodgson, A.B. & Mauseth, G.S. 2005.
Project facilitates the natural reseeding of mangrove
forests (Florida). Ecol. Rest., 23(4):276–277.
McKee, K.L. & Faulkner, P.L. 2000. Restoration of
biogeochemical function in mangrove forests. Rest.
Ecol., 8(3): 247–259.
Proffit, C.E. & Devlin, D.J. 2005. Long-term growth and
succession in restored and natural mangrove forests
in southwestern Florida. Wetl. Ecol. Manage., 13:
531–551.
Saenger, P. 2002. Mangrove ecology, silviculture and
conservation. Dordrecht, The Netherlands, Kluwer
Academic Publishers.
Saenger, P. & Siddiqi, N.A. 1993. Land from the seas:
the mangrove afforestation program of Bangladesh.
Ocean Coast. Manage., 20: 23–39.
Samson, M.S. & Rollon, R.N. 2008. Growth perfor-
mance of planted red mangroves in the Philippines:
revisiting forest management strategies. Ambio,
37(4): 234–240.
Spalding, M.D. 1997. The global distribution and
status of mangrove ecosystems. Intercoast Network
Newsletter Special Edition #1: 20–21.
Stevenson, N.J., Lewis, R.R. & Burbridge, P.R. 1999.
Disused shrimp ponds and mangrove rehabilitation.
In W.J. Streever, ed. An international perspective on
wetland rehabilitation, pp. 277–297. Dordrecht, The
Netherlands, Kluwer Academic Publishers.
Tomlinson, P.B. 1986. The botany of mangroves.
Cambridge Tropical Biology Series. New York, USA,
Cambridge University Press.
Turner, R.E. & Lewis, R.R. 1997. Hydrologic restoration of
coastal wetlands. Wetl. Ecol. Manage., 4(2): 65–72.
14.2. Forest restoration in
degraded tropical peat
swamp forests
Laura L.B. Graham and Susan E. Page
Department of Geography, University
of Leicester, United Kingdom
Tropical peat swamp forests and their
degradation
Tropical peatlands in Southeast Asia are the most
extensive in the world; they contain ~69 Gt of car-
bon, equivalent to 11–14 percent of global peat-
land carbon, and cover 247 778 km2, the majority
being in Indonesia (206 950 km2; 57 Gt of carbon;

GENET IC CONS IDERATI ONS IN ECO SYSTEM RE STORATI ON USING N ATIVE TRE E SPECIE S
201
Page, Rieley and Banks, 2011). In addition, tropi-
cal peat swamp forests provide a range of other
environmental services including a habitat for en-
demic, endangered and rare species (Posa, Wije-
dasa and Corlett, 2011) and have a role in regional
hydrological regulation (Wösten et al., 2006).
Less than 4 percent of tropical peat swamp
forest in Southeast Asia remains in a near-intact
condition, with 37 percent classified as degraded
forest (logged-over), 24 percent as deforested,
burnt or both, and 32 percent as under agricul-
ture (Miettinen and Liew, 2010). Degradation of
peat swamp forest leads to the loss of most, if not
all, ecosystem services, including carbon storage
and hydrological regulation (Page et al., 2009).
Total annual CO2 emissions from Indonesian
peatlands (from peat oxidation, fire and loss of
biomass) over the period 2000 to 2006 have been
estimated at 640 Mt of CO2, equivalent to 2.1per-
cent of current annual global fossil fuel emissions
(Bappenas, 2010).
Given the extent of degraded peat swamp
forest in Southeast Asia and the implications for
increased carbon emissions to the atmosphere,
there is an international effort to promote ecosys-
tem rehabilitation. The Indonesian Government is
collaborating internationally to initiate large-scale
restoration programmes on degraded peatlands.
Overcoming the barriers to the restoration
of tropical peat swamp forest
When intact tropical peat swamp forest is dis-
turbed, environmental effects include changes in
microclimate (higher temperatures, reduced hu-
midity, altered light levels); lowering of the water
table, which in undisturbed peat swamp forest
is close to or above the peat surface; reduction
or loss of the hummock–hollow peat surface to-
Figure 14.2.
The distribution of the main peat deposits in Southeast Asia (shaded areas). Most peatlands occur on
the islands of Sumatra and Borneo (Kalimantan, Sarawak and Brunei) and in peninsular Malaysia.
Source: derived from Miettinen et al. (2012).

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pography, and hence loss of hummock surfaces
on which tree seedlings establish; and increased
peat oxidation and fire occurrence, both resulting
in peat surface subsidence and increased risk of
flooding (Page et al., 2008).
Disturbance history will be unique to each loca-
tion and will influence the rate and type of for-
est regeneration. Altered hydrological conditions
and fire are the main barriers to forest regenera-
tion (Page et al., 2009); therefore, a preliminary
assessment should be undertaken of land cover,
drainage history (time of installation, location,
size of drainage features) and fire history (fire lo-
cation, frequency and severity). If anthropogenic
impacts are minimal, forest regeneration capac-
ity is likely to be relatively high. For example,
in selectively logged forest or sites subject to a
single, low-intensity fire, secondary peat swamp
forest can establish over approximately ten years
(Hoscilo et al., 2011). With increasing intensity of
degradation (e.g. following intensive drainage
and frequent and/or severe fires), regeneration
of woody species is limited or entirely suppressed
and species diversity greatly reduced (Wösten
et al., 2006). Thus, active restoration measures
will be required to facilitate reforestation. In ad-
dition, information on annual hydrological varia-
tion (duration and extent of low and high water
levels, which will influence survival of transplants,
especially at the critical seedling stage), natural
seed dispersal by mammals and birds, levels of
competition from invasive, non-woody species
(ferns, sedges) and availability of nutrients and
mycorrhizal fungal symbionts in the surface peat,
all need to be assessed as potential barriers to
tree regeneration. Active barriers identified in
recent studies include wet-season flooding, high
light intensity, low levels of mycorrhizal fungi in
the surface peat and low levels of seed dispersal
(Page et al., 2008; Graham and Page, 2012).
Species selection
In Southeast Asia, peat swamp forest supports at
least 1500 plant species (Posa, Wijedasa and Cor-
lett, 2011). Of these, approximately only 11 per-
cent are endemic to peat swamp forest, although
all are adapted to the wet peatland environ-
ment. In stark contrast, only two or three pioneer
woody species occur on heavily degraded sites
(Page et al., 2009). These typically include wind-
dispersed species (e.g. Combretocarpus rotunda-
tus (Miq.) Danser), species dispersed by small ani-
mals, (e.g. Cratoxylon glaucum Korth., Syzygium
spp.) or fire-resistant species (e.g. Melaleuca spp.).
There is little formally published literature on
the selection of appropriate species for restora-
tion of tropical peat swamp forest, but a reason-
able wealth of grey literature, including confer-
ence proceedings and internal project reports.
The Kalimantan Forest Climate Partnership,
a tropical peatland restoration demonstra-
tion project supported by the United Nations
Collaborative Programme on Reducing Emissions
from Deforestation and Forest Degradation in
Developing Countries (REDD) and developed un-
der an Australia-Indonesia Government partner-
ship, has pooled available literature into a silvi-
cultural review of peat swamp forest tree species
occurring in Central Kalimantan. This is an impor-
tant stage in identifying potential transplant spe-
cies for this region while also high lighting gaps in
silvicultural knowledge that need to be addressed
in order to develop restoration best practice
(Graham, 2009). This is the first step in identify-
ing species with appropriate ecological traits for
use across the range of environmental conditions
characteristic of degraded tropical peatlands.
To date, a small proportion (< 5 percent) of the
peat swamp forest tree species have been used
in restoration trials; frequently used species are
Shorea balangeran Burck, Alstonia spathulata
Blume and Dyera polyphylla (Miq.) Steenis. The
last of these is popular because the bark can be
tapped for latex, which provides a source of in-
come for local communities. Ideally, species to be
planted should be selected based on their eco-
logical tolerance of site-specific conditions, but
since there is limited autecological knowledge of
the vast majority of tropical peat swamp forest
tree species, restoration efforts have proceeded
largely by trial and error. Peat swamp forest that
has been logged and drained is exposed to high

GENET IC CONS IDERATI ONS IN ECO SYSTEM RE STORATI ON USING N ATIVE TRE E SPECIE S
203
light levels and drought during the dry season,
but does not suffer from flooding. Other sites,
such as those closer to waterways or where the
peat surface has subsided as a result of peat
oxidation and fire, may flood regularly (Wösten
etal., 2006). Some peat swamp forest tree species
have a broad tolerance of this range of degrad-
ed conditions; Shorea balangeran and Syzygium
spp., for example, are species of riverine forest at
the edge of the peat dome and tolerant of both
high light and high water levels, making them
suitable species for planting on open areas sub-
ject to flooding. In contrast, Tetramerista glabra
Miq., a tree of mixed swamp forest on deeper
peat, will grow in degraded areas but only if
there is adequate shade, and thus is best suited to
enhancement planting or planting at the forest
edge. Other species showing high survival rates in
restoration trials include Combretocarpus rotun-
datus, Koompassia malaccensis Benth., Melaleuca
cajuputi Powell, Syzygium oblatum (Roxb.) Wall.
ex A.M.Cowan & Cowan and Tetramerista glabra.
Nursery seedlings propagated from seed or wild-
lings are commonly used as propagation material.
Seedlings are normally cultivated in nurseries for
at least six months before planting out on the de-
graded areas.
In practice, the greatest risks to the long-term
success of restoration trials on tropical peat
swamp forest to date have proved to be pro-
longed deep flooding and fire rather than poor
choice of species for planting. Efforts to overcome
seedling mortality during periods of high water
level have included constructing artificial plant-
ing mounds, while reducing fire damage requires
active fire prevention during the annual dry sea-
son when drained peatlands are at greatest risk
of ignition (Giesen, 2009; Page et al., 2009).
Community knowledge and participation
Local communities can provide knowledge of site
history and insights into the social impediments
to forest restoration. A study in Central Kaliman-
tan that employed community participation, fo-
cus groups and interviews investigated the under-
standing and attitudes of local people relating to
forest degradation, their current dependency on
and attitudes towards the forest and their hopes
for its future (Graham, unpublished data). The
study highlighted important issues that would
need to be addressed in a local forest-restoration
action plan, and demonstrated the wealth of for-
estry and ecological principles understood by the
community, their desire to be involved with res-
toration activity and their insights on how trans-
planted tree species should be selected and used.
Concluding remarks
Tropical peatland degradation is now widely ac-
cepted as a matter of international concern, and
restoration is seen as essential. Key stages neces-
sary to achieving restoration in topical peatlands
include improved knowledge of how to tackle
fire management, hydrological rehabilitation and
species selection and better assessment of the di-
verse array of secondary regeneration barriers.
Knowledge gained from long-term monitoring of
restoration sites will also be critical to develop-
ment of pathways to more efficient landscape-
scale restoration, as will be effective local com-
munity engagement.
References
Bappenas. 2010. Reducing carbon emissions from
Indonesia’s peatland. Jakarta, National Agency for
Planning and Development.
Giesen, W. 2009. Guidelines for the rehabilitation of
degraded peat swamp forests in Central Kalimantan.
Master plan for the conservation and develop-
ment of the ex-mega rice project area in Central
Kalimantan. Arnhem, The Netherlands, Euroconsult
Mott MacDonald.
Graham, L.L.B. 2009. Internal report: a literature review
of the ecology and silviculture of tropical peat
swamp forest tree species found naturally occurring
in Central Kalimantan. Palangka Raya, Indonesia,
Kalimantan Forest and Climate Partnership.

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Graham, L.L.B. & Page, S.E. 2012. Artificial bird perches
for the regeneration of degraded tropical peat
swamp forest: a restoration tool with limited poten-
tial. J. Restor. Ecol., 20(5): 631–637.
Hoscilo, A., Page, S.E., Tansey, K.J. & Rieley, O.J.
2011. Effect of repeated fires on land-cover change
on peatland in southern Central Kalimantan,
Indonesia, from 1973 to 2005. Int. J. Wildland Fire,
20: 578–588.
Miettinen, J. & Liew, S.C. 2010. Degradation and devel-
opment of peatlands in Peninsular Malaysia and in
the islands of Sumatra and Borneo since 1990. Land
Degrad. Dev., 21: 285–296.
Miettinen, J., Hooijer, A., Shi, C., Tollenaar, D.,
Vernimmen, R., Liew, S.C., Malins, C. & Page, S.E.
2012. Extent of industrial plantations on Southeast
Asian peatlands in 2010 with analysis of historical
expansion and future projection. Glob. Change Biol.
Bioenergy, 4: 908-918.
Page, S.E., Rieley, J.O. & Banks, J.C. 2011. Global and
regional importance of the tropical peatland carbon
pool. Glob. Change Biol., 17: 798–818.
Page, S.E., Graham, L.L.B., Hoscilo, A. & Limin, S.
2008. Vegetation restoration on degraded tropical
peatlands: opportunities and barriers. Paper pre-
sented at the International Peatland Symposium,
Tullamore, Ireland.
Page, S.E., Hoscilo, A., Wosten, H., Jauhiainen, J.,
Ritzema, H., Tansey, K., Silvius, M., Graham,
L., Vasander, H., J. Rieley & Limin, S. 2009.
Ecological restoration of lowland tropical peatlands
in Southeast Asia – current knowledge and future
research directions. Ecosystems, 12: 288–905.
Posa, M.R.C., Wijedasa, L.S. & Corlett, T.R. 2011.
Biodiversity and conservation of tropical peat swamp
forests. BioScience, 61: 49–57.
Wösten, J.H.M., van der Berg, J., van Eijk, P., Gevers,
G.J.M., Giesen, W.B.J.T., Hooijer, A., Idris, A.,
Leenman, P.H, Rais, D.S., Siderius, C., Silvius,
M.J, Suryadiputra, N. & Wibisono, T.I. 2006.
Interrelationships between hydrology and ecology
in fire degraded tropical peat swamp forests. Int. J.
Water Resour. Dev., 22: 157–174.
14.3. Support to food security,
poverty alleviation and
soil-degradation control
in the Sahelian countries
through land restoration
and agroforestry
David Odee and Meshack Muga
Kenya Forestry Research Institute
Sahelian countries are severely affected by
drought and desertification, with a significant
southward expansion of the desert into gum-
and resin-producing zones. Since the 1970s
countries within the sub-Saharan Sahelian re-
gion have experienced droughts that adversely
affect livestock production, agriculture and
woodlands. A regional project implemented
in 2003–2010 applied a coordinated strategy
for restoration of degraded lands to support
agrosilvipastoral activities. The development
objective of the project was to contribute to
sustainable development, food security and the
fight against desertification through the pro-
motion and integration of gum and resin pro-
duction into rural economic activities in Africa.
The immediate objective was to strengthen the
analytical and operational capacity of six pilot
countries (Burkina Faso, Chad, Kenya, Niger,
Senegal and the Sudan) to address food security
and desertification problems through the im-
provement of agrosilvipastoral systems and the
sustainable development of the gum and resin
sectors.

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Site information
The project sites were located mainly in degraded
silvipastoral sites in arid and semi-arid ecological
zones of sub-Saharan Africa. Restoration plots
ranged from two to 850 hectares, with a total
area of approximately 13 000 ha. Location of the
sites and the landscape context in each country
were as follows:
• Burkina Faso: 27 sites (no site information).
• Chad: degraded land with low shrub and
tree cover with natural indigenous species.
• Kenya: low isolated shrubs of natural Acacia
and Commiphora species in Marsabit District
(Merille, Logologo and Laisamis sites), and in
the southern rangelands in Makueni District
(Kibwezi and Kiboko sites).
• Niger: natural woodlands occurring in the
Departments of Mirriah, Aguié, Madaoua,
Say, Kollo and Plateau de Kouré.
• Senegal: isolated acacia woodlands in the
main administrative regions in Louga,
St Louis, Matam, Tambacounda and
Diourbel.
• The Sudan: in the Acacia senegal belt and
sites adjacent to forest reserves in North
Kordofan, Sennar and Blue Nile states.
Site conditions prior to restoration activities were
dry, open and highly degraded, with isolated
shrubs or woody species; some sites were devoid
of vegetation. Key sources of livelihoods for the
local communities were landscape restoration
for livestock grazing, agroforestry and provision-
ing of other ecosystem services such as fuelwood
and soil fertility improvement. Gum production
was the strategic income-generating activity and
driver for restoration activities. There was no
involvement in carbon offset schemes or activi-
ties targeted at reducing emissions from defor-
estation and forest degradation (REDD+), despite
their strong relevance.
Restoration activities
A participatory approach was adopted by the
project, with high interest and ownership shown
by local communities. Sites for restoration were
selected through participatory approaches with
local communities. Preliminary surveys and
collection of baseline information helped to iden-
tify target groups. Main interest groups for the
project were local communities, local government
departments for environment and agriculture,
FAO, the Network of Natural Gums and Resins
in Africa (NGARA) and the Inter-Departmental
Working Group on the Convention to Combat
Desertification.
Restoration activities were initiated in 2004.
The project employed a mechanized water-
harvesting technology (the Vallerani system) that
digs microbasins while ploughing degraded soils,
and used this to develop acacia-based agrosil-
vipastoral systems and reverse land degradation.
The technology has good potential to harvest and
store water for associated vegetation (trees and
crops) long after rainfall. The mechanized system
is efficient and can be used to prepare swathes
of degraded lands within a shorter time than can
be achieved with local traditional systems, which
employ hand hoes or oxen-powered ploughs. This
system has been used successfully in rehabilita-
tion of degraded lands in North Africa.
Native tree species were preferred because of
their adaptation to the local environment, avail-
ability of planting material and awareness and
use of the species by local communities. The key
native species used across the sites was the gum
arabic tree, Acacia senegal (L.) Willd., because of
the commercially valuable gum it produces and
thus the potential for income generation. In ad-
dition to producing gum, it is a nitrogen-fixing
tree, enriching the soil through the nitrogen-rich
litter it produces, producing fodder during the
dry season and stabilizing the soil. Other species
used were:
• the drought-tolerant timber tree species,
Melia volkensii Gürke (used at Kibwezi and
Kiboko sites, Kenya)
• Acacia seyal Delile, Acacia nilotica (L.)
Delile, Bauhinia rufescens Lam. and Ziziphus
mauritiana Lam. in Niger
• Acacia mellifera (M.Vahl) Benth. in Senegal
• Acacia seyal, Acacia mellifera, Acacia nilotica
and Albizia spp. in the Sudan.

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Some exotic or non-autochthonous species
were also used in agroforestry systems, namely
Jatropha curcas L. (biofuel) and Mangifera indica
L. (mango) for income generation in Kenya and
Azadirachta indica Adr. Juss. for medicinal pur-
poses and fodder in the Sudan.
Seeds of Acacia senegal were collected from
local populations or the nearest possible sourc-
es using standard procedures to ensure ge-
netic diversity. Seedlings were raised in local or
community- based nurseries. In Niger, Acacia sen-
egal seeds from Kordofan provenance were also
used but it is not clear in which site they were
planted.
Acacia senegal trees were planted in rows
along the microbasins with an interrow spacing
of 4–6 m. Annual crops such as millet and maize
were planted between the tree rows to create
agroforestry systems. The plots were located near
villages and ranged in size from five to 100 ha.
Monitoring and experiences gained
Management and monitoring of the trees were
carried out by the local communities. Trees were
maintained and intercrops planted annually dur-
ing rainy seasons. Performance assessment was
based on survival of seedlings and growth and
yields of annual intercrops. Seven years after
planting, Acacia senegal and other species were
reported to have survived and established suc-
cessfully, but no statistics are available.
The overall conditions of the sites are also re-
ported to have been restored, including improved
soil health and fertility and vegetation cover. The
local communities are using the restored land
for agroforestry activities. Agricultural inter-
crops include sorghum, beans, pigeon pea, cow-
pea, green gram, pearl millet, watermelon and
maize. The local communities are in accord with
the need to adopt an integrated approach to ad-
dressing the problems of land degradation and
poverty. Technological aspects of production and
processing should be linked to marketing, while
livelihood diversification should be introduced to
ensure sustainability. The participating countries
were highly receptive to the experiences of reha-
bilitating degraded lands and rationalizing the
production of gums and resins as the key driver to
restoration in the drylands.
Overall, the restoration activity was deemed
a success, considering the good establishment
and survival of the trees planted, as well as in-
tegration and adoption of agroforestry activities
within the restored areas by the normally pasto-
ral communities. Problems encountered included
prolonged droughts, especially during planting
and establishment, damage to planted trees by
livestock and wild animals, and the communal
land-tenure system, which affects ownership
and responsibility over restored areas. Income-
generating activities should be identified, pro-
moted and supported to offer incentives to the
local communities to backstop the sustainability
of the restoration activities.
There are plans to upscale the restoration pro-
gramme, and to link to new initiatives such as
the Great Green Wall for the Sahara and Sahel
initiative supported by the African Union and the
European Union.

GENET IC CONS IDERATI ONS IN ECO SYSTEM RE STORATI ON USING N ATIVE TRE E SPECIE S
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14.4. The use of native species
in restoring arid land and
biodiversity in China
Lu Qi and Wang Huoran
Chinese Academy of Forestry, Beijing,
China
Generally, desertification is mainly caused by
climate change and human activities. Overgraz-
ing, inappropriate agricultural use and exces-
sive collection of fuelwood and medicinal herbs
reduce vegetative cover, rangeland capacity and
biological diversity, and expose the soil to erosion
(CCICCD, 1997). China is one of the countries
worst affected by desertification. The areas most
seriously affected are found in the northern,
northeastern and northwestern parts of the coun-
try. The Chinese Government, scientists and local
communities have made great efforts to restore
land and biodiversity affected by desertification.
The climate and soils of northwestern China
pose extremely severe challenges to the develop-
ment of vegetation. The climate is characterized
by frequent, strong winds, very low rainfall and
high evapotranspiration. Sandstorms are destruc-
tive and hamper agriculture, forestry and animal
husbandry. Most of the arid areas experience a
typically continental climate, with annual precipi-
tation less than 250 mm. Temperature changes
dramatically, daily, seasonally and spatially; for
example, the maximum temperature recorded in
the Xinjiang region is 47.5 °C in Turpan and mini-
mum temperature is –43 °C in the Zunger Basin.
Desert soils are relatively undeveloped. Soil tex-
ture is usually coarse sandy loam, sand and gravel.
Organic matter and nitrogen contents are low,
but abundant soluble mineral salts are accumu-
lated at the soil surface, resulting in high pH and,
in many cases, large areas covered by saline crusts.
Plant genetic resources in the arid areas
of China
It is estimated that there are over 1000 species of
trees and shrubs in the arid and semi-arid areas
in northern China (Li, 1992). Many of these, es-
pecially the shrubs, have been used for environ-
mental improvement, dune stabilization, affores-
tation, and production of food and other human
requirements.
A number of tree species have been used in
dune plantings in the arid and semi-arid regions
of China. Pinus tabuliformis Carrière, P. sylvestris
var. mongolica Litv., Platycladus orientalis (L.)
Franco, Sabina chinensis (L.) Antoine, Juniperus
rigida Siebold & Zucc. and Picea crassifolia Kom.
are native coniferous species adapted to a cold
and dry environment with annual rainfall of less
than 400 mm (Zhao, 2005). Picea crassifolia, for
example, is naturally distributed up to the tree-
line and is critical for watershed management
in the glacier-covered Qilianshan Mountains in
Gansu Province. Sabina chinensis and S. vulgaris
Antoine are useful for ground cover to protect
soil from erosion on harsh sites.
Only a few indigenous hardwoods can survive
in the harsh environment of northwestern China.
Populus simonii Carr. and P. alba pyramidalis
(Bunge) W.Wettst. are mostly used for establish-
ing windbreaks or shelterbelts on agricultural
land or around oases (Zhao, 2005). Of the spe-
cies used in shelterbelts established in the 1970s
in northwestern China, only Ulmus pumila L.
survived drought and insect pests. Other hard-
woods, such as Sophora japonica L., Salix matsu-
dana Koidzumi and Ailanthus altissima (P. Mill.)
Swingle are also often planted as shade trees
around villages in northwestern China.
The diversity of shrubs is, however, relatively
richer than that of trees (SSG, 2003). Efforts to
maintain biological diversity and reclaim land

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have in recent years increasingly used shrub spe-
cies as well as trees. Shrub communities have
important ecological functions in conservation
of soil, water and biodiversity. Most indigenous
shrubs are better adapted to dry soil, poor nutri-
ent availability and temperature extremes than
are trees. Moreover, several shrubs have high eco-
nomic value as medical plants (e.g. Sophora alo-
pecuroides L. and Lycium chinense P. Mill). There
is great potential in the utilization of the genetic
resources of shrub species evolved in the arid and
cold environment in northern China.
A number of shrub species have been used
to stabilize sand dunes and for other ecological
and economic uses. Caragana korshinskii Kom.,
a leguminous shrub, is used extensively for dune
stabilization, bioenergy, fodder, and soil and wa-
ter conservation in arid and semi-arid areas in
the middle reaches of the Yellow River. It is xero-
phytic, capable of prolific coppicing and tolerant
of animal browsing. It is native to northwestern
China, where it grows in areas with annual rain-
fall of approximately 200 mm and air tempera-
tures ranging from 70 °C above the sandy ground
in summer to −30 °C in winter. Other species in
the genus Caragana with similar biological traits,
such as C. intermedia Kuang & H.C. Fu and C.
microphylla Lam., are also commonly used for
land reclamation and dune stabilization (Li and
Bao, 2000). Hedysarum scoparium Fisch. & C.A.
Mey. and H. mongolicum Turcz., also leguminous
shrubs, are more adapted to moving dunes, semi-
fixed land or fixed sandy land in areas with an-
nual rainfall between 150 and 300 mm. They are
tolerant of wind erosion, with deep root systems
and strong coppicing ability. Hedysarum mongoli-
cum colonizes sites by producing thick suckers
around main stems. Glycyrrhiza uralensis Fisch.
(Chinese liquorice) is a perennial herb or semi-
woody shrub of the Paplionaceae family that has
been successfully grown in the desert, providing a
good income to farmers as a traditional Chinese
medicinal plant.
Ammopiptanthus mongolicus (Kom.) S.H. Cheng
is unusual in that it is an evergreen species in the
cold desert area. It forms symbioses both with nod-
ule-forming bacteria and mycorrhizal fungi, which
together help it fix nitrogen and take up other
nutrient elements in the soil. This species grows
well in soil with pH values of 8 to 9 (SSG, 2003).
Tamarix chinensis Lour., Hippophae rhamnoides
L., Elaeagnus angustifolia L., Salix psammophila
Z.Wang & Chang Y.Yang, Haloxylon ammoden-
dron (C.A. Mey) Bunge ex Fenzl and many other
species are used to stabilize moving sands and es-
tablish shelterbelts as understorey or living hedg-
es. Research in genetic improvement with H. rham-
noides has been carried out over the last 30 years
and good seed resources have been established
and utilized (Lian and Chen, 1992; SSG, 2003).
Methodology of tree planting in arid areas
Generally, three approaches are used to estab-
lish plantations of woody plants in China (Zhao,
2005). The first approach is to plant trees by hand
or machine. This is commonly used to establish
plantations, especially for shelterbelts. Seedlings
or cuttings used are normally of a large size; for
instance, it is quite common to use cuttings of
poplars over 2 metres in length. The seedlings or
cuttings may be watered and fertilized at plant-
ing time to assist survival. The second approach is
to sow seed by airplane. This is the usual method
when dealing with large areas of sand dunes or
wild lands, usually during the rainy season. Timing
is critical for aerial sowing. In the west of north-
eastern China and the Loess Plateau, plantations
of pines, larches or shrubs are sometimes estab-
lished by this method. Approximately 30–50 per-
cent of the seeds sown germinate and establish
(Qi, 1999). The third method is to close a mountain
area or sandy land to human use or to mitigate
animal pressure. This allows natural vegetation to
re-establish and the land to recover. It takes many
years to restore vegetation this way in the arid
and semi-arid areas, but the approach requires lit-
tle in terms of direct inputs. This concept of land
management has been used over a long period
and has been shown to be efficient and economi-
cal (Zhao, 2005).
Prior to planting, sites are usually prepared to
reduce runoff of water and increase infiltration.

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Many different methods have been developed
and used depending on the situation. However,
two methods of site preparation are widely used.
On shallow slopes or relatively flat sites a 1×1m
square pit is dug to a depth of 20–30cm and a
seedling is planted in the centre. On steeper
slopes or stony sites a fish-scale-shaped pit, or
level terrace, is made and a seedling is planted
at the lowest point of the pit. On sandy land or
mobile dunes seedlings or cuttings are planted on
unprepared land.
Straw checkerboards and barriers have been
shown to improve establishment of trees or shrubs
planted on dunes (Fan, Xiao and Liu, 1999). Newly
planted seedlings or cuttings are easily damaged
by wind or buried by sand. To prevent this, artifi-
cial barriers and straw checkerboards are estab-
lished prior to planting trees in order to reduce
wind velocity near the ground and slow down the
movement of sand. Sand barriers can also prevent
sand from drifting inside fixed-sand areas. Wheat
or rice straw for making checkerboards is readily
available in rural areas. Artificial sand barriers are
usually made with shrub stems or other similar
materials. Trees are planted in each grid of the
checkerboards. Plantations usually take two or
three years to establish after planting; the straw
checkerboards and artificial barriers help stabi-
lize dunes while the trees are establishing. This
method has been widely used for dune plantings.
Recently, drip irrigation systems have been more
widely used to maintain plantations in extremely
dry areas, in some cases where annual precipita-
tion is not more than 400 mm (Xu et al., 1999). Such
systems are very expensive to set up, but are highly
efficient and save water compared with traditional
irrigation systems. Drip systems have, for example,
been used in the roadside planting project along
the desert highway in Takelamagan, Xinjiang
Uyghur Autonomous Region, and a plantation
project in the vicinity of Yinchuan City, Ningxia Hui
Autonomous Region. Research is also being carried
out to identify plant species and accessions with
lower water requirements. Competition for water
between trees and agricultural crops must be mini-
mized or avoided if possible.
Figure 14.3.
Planting methods (left) and species (right) used in arid land restoration in China

THE STATE OF THE WORLD’S FOREST GENETIC RESOURCES – THEMATIC STUDY
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210
Lessons learned from the past
Species selection is a key to the success of biodi-
versity recovery or land reclamation. In the past,
too much emphasis was given to planting tree
species in monocultures, which led to poor ad-
aptation and problems with pests and diseases.
One or very few clones of poplars have com-
monly been planted across landscapes without
having tested the clone prior to mass planting.
Some poplars, such as Populus simonii and P. alba,
even though native to northwestern China, are
not suitable for areas receiving less than 400 mm
annual rainfall, yet shrub communities develop
well under such conditions. More shrub species
are now being used to stabilize dunes, for bioen-
ergy production and for conserving soil and wa-
tersheds in the dry areas in northwestern China.
Another change in approach is to use more fruit
and nut trees to support the market economy.
Chinese chestnut (Castanea mollissima Blume),
Siberian apricot (Prunus armeniaca L.) and pear
(Pyrus betulaefolia Bunge), for example, are tol-
erant of harsh environments and are able to pro-
duce nuts and fruits under such conditions (SSG,
2003; Zhao, 2005).
Insect damage has in recent years drawn more
attention to the management of dune planta-
tions, since it has become more serious in the dry
environment in northern China. Species diversifi-
cation is effective in controlling insect damage:
a single species or clone must not be planted
over large areas. Poplars are easily damaged by a
number of insects and diseases, such as the Asian
long-horned beetle (Anoplophora glabripennis),
poplar grey spot (Coryneum populineum) and
poplar leaf rust (Melampsora laricis-populina)
(FAO, 2007). Compared with poplars, Ulmus pumi-
la (Manchurian elm) and Salix matsudana are at-
tacked much less by insects. In some counties of
the Ningxia Hui Autonomous Region, among
shelterbelts that were established in 1970s, only
those established with elms remain. From an eco-
logical point of view it is better not to establish
windbreaks using tall tree species. Shrubs such as
Caragana spp. and Hippophae rhamnoides would
be a good alternative.
Grazing animals and fuelwood collection pose
problems in the management of dune planta-
tions, but this can be addressed by increasing the
awareness of local residents of the environmental
importance of shelterbelts and other plantings on
dunes.
Changes in land tenure in recent years have im-
plications for land restoration work in northern
China. In the past, windbreak systems were es-
tablished in very regular networks extending over
long distances. However, this is no longer possible
because agricultural land has been divided into
small lots, each owned by an individual. Under
the new system farmers are entitled to grow any
tree species they wish. Choice of species is mostly
driven by market demand. The government can
lead planting trends only by adopting appropri-
ate policies and providing technical and financial
help.
Greenness, a new concept to express coverage
of vegetation in a given geographic area, has
been proposed as a replacement for the concept
of forest coverage. Greenness can be calculated
by remote sensing from satellite images. It is ar-
gued that greenness is a better measure of veg-
etation status than forest coverage since forests
cannot establish or develop well in the arid and
semi-arid areas, and shrub or grass communities
also have environmental significance. This con-
ceptual shift will lead to change in the strategy
of land reclamation, especially in terms of species
composition and shelterbelt or stand structure.
Disturbed shrub communities and grassland will
receive more attention for conservation and man-
agement in arid areas (Shan, 2007).
Research gaps identified
It is foreseen that environmental issues, espe-
cially those related to climate change, will be-
come more serious in the near future, affecting
human living conditions and social and economic
development. In particular, sustainable social de-
velopment can be expected to be severely con-
strained by the decline in availability of biological
resources. Obviously, arid and semi-arid areas do
not hold potential for the development of com-

GENET IC CONS IDERATI ONS IN ECO SYSTEM RE STORATI ON USING N ATIVE TRE E SPECIE S
211
mercial forestry, and therefore the focus should
be on rehabilitation and sustainable manage-
ment of arid land resources through biological
approaches. Research and development in the
conservation of biodiversity and land reclamation
should emphasize the exploitation of genetic re-
sources of native species, especially shrubs and
perennial herbs with economic values. Water
availability is a critical factor for species survival,
and plant species’ water-saving strategies should
receive more attention in research programmes.
References
CCICCD (China National Committee for the
Implementation of United Nations Convention
to Combat Desertification). 1997. China country
paper to combat desertification. Beijing, Secretariat
of the CCICCD, Forestry Publishing.
Fan, H., Xiao, H. & Liu, X. 1999. Techniques for estab-
lishing straw checkerboards, sand barriers and for
revegetating sand dunes. In CCICCD, ed. Traditional
knowledge and practical techniques for combating
desertification in China, pp. 116–123. Beijing, China
Environmental Science Press.
FAO (Food and Agriculture Organization of the
United Nations). 2007. Overview of forest pests,
People’s Republic of China. Working paper FBS/13E.
Rome.
Li, J. 1992. Study on the genetic resources of woody
plants in “the three north” areas of China. J. Forest.
Res. (China), 5(special issue): 1–11.
Li, Z. & Bao, Y. 2000. Study on changes of population
pattern and inter-species relationship of Caragana
in Inner Mongolia steppe and desert region. J. Arid
Land Resour. Environ., 14(1): 64–68.
Lian, Y. & Chen, X. 1992. Ecogeographic distribution
and phytogeographic significance of Hippophae
rhamnioides. J. Plant Taxon., 30(4): 349–355.
Qi, J. 1999. Air-seeding for afforestation and rangeland
management in desert and the Loess Plateau. In
CCICCD, ed. Traditional knowledge and practical
techniques for combating desertification in China,
pp. 102–109. Beijing, China Environmental Science
Press.
Shan Zhiqiang. 2007. A discussion on “greenness” and
“forest coverage”. Chinese Nat. Geogr., General
issue No. 564: 24.
SSG (Scientific Study Group of Baijitan National
Nature Reserve, Lingwu County, Niningxia Hui
Autonomous Region). 2003. Ningxia Lingwu
Baijitan National Nature Reserve Scientific Survey
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saving irrigation technologies in arid areas. In
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techniques for combating desertification in China,
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Zhao, L., ed. 2005. Report of the assessment at the
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14.5. Using native shrubs to
convert desert to grassland
in the northeast of the
Tibetan Plateau
Yang Hongxiao1 and Lu Qi2
1 College of Resources and
Environment, Qingdao Agricultural
University, China
2 Institute of Desertification Studies,
Chinese Academy of Forestry, China
The Tibetan Plateau lies at an average of more
than 4500 m above sea level. The climate is much
colder and drier on the northern slope than on
the southern one. Mean annual temperature var-
ies from −5.8 °C to 3.7 °C and annual precipitation
is usually less than 400 mm. Such conditions ad-
versely affect plant growth and maintenance of
vegetation. Adding to this, overgrazing and poor
farming practices have destroyed the vulnerable
vegetation. Desertification of former grassland
intensified in recent decades and secondary de-
serts increased. Secondary deserts look like pale
scars on the grassland, where the sand is bare and
exposed to the wind. When strong winds blow,
such sand moves easily to form sand drifts and
create sand storms. The government and local
people are anxious to revegetate these degraded
areas to improve environmental health and sup-
port sustainable development. However, this is
not easy; drifting sand restricts plant coloniza-
tion. Fortunately, researchers have found that
native shrubs can help solve this problem (Yang
et al., 2006).
Research results
Shazhuyu is a village in the northeast Tibetan Pla-
teau (100°13’53.97”E, 36°14’08.41”N), with an-
nual precipitation of about 280 mm and mean an-
nual temperature of about 5 °C. There are many
patches of secondary desert around the village.
Experiments were started in 1958 to investigate
ecosystem restoration. These demonstrated the
following:
• The secondary deserts are restorable
because the climate is not as dry as in
primary deserts, such as Taklimakan in
Xinjiang, China.
• Frequent sand drift driven by strong wind
constrains the process of grass colonization
and ecosystem restoration.
• Transplanting native shrubs into the
secondary desert patches reduces sand drift,
allowing grass to quickly colonize the area.
• With the help of the shrubs, grass
communities develop smoothly over a
period of several years.
• Once the grass communities are strong
enough to persist by themselves, the shrubs
begin to disappear from well-restored
ecosystems that have become grassland with
high vegetation cover.
• The grassland can persist once the shrubs
have gone because most of the grasses are
perennial, stabilizing the sand with their
roots, stems and leaves even in winter.
• Upon reaching this stage, the ecosystem can
tolerate some grazing.
• If grazing is well managed, desertification
need not occur.
Thus, native shrubs can play an important role in
converting secondary desert to grassland (Yang
et al., 2006; Lu et al., 2009).
Practical application
Artemisia ordosica Kraschen. and Caragana
korshinskii Kom. are two common species of na-
tive shrubs in northwest China, mainly appear-
ing in arid deserts or on sandy land in the north-
east of the Tibetan Plateau and the west of the
Mongolian Plateau (Yang et al., 2006; Lu et al.,

GENET IC CONS IDERATI ONS IN ECO SYSTEM RE STORATI ON USING N ATIVE TRE E SPECIE S
213
2009). Shazhuyu falls within their natural distri-
bution, and they have been found to grow well
in secondary deserts in the area. Researchers in
charge of ecosystem restoration selected them as
the main species to stabilize sand and facilitate
ecosystem restoration. Biologically, they are well
adapted to desert conditions. Nevertheless, they
spread through the deserts very slowly. Harmful
sand drift greatly slows the spread of A. ordosica
populations, although their tiny seeds are read-
ily dispersed by wind. Spread of C. korshinskii is
limited by the plant’s heavy seeds, which are not
dispersed by wind (Yang et al., 2009). The slow
natural spread of the two species limits their
contribution to ecosystem restoration. To over-
come this, researchers developed a method to
transplant saplings or mature individuals of A. or-
dosica and C. korshinskii or to sow their seeds into
bare deserts. Researchers also adopted two other
methods using enclosures and mechanical barriers
(Lu et al., 2009). Enclosures protect target deserts
from grazing, while mechanical barriers made of
clay, dry straw, dry twigs or rock fragments pre-
vent sand from drifting. The two methods signifi-
cantly increase the survival rate of A. ordosica and
C. korshinskii. To some extent, they can also facili-
tate grass colonization on the deserts, but this ef-
fect is not as prominent as that of mature shrubs.
Application to arid-ecosystem restoration
Using these methods, researchers have succeeded
in converting some sand deserts back to grassland
dominated by the perennial grass Leymus seca-
linus (Georgi) Tzvel., which is palatable to yaks,
sheep, cattle and horses. This conversion takes
about 50 years. In contrast, untreated areas are
still bare and consist of mobile sand, showing no
signs of restoration.
Experiences from Shazhuyu set a good exam-
ple for other arid regions in the Tibetan Plateau
and elsewhere, where native shrubs play an
important role in restoring secondary deserts.
Throughout the Tibetan Plateau, native shrubs
that adapt well to a sandy environment include
A. ordosica, C. korshinskii, Artemisia xigazeensis
Ling et Y.R. Ling, Sophora moorcroftiana (Benth.)
Baker, Tamarix ramosissima Ledeb., T.hohenack-
eri Bunge, T. laxa Willd., T. elongate Ledeb.,
Hippophae rhamnoides Linn., Nitraria tanguto-
rum Bobr., N. sibirica Pall., Lycium ruthenicum
Murray, L.dasystemum Pojarkova, L. cylindricum
Kuang, Haloxylon ammodendron Bunge and
Ephedra przewalskii Stapf (Chen and Liu, 1997;
Liu, Gao and Jiang, 2003). Different regions will
have different native shrubs, and local people
should be asked to recommend the most promis-
ing for ecosystem restoration in their locale.
References
Chen, H. & Liu, Z. 1997. The characteristics of the veg-
etation composition of Jiangdang and its adjacent
area at Rigaze County in Tibet. J. Desert Res., 17:
63–69.
Liu, Z., Gao, H. & Jiang, D. 2003. Issues concerning
shifting sand stabilization at Zigaze, Tibet. J. Desert
Res., 23: 665–669.
Lu, Q., Wang, X., Wu, B. & Yang, H. 2009. Can
mobile sandy land be revegetated in the cold and
dry Tibetan Plateau in China? Front. Biol. China, 4:
62–68.
Yang, H., Lu, Q., Wu, B., Yang, H., Zhang, J. & Lin,
Y. 2006. Vegetation diversity and its application in
sandy desert revegetation on Tibetan Plateau. J. Arid
Environ., 65: 619–631.
Yang, H., Wu, B., Deng, H. & Lu, Q. 2009. Mechanism
of natural revegetation in sandy lands of agro-pas-
toral ecotone in north China and its application to
sand control. World Forest. Res., 22(4): 29–33.

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14.6. Reforestation of highly
degraded sites in Colombia
Luis Gonzalo Moscoso Higuita
FORESTPA S.A.S., Colombia
The method presented here was developed by
the author based on observations of natural suc-
cession taking place on land that was abandoned
by rural communities as a consequence of armed
conflicts. Study sites included land that had been
highly degraded by agricultural practices, cattle
breeding and mineral extraction. Natural succes-
sion was thoroughly studied on the sites, partic-
ularly with respect to the dominance of certain
species, the abundance and diversity of weeds
in the understorey, the formation of soil organic
matter, the appearance of micro-organisms and
recolonization by natural fauna.25
Based on these observations, the author de-
veloped strategies to stimulate natural processes
relating to wind-, water- and animal dispersal of
seeds. This involved selection of the site, estab-
lishment and maintenance of fences, selection of
elite mother trees as candidates for future seed
collection, propagation in transitory nurseries,
and establishment and maintenance of planta-
tions. Full details of the approach are presented
in Moscoso Higuita (2005). So far, this method
has been applied on 1292 ha of land, of which
751.79 ha are degraded by gold mining, intensive
25 The Cáceres I and II projects, quoted as a case study, account
for 751.79 hectares, located at 7°34’38.6’’N, 75°19´47.4’’W, and
are characterized as tropical rainforest according to the Holdridge
classification.
cattle-raising and unsustainable cultivation of ag-
ricultural crops in the Colombian municipality of
Cáceres, Antioquia.³
Methodology
Various forest restoration strategies were devel-
oped in accordance with the type and degree
of soil degradation, including intensive and ex-
tensive cattle-raising, unsustainable agricultural
practices and opencast gold mining. For this
reason, the methodology is initially described
through a two-track approach, whereby some
common themes such as social and administrative
aspects are described jointly.
The starting point for the development of this
methodology involved the traditional knowl-
edge of the ethnic groups living in the project
area. Colombia is a highly diverse country, both
in terms of biota and societies of different ori-
gins, including Afro-Colombians, indigenous peo-
ple, mestizos and mulattos. Valuable traditional
knowledge and practices were combined with
external technical and administrative knowl-
edge, always in a context of respect and equity
in terms of goods and services. Second, it was
necessary to know what environmental resources
were present in the buffer zone surrounding the
degraded area. This involved selection of target
species, understanding existing opportunities to
connect vegetation fragments, identification of
superior or elite mother trees for collection of
germplasm, and assimilation of basic information
on soils, vegetation, hydrology, fauna and micro-
organisms, among other aspects. Once sufficient
knowledge had been collected to allow charac-
terization of both the degradation state and the
available natural resources, the forest restoration
strategy was put into practice.
The project area was divided into zones accord-
ing to prevailing conditions of soil degradation
using tools such as aerial photography and maps
covering several decades. This was done to get
a better understanding of the sites’ history.26 In
26 Maps and aerial photographs were obtained from Instituto
Geográfico Agustín Codazzi (IGAC).

GENET IC CONS IDERATI ONS IN ECO SYSTEM RE STORATI ON USING N ATIVE TRE E SPECIE S
215
addition, a topographic survey was undertaken
with the help of a GPS to determine the location
of all landforms, including paths, roads, forest rel-
icts, crop residues, constructions and others. Based
on this information, the landscape was divided
into plantation areas that were isolated from one
another by means of living hedges consisting of
fast-growing and vigorously rooting plant species
(planted by means of vegetative reproduction,
and as grass seedlings) and barbed-wire fences.
Soil samples were collected and characterized
chemically and physically with the help of special-
ized laboratories. Based on these analyses, com-
panies producing organic fertilizers prepared a
special fertilizer mix based on compost to restore
nutrient deficiencies (Table 14.1). Fungi and bac-
teria were added to the soil to promote biodeg-
radation of organic materials and increase avail-
ability of nutrients. Ultimately, this is expected
to lead to an improved nutritional balance and
improved soil-water retention.
Standard protocols were followed to iden-
tify good seed sources and planting material for
production of seedlings in transitory nurseries.
Selection was based on:
• research conducted at other locations with
similar edaphic and climatic conditions;
• volumetric efficiency of the species;
• nutritional requirements of the species;
• geographic distribution of the species;
• adaptive capacity of foreign or exotic
species;
• occurrence of the species in the region;
• economic, cultural and ecological value;
• ease of accessibility of seeds, high
germination percentage and broad genetic
base; and
• resistance to pests and diseases.
In the Cáceres II project, plantations were estab-
lished with seeds from seed orchards and from
elite trees selected from areas near the restora-
tion site and from other locations that were part
of the species’ natural distribution. A register
was assembled of every elite tree, containing
details of its morphology, phytosociology and
health, and the location and topography where
it was found. For each of the selected species,
protocols were followed for germination and
growth in transitory nurseries. Planting bags
were filled with a mix of mineral soil, fine sand
and the fertilizer mix detailed in Table 14.1, in a
ratio of 3:1:1. The substrate was further enriched
with sprays composed of fungi and bacteria that
were produced and supplied by a laboratory spe-
cialized in biotechnology. The bacteria included
in these sprays belonged to the genera Bacillus,
Pseudomonas, Azotobacter, Azospirillum, Beijer-
inckia, Nitrosomonas, Nitrobacter, Clostridium,
Thiobacillus, Lactobacillus and Rhizobium. These
are among the most important bacterial genera
involved in the degradation of organic and inor-
ganic compounds, and hence promote availability
of nutrients to plants. The most important genera
of soil actinomycetes used for enhancing plant
nutrient uptake were Streptomyces, Nocardia, Mi-
cromonospora, Thermoactinomyces, Frankia and
Actinomyces. Finally, mycorrhizal fungi from the
genera Rhizophagus (ex Glomus), Acaulospora,
Entrophospora and Gigaspora (which enter in
symbiosis with the roots; Delgado Higuera [1999])
were added to the substrate. The same substrate
was used in the planting bags, while transplant-
ing the germinated plants, and as fertilizer in the
established plantation.
TABLE 14.1.
Composition of fertilizer mix applied
Component Percentage of
content
Chicken manure 40.0
Mushroom compost 20.5
Mycorrhizae 18.0
Phosphoric rock 10.0
Gypsum 7.0
Magnesium sulphate 3.0
Agrimins 1.0
Humiplex 50G 0.5
Total 100

THE STATE OF THE WORLD’S FOREST GENETIC RESOURCES – THEMATIC STUDY
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216
Light vehicles such as pickup trucks were used
to transport plant material to the planting sites.
Within sites, seedlings were transported on the
backs of horses and mules to reduce soil compac-
tion. Sites were weeded manually or, in some cas-
es, using mowers, especially on sites with a grass
cover. During weeding some isolated trees were
left standing and were pruned; these act as natu-
ral seed sources, attract dispersal agents and serve
as biological corridors. Bordering groves, marshes,
streams, creeks and rivers were left untouched.
This approach is categorized as enrichment; a
combination of sparing interesting species al-
ready available on site during initial weeding ac-
tivities and planting new trees in clearings.
The planting design varied according to the
number and distribution of trees already present
in the sites. On land without any remaining trees,
seedlings were planted at a spacing of 3×3m. At
sites that were previously used for livestock ranch-
ing and agriculture, holes of 0.4 × 0.4 × 0.4 m
were dug and filled with crumbled soil to counter
soil compaction. In soils degraded by alluvial gold
mining, holes were dug up to 70 cm deep to reach
deeper, less or non-polluted soil layers. On sites
with a grass cover, areas of 1-m diameter around
each plantlet were weeded. In soils that were pol-
luted from mining this was not necessary as there
was generally no vegetation left. Soil from each
planting hole was mixed with 120 g of fertilizer
mix (Table 14.1), 5 g of Terracotem (soil condi-
tioner with hydro-retainer) and organic–mineral
fertilizer. Seedlings that failed to establish were
replanted 45 days later, but mortality did not ex-
ceed 10 percent.
Maintenance depended on the nature of the
weeds emerging. Naturally regenerating tree,
shrub or herbaceous plants of interest were iden-
tified and left standing. Degree of interest was
based on wood quality, proximity to the plantlets,
ecological value, growth, abundance, incidence
and presence in the plantation. On sites with low-
er vegetation cover, regeneration was assisted or
promoted by dispersing seeds of pioneer species
(seed collected from marked trees, seed stands
and clonal seed orchards). Assisted regeneration
refers to methods that enhance the in situ regen-
eration of plant species, either by promoting the
health and growth of plants already available on
site or by distributing seeds of desired species.
The species that were initially established in the
Cáceres II project were acacia and teak (Acacia
mangium Willd. and Tectona grandis L. f., both
exotic species), Colombian mahogany (Cariniana
pyriformis Miers), red cedar (Cedrela odorata
L.), roble (Tabebuia rosae DC), Brazilian firetree
(Schizolobium parahyba (Vell.) S.F. Blake), amami
gum (Hymenaea courbaril L.), choiba or almendro
(Dipteryx oleifera Benth.), melina (Gmelina arbo-
rea Roxb., another exotic), downtree (Ochroma
pyramidale (Cav. ex Lam.) Urb.), Spanish elm
(Cordia gerascanthus L.), Devil’s ear (Enterolobium
cyclocarpum (Jacq.) Griseb.) and ceiba tolúa
(Bombacopsis quinata (Jacq.) Dugand), account-
ing for a total of 359830 individuals.
In the 190 monitoring sites that have been in-
stalled so far using this technique, 122 tree species
in various families and genera have been shown
to have good CO2 sequestration capacity. Among
these are species that produce tannins, dyes, elas-
tomers, herbal medicines, food and other prod-
ucts. These trees have contributed to attracting
a large number of mammals, birds, insects, bats,
amphibians and reptiles back to areas that were
once part of their natural habitat (Asorpar Ltda
and South Pole Carbon, 2011).
Restoration of soils degraded by gold mining
Over 100 ha of soils degraded by alluvial gold
mining have been reforested and restored by the
present project.
Opencast gold mining had been executed on
the sites of interest (located in the Lower Cauca
region of Antioquia Department, Colombia) us-
ing tools ranging from bulldozers, excavators
and tractors to small tools such as sieves, shovels
and picks. Gold miners typically mixed the or-
ganic layer with the rest of the soil, in most cases
digging down to the bedrock. This mixture was
then washed with water under pressure to pass it
through a funnel and a canoe-shaped sieve con-
taining quicksilver (mercury) to catch available

GENET IC CONS IDERATI ONS IN ECO SYSTEM RE STORATI ON USING N ATIVE TRE E SPECIE S
217
gold particles. The contaminated and sediment-
loaded water then typically flowed into creeks,
streams and rivers.
The mining activities left behind gullies, slag
heaps (sterile accumulation zones) and lagoons
(Figure 14.4). However, even under such degrad-
ed conditions life can show its power to regener-
ate ecosystems seemingly lost, with the appear-
ance of rustic and fast growing species such as
downtree, pedro tomín (Cespedesia macrophylla
Seem.), yagrumo macho (Schefflera morototoni
(Aubl.) Maguire, Steyerm. & Frodin), yarumos
(Cecropia sp.), capulin (Trema micrantha (L.)
Blume) and jacaranda (Jacaranda copaia D.Don).
Several steps were followed in the restoration
of these sites, beginning with landscaping, fol-
lowed by subsoil preparation and finally fertiliz-
ing and planting.
Landscape
At this stage the objective was to establish a
topography as similar as possible to that of the
surrounding area. The waste piles left behind
by mining activities, consisting of stones, sand,
gravel, silt and, in some cases, organic matter de-
posited on the sides of gullies, were levelled using
a bulldozer. This also improved the soil texture
and structure (Figure 14.5).
Earlier projects used heavy machinery (a
Caterpillar D-6 bulldozer and a CASE 1450B), but
more recent projects used lighter and more ver-
satile machinery (a Caterpillar D4 bulldozer) to
reduce soil compaction.
Subsoil preparation
The bulldozer was equipped with a 0.7-m-long
hydraulic hook to make grooves while levelling
the soil. This broke up compacted soil layers and
improved soil physical properties. This encourages
root development and improves the establish-
ment and growth of the trees and shrubs planted.
Fertilizing and planting
Fertilizers were applied in the grooves in accord-
ance with recommendations derived from the soil
analyses. After this, the trees were planted in the
grooves.
Results
The post-restoration project area can be charac-
terized as a polyculture dominated by native spe-
cies. The exotic species Acacia mangium and teak
were planted on soils where abiotic conditions
impeded the growth of other flora, in order to
create favourable soil conditions. Only 21 species
Figure 14.4.
Initial state of the restoration site as left by opencast gold-mining activities

THE STATE OF THE WORLD’S FOREST GENETIC RESOURCES – THEMATIC STUDY
PART 3
218
were found on the project area prior to restora-
tion; currently, over 100 native species are found
in the plantations, established through natural
and assisted regeneration. This approach has
enabled the restoration of ecosystems in histori-
cally degraded areas. The plantations are cur-
rently nine years old. The project offers a unique
opportunity to gain valuable knowledge about
management practices and appropriate selec-
tion of native tree species for commercial for-
estry plantations, particularly with respect to
seed selection, planting, management and main-
tenance.
The reforestation and restoration activities of
the project increased economic activity through
the sustainable nature of practices being imple-
mented, and the creation of employment for local
communities. Even people from local communi-
ties who previously engaged in mining practices,
were encouraged to participate in this project,
which provided economic stability and contrib-
uted to improving people’s living standards.
Today, the area is a positive example of what
can be achieved in a short term by taking advan-
tage of voluntary carbon markets. With around
1292 hectares already planted, the project not
only generated 198 000 VCUs (verified carbon
units, equivalent to 198 000 tonnes of CO2 fixed,
or the annual average emissions of more than
90 000 Colombians), but also delivered climate,
community and biodiversity co-benefits recog-
nized by the Climate, Community and Biodiversity
Standards. At the time of this project, the trading
price of carbon was more than €8/t (Asorpar Ltda
and South Pole Carbon, 2011).
Figure 14.6 provides an overview of the
volume of wood and amount of biomass pro-
duced and CO2 sequestered by individuals of dif-
ferent tree species between 2002 and 2010 in
areas degraded by mining and intensive cattle-
raising.
Acacia mangium showed the strongest growth,
greatest uniformity in plantations and best adap-
tation to soils damaged by mining. It is also legu-
minous, forming a symbiosis with nitrogen-fixing
bacteria and producing abundant leaf litter; this
in turn adds organic matter to the soil. As such,
the species creates a microclimate in which many
Figure 14.5.
The restoration site after levelling of the soil with a bulldozer. The little naturally occurring
vegetation visible in the picture was preserved

GENET IC CONS IDERATI ONS IN ECO SYSTEM RE STORATI ON USING N ATIVE TRE E SPECIE S
219
other plant and animal species can establish and
thrive and further enrich the understorey. For
these reasons, Acacia mangium was combined
with fast-growing native species, some estab-
lished by means of the enrichment method and
others through dispersal of collected seeds.
The non-contaminated lagoons left by gold
mining were used for fish farming with species
Figure 14.6.
Volume of wood and CO2 sequestered per hectare in project areas previously degraded by mining
activities (A) and intensive cattle raising (B)
(A)
(B)

THE STATE OF THE WORLD’S FOREST GENETIC RESOURCES – THEMATIC STUDY
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220
such as red-bellied pacu (Colossoma bidens), red
tilapia (Oreochromis sp.), carp (Cyprinus carpio)
and small mouth (Prochilodus magdalenae reticu-
lata). The lagoons are now surrounded by trees,
the water is cleaner, there is less evaporation,
temperatures are cooler and the aesthetic aspect
of the sites has improved.
The approach described has been replicated at
other locations. Of the 12000 hectares previously
degraded by gold mining, more than 50 percent
have already been recovered.
The success of this approach depends in large
part on the direct involvement of communities liv-
ing in project areas, respecting their culture, and
contributing to the improvement of their living
environment and standard of living. After all, one
of the most valuable results of the project is to
generate alternative ways of life, and at the same
time create new sources of income. Thanks to the
more than 40000 days of paid work generated
by the project, people have been able to improve
their health, quality of life and purchasing pow-
er. In addition, the project provided people with
guidelines about administrative, environmental,
social and financial matters. As the ecosystem re-
covered, natural control systems against the pro-
liferation of malaria vectors were restored, reduc-
ing the impact of the disease.
Conclusions
• The basis and key to success in reforestation
projects lies in the quality of the seed
being used. Therefore, seed trees must be
carefully selected and conserved through
vegetative propagation in a clone bank and
seed orchard. In addition, seeds should be
used from areas belonging to the natural
distribution of each species.
• Forest remnants occurring in the vicinity
of the plantation should be studied to
assess natural regeneration, the dominant
species, species phenology, associated fauna,
the most common pests and diseases, and
other basic information necessary for the
Figure 14.7.
Current status of restoration site after 96 months. The site is dominated by pedro tomín (Cespedesia
macrophylla).

GENET IC CONS IDERATI ONS IN ECO SYSTEM RE STORATI ON USING N ATIVE TRE E SPECIE S
221
development of the restoration project.
• Only healthy seeds of known origin and
from selected seed trees should be planted.
• Undesirable individuals in the forest nursery
should be discarded.
• All management oriented to the
establishment and maintenance of
plantations should be performed in
the most natural way possible: always
remember that we should be working with
nature.
• Combine technology with empirical and
natural knowledge. The environment should
not be modified to any significant extent,
as this will usually damage the project
outcome.
• Align management and technical guidelines.
Operational resources are already available
in the region; farmers are generally innate
planters.
• The initial establishment practices (weeding,
digging, fertilizing, planting) and the origin
of the plant material largely determine the
success of the plantation.
• After the plantation has been established,
proper maintenance activities (weeding
and fertilizing) are indispensable for the
development of trees planted.
• Weeding contributes directly to the growth
of trees and to the control of pests and
diseases.
• Take advantage of the land between the
tree stands, for example, by means of
agroforestry or silvipastoral systems. Such
practices result in resources being reinvested
in the project and make it more profitable.
They also help to control pests and diseases
and contribute to making fuller use of
fertilizers.
• Encourage local people to be involved in
the establishment and management of the
plantation, so as to enhance the chances
of success and significance of the project.
Local communities are the beneficiaries of
the project, receiving the vast economic and
social benefits from reforestation.
• Care for wells, channels, creeks and rivers. No
weeding or cutting of forests surrounding
swamps and lagoons should be allowed.
• Protecting forests by means of fences and
surveillance by rangers guarantees the
longevity of future forests.
Recommendations
• More research should be conducted by
government agencies on forest-tree
breeding and production. If this is not
done, valuable and fast-growing species for
potential use in reforestation pilot schemes
will disappear.
• Continue to select superior or plus trees at
various sites within the natural distribution
of species to obtain a broad genetic base
of planting material with a wide range of
shapes, volumes and physical, chemical and
mechanical properties.
• There is a need to collect as much
information as possible about the selected
plus trees in order to determine their
edaphic requirements, production potential,
best planting distances, susceptibility to
pests and diseases, enrichment methods,
behaviour of the species, phenology and
germination potential of seeds, nursery
propagation techniques and plantation
systems.
• Insurance policies covering abiotic and biotic
risks in the plantations should be promoted
to encourage and secure investment while
attracting capital.
• Advantage should be taken of forest
plantations and their surroundings to
install research programmes funded by the
government to encourage scientific work.
Acknowledgements
The author thanks Asorpar Ltda. for manag-
ing the project and encouraging both individu-
als and legal partners to contribute financially
to its implementation. Thanks also go to South
Pole Carbon for being in charge of preparing the
studies and selling carbon assets, and to forest

THE STATE OF THE WORLD’S FOREST GENETIC RESOURCES – THEMATIC STUDY
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222
engineer Victor David Giraldo Tirado for his in-
valuable assistance in the interpretation of the
results.
References
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Reports ASORPAR Ltda., 1995–2000. Medellin,
Colombia, Asorpar Ltda.
Delgado Higuera, M. 1999. Proposals for agriculture
sustainability. Boletín informativo. Colombia, Orius
Biotecnología.
Moscoso Higuita, L.G. 2005. Reforestación – un pro-
ceso natural. Medellin, Colombia, Eurotex.
Further reading
Aubad Echeverri, J. & Estrada Lozano, J. 1999.
Snakes, recognition and conservation. Physical Biotic
Monitoring Series Porce II N° 4. Medellin, Colombia,
Empresas Públicas de Medellín, E.S.P. Unidad de
Comunicaciones y Relaciones Corporativas.
Aubad Echeverrí, J. & Londoño Florida, C. 1999.
Birds, biology and conservation. Physical Biotic
Monitoring Series Porce II N° 5. Medellin, Colombia,
Empresas Públicas de Medellín, E.S.P. Unidad de
Comunicaciones y Relaciones Corporativas.
Bernal Euse, J. 1994. Tropical pastures and forages –
production and handling. 3rd ed. Colombia, Banco
Ganadero.
Botina, P., Rodrigo. 1998. Non timber products of the
Putumayian and Caucan Amazonia. Lima, Asociación
Bosques y Desarrollo.
Carreño Sandoval, E. & Martínez Barón, A. 1983.
Answers of 10 forestry species to different pre-
germinative treatments and repetition in green-
house. Bogotá, Universidad Distrital Francisco José
de Caldas, Facultad de Ingeniería Forestal.
Castaño Uribe, C. 1989. Guide to the system of
National Parks of Colombia. Bogotá, INDERENA.
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15.1. Species restoration
through dynamic ex
situ conservation: Abies
nebrodensis as a model
Fulvio Ducci27
Consiglio per la Ricerca e la sperimenta-
zione in Agricoltura, Forestry Research
Centre (CRA-SEL), Arezzo, Italy
Abies nebrodensis (Lojac.) Mattei, or Sicilian fir,
is an endemic species of Sicily, Italy, growing on
the Madonie range at 1700–1900 m above sea
level. It is a highly endangered species (Conseil
de l’Europe, 1977; Morandini, 1986; Raimondo
et al., 1992; IUCN, 2007; Thomas, 2011), compris-
ing a single relict population of approximately 30
adult trees spread over an area of 150 hectares
(Morandini et al., 1994; Virgilio, Schicchi and La
27 Coordinator of FAO Silva mediterranea WG 4 “Forest Genetic
Resources in the Mediterranean Region”.
Mela Veca, 2000; Figure 15.1). Specific measures
are needed to conserve and restore genetic varia-
bility of the species in a dynamic way that enables
continued evolution of the species, particularly as
climate changes.
Although an extreme example, Abies nebro-
densis can provide useful insights for the con-
servation and restoration of other species and
their genetic variation. The populations of many
tree species are fragmented in the southern part
of their distribution range, yet these popula-
tions may represent valuable sources of genetic
information under changing environment. The
factors that affect their genetic patterns and vi-
ability must be understood and managed in order
to develop methods for dynamically preserving
such gene pools. Climate change and ecological
changes in the Madonie range are rendering the
locale unsuitable for the long-term maintenance
of the existing gene pool of Sicilian fir, indicat-
ing the need to adopt pragmatic strategies to
preserve the species. Environmental or practical
considerations do not always allow such conser-
vation in situ, in natural populations. This article
presents options for species restoration and dy-
namic gene conservation ex situ, based on recent
genetic studies on Abies nebrodensis.
Status of Abies nebrodensis
Abies nebrodensis was more widely distributed in
the past in the Madonie range in Sicily, as shown
by fossil wood samples of about 9000 years old
(Biondi and Raimondo, 1980). The range of the
species declined rapidly at the end of the seven-
Chapter 15
Species restoration approaches

THE STATE OF THE WORLD’S FOREST GENETIC RESOURCES – THEMATIC STUDY
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226
teenth century, when demand for timber in the
neighbouring villages increased at the same time
as grazing expanded to higher elevations.
Local botanists have known of the existence of
a peculiar population of Abies in Sicily for several
centuries. Indeed, samples of the Sicilian fir can be
found relatively frequently in herbaria dating from
the end of the eighteenth century (Raimondo,
Giancuzzi and Schicchi, 1990). Abies nebroden-
sis was first described as a separate species in the
beginning of the twentieth century (Lojacono
Pojero, 1907; Mattei, 1908). Several authors re-
lated A. nebrodensis to Abies alba Mill. (Silver fir),
and in particular to the nearby Calabrian A. alba
provenances, on the basis of morphology (Arena,
1959; Gramuglio, 1960; Morandini, 1969) and us-
ing needle anatomical traits (Bottacci, Gellini and
Grossoni, 1990). However, A. nebrodensis seems to
derive from the pre-glacial group of several fir spe-
cies which are the ancestors of A. nordmanniana
(Steven) Spach (Caucasian fir), A. cilicica (Antoine
& Kotschy) Carrière (Cilician fir), A. cephalonica
Loudon (Greek fir), A. numidica de Lannoy ex
Carrière (Algerian fir) and A. alba in southern
Calabria. Genetic distance analysis (anatomical, bi-
ochemical and molecular) showed a clear discrimi-
nation of A. nebrodensis from the other species
(Ducci et al., 2004; Camerano et al., 2012). At the
same time, this species has preserved some traits
typical of more eastern fir species (Ducci, Proietti
and Favre, 1999; Ducci et al., 2004) as well as traces
of ancient exchanges with geographically nearer
species (A. alba, and A. numidica). In some cases, A.
nebrodensis showed an allelic pattern completely
different from that of neighbouring populations
of A. alba but closer to that of the oriental spe-
cies. Results of the genetic analyses were similar
for needle morphology and anatomical traits.
Together, they suggest that A. nebrodensis could
be the focal point where A. alba traits, A. cepha-
lonica traits and A. numidica traits converged in
the past. For phylogenetic and genetic studies
Figure 15.1.
Current natural distribution of Abies nebrodensis. The within-population genetic structure increases
progressively from the centre towards the peripheral rings in the diversity core zone (orange
concentric rings, Vallone Prato). The extinction phase was detected in the Contrada Timpa Rossa
area, while the recolonizing phase (blue line) is associated with only two trees near the Vallone della
Madonna degli Angeli.
Source: modified after Ducci et al. (1999); De Rogatis, 2011, unpublished data.

GENET IC CONS IDERATI ONS IN ECO SYSTEM RE STORATI ON USING N ATIVE TRE E SPECIE S
227
on Abies nebrodensis, see Vicario et al. (1995),
Parducci et al. (1996, 2001) and Ducci et al. (2004).
Research aimed at conservation of the relict
A. nebrodensis population started in the 1940s,
and initially focused mainly on the biology of the
species, monitoring the population and restarting
the local forest ecosystem succession. Scientific
techniques used in this period included invento-
ries of existing trees, periodic surveys of sexual
reproduction, anatomical studies on adaptive
traits on needles, seed germination, vigour and
phyto-ecological aspects in the area. Those stud-
ies supplied clear evidence for the high potential
of the population to restore its dynamism, which
has since been confirmed by phylogenetic and
genetic studies within this population. In the in-
ventories carried out in 1992 (Morandini, Ducci
and Menguzzato, 1994) new trees were recorded
in situ, and a very slow but progressive restarting
of the forest ecosystem was documented. Starting
from only two or three flowering trees in 1960s,
the amount of pollen and cones produced by the
population has significantly increased (Arena,
1960; Gramuglio, 1962).
Genetic diversity within the A. nebrodensis
population was estimated using allozyme markers
(11 loci, 32 alleles; Ducci, Proietti and Favre, 1999).
The within-population genetic structure was com-
pared with other Mediterranean fir species and
another reference set of 16 Italian populations of
A. alba. These results showed how certain alleles
have contributed to differentiating the A. nebro-
densis population from the reference populations
of the other Mediterranean fir species. Moreover,
it was discovered that the genetic diversity with-
in the A. nebrodensis population, despite there
being only 30 remaining adult trees, was similar
to the wider and dynamic populations of A. alba
growing in analogous isolation and progressive
drifting situations (i.e. the forests of Gariglione in
Calabria and La Verna in Tuscany). Simultaneously,
a very high excess of homozygotes was detected
in the A. nebrodensis population. The microenvi-
ronmental diversity in situ has allowed the main-
tenance of very high within-population diversity
parameters and relatively good genetic structur-
ing compared with wider ranging and more nu-
merous populations of Mediterranean firs. The
species has a well-defined genetic structure that
is characterized at topographic level within the
150 ha of the present natural range. Genetic
analyses have shown clearly that the closer the
ecological conditions are to the species optimum,
the higher the genetic variation and the greater
the ability of the population to preserve its gene
pool (Vendramin, 1997; Ducci, Proietti and Favre,
1999).
On the basis of the genetic analyses, three
distinct zones were identified within the A. ne-
brodensis population: the diversity core of the
species, one site in recolonizing phase and one
site in an extinction phase (Figure 15.1). The com-
parison of results of enzyme analyses carried out
in 1999 (Ducci, Proietti and Favre, 1999) and am-
plified fragment length polymorphisms (AFLPs)
(De Rogatis, unpublished data) confirmed the
geographic distribution of the variation of the
population in situ. The genetic situation of the
Sicilian fir within each zone reflects the microen-
vironmental conditions, which can be classified as
favourable, highly favourable and unfavourable
for survival and natural regeneration in the three
zones.
The relatively good situation of the population
from the biological point of view is in contrast
with the generally unfavourable site character-
istics, which limit the survival possibilities of the
present A. nebrodensis population. The present
refuge is restricted to a mountain top, which pre-
cludes natural spread. The ecosystem has been
severely affected by human activities, particularly
deforestation and grazing by goats. Only the re-
colonization zone of Vallone della Madonna degli
Angeli is really suitable for natural regeneration.
Furthermore, climate change is expected to mod-
ify the local environmental conditions as climate
belts move upward. In the case of Vallone della
Madonna degli Angeli (about 1800 m), the trees
cannot migrate to a more suitable environment.
Indeed, rocky soils do not allow natural regenera-
tion and migration outside the Vallone and the
trees at the mountain top have no possibility of

THE STATE OF THE WORLD’S FOREST GENETIC RESOURCES – THEMATIC STUDY
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228
migrating to a higher elevation. Other adverse
ecological factors include a progressively increas-
ing risk of fires.
Genetic effects add to the constraints imposed
by environmental factors on regeneration. Self-
pollination reduces genetic variability across the
population, and the long distances between the
trees encourage selfing, leading to progressive ge-
netic erosion. The few seedlings currently grow-
ing in the core zone all originated from a single
mother tree (no. 21, Figure 15.1). Thus, the only
area where the population might regenerate is
characterized by a severe “founder effect,” the
loss of genetic variation that occurs when a new
population is established by a very small number
of individuals from a larger population. The same
applies to the new tree generations in the zone
undergoing recolonization because of the lack of
mother trees that produce viable seed. Even if the
present adult trees can survive, microenvironmen-
tal conditions that adversely affect natural regen-
eration and survival of seedlings severely threaten
the population, except for the small recolonizing
zone of Vallone della Madonna degli Angeli. Here,
the genetic variation will be low and influenced
by the above-mentioned founder effect.
Action needed for species restoration
Previous work has shown the urgent need to re-
duce the genetic erosion of the important gene
pool of A. nebrodensis, which contains traits
common to the Mediterranean firs. The first step
should be to stop the loss of rare alleles and phe-
notypic traits of possible adaptive significance.
Another goal should be to reduce the strong
excess of homozygotes, the presence of which
implies constraints in maintaining the high poly-
morphism observed in the population. This will
require random mating among all genotypes.
The present genetic situation of A. nebrodensis
in situ is potentially suitable for re-establishing
the dynamics of the species, but the environmen-
tal conditions are not conducive to long-term
spontaneous survival. When the present genera-
tion concludes its biological cycle, the species will
be at real risk of extinction.
Ducci, Proietti and Favre (1999) studied the in-
trapopulation genetic structure of Abies nebro-
densis to explore the possibilities of conserving
the species in situ. The aim of their work was to
develop a strategy for the re-establishment of the
biological dynamics of the existing gene pool, set
bases for future spreading of propagation materi-
als of A. nebrodensis in the Madonie range and,
in view of the effects of climate change, to es-
tablish new and dynamic populations in suitable
locations. The authors concluded that practically
the only way to achieve these goals, and the one
involving least risk, would be ex situ conservation
using seed orchards devoted to increasing mixed
mating in the next generations.
There have been some previous attempts to
propagate A. nebrodensis through seed collec-
tion and grafting. However, even if it is possi-
ble to obtain large numbers of seedlings under
nursery conditions, moving seed from the natural
population reduces the number of seeds available
to support species survival in situ and increases
genetic erosion. Moreover, the small in situ areas
still suitable for reforestation programmes would
then be planted with materials mainly from only
three or four genotypes that have regularly re-
produced in situ. This approach would remove
any possibility of reducing genetic erosion by
introducing reproductive materials with higher
levels of diversity. Collection of cones within the
Strict Reserve of the Regional Park of Madonie
Mountains should be completely forbidden or at
least strictly controlled and monitored.
A few grafts of A. nebrodensis have been pro-
duced and distributed among several European
arboreta in the past. Regrettably, the source of
the grafts has not been documented and it is pos-
sible that they originate from a single specimen
(growing within the garden of the Villa of Casale
baron in Polizzi Generosa). To our knowledge, the
following grafts are still growing:
• Five grafts (unknown original genotype)28 in
the arboreta of Barres and Amance, France,
prepared by Dode in 1930 (Morandini,
28 Technically defined as an ortet.

GENET IC CONS IDERATI ONS IN ECO SYSTEM RE STORATI ON USING N ATIVE TRE E SPECIE S
229
1969). The grafts produced viable seedlings
through open pollination (Fady, personal
communication).
• Three grafts (only one mother tree [ortet],
identified by genetic analyses) within the
garden of Villa Lanza near Gibilmanna
(Palermo, Sicily) (Morandini, 1969).
• One graft, probably from the same ortet
as above, at the arboretum of Borde
Hill (Sussex, United Kingdom) of which,
however, no recent news is available.
• In addition, seedlings of A. nebrodensis are
known from the following locations:
• About 40 seedlings, 10 to 12 years old
(provenance Vallone Madonna degli
Angeli, unidentified mother trees), were
planted near Papiano (province of Arezzo,
43°48’55.008’’N, 11°41’46.68’’E, 752 m above
sea level, southwest aspect).
• Eighty seedlings, of the same age as above,
near the State Forest Service nursery in
Pieve Santo Stefano (province of Arezzo,
43°39’16.5954’’N, 12°3’22.68’’E, 700 m above
sea level, northwest aspect). They are now
monitored annually for pollen and cone
production.
• Three 25-year-old trees in the arboretum
of Vallombrosa (province of Florence,
43°43’52.392’’N, 11°33’18.3594’’E, 1000 m
above sea level, westerly aspect).
Other grafts were distributed to private arboreta
in France (M. Albert Dumas) and in Germany (Herr
Klaus Albert Höller).
All these collections can contribute to ex situ
conservation of A. nebrodensis, but they are
characterized by the absence of information
about mother trees or ortets. Genetic analyses
are being carried out to determine the parental
origin of these materials, but most of them
have to be excluded as seed sources because of
possible hybridization with other species or self-
pollination. The importance of establishing seed
orchards devoted to re-establishing a dynamic
genetic structure of the species is clear. The
materials should be characterized by greater
heterozygosity and diversity than the previously
used seedlings. The ex situ conservation approach
should be integrated with in situ conservation
approaches.
Species restoration through dynamic ex situ
conservation
A specific programme for the ex situ conservation
and restoration of Abies nebrodensis was started
in 1992 in response to the endangered status of
the species and the anticipated effects of climate
change. The aim of the programme was also to
use A. nebrodensis as a model for studying ex situ
conservation and species restoration methods for
small or marginal population gene pools.
To produce reproductive material able to re-
store the genetic dynamism in A. nebrodensis, two
orchards (clone archives) were established in 1994
in Tuscany near Arezzo (Pomaio 43°28’39.36’’N,
11°57’0.72’’E, 690 m above sea level and Caprile,
43°43’34.212’’N, 12°7’6.96’’E, 910 m above sea lev-
el), 1200 km north from the original population
(Figure 15.2).
The experiment represents a model strategy to
conserve ex situ the entire gene pool of a species
through massive clonal replication. The orchards
were established with grafts from the 27 adult
trees of the original population. Materials were
grafted onto four-year-old rootstocks of Abies
alba (provenance Serra S. Bruno, Calabria) in pots.
Each mother tree (ortet) is replicated at least
five times. In total, about 200 grafts grow in the
orchards, and each year some of them are reju-
venated in order to have new clonal copies. The
orchards were established according to a com-
plete single-tree random design to improve pol-
len exchange among genotypes and increase het-
erozygosity, which in the natural population are
constrained by the very low density of adult trees.
The clones in the lower-altitude orchard start-
ed to reproduce in 1997, with 80 percent of the
clones producing male flowers. Two clones (1and
22) produced the first cones in spring 2000. In
total, 110 seedlings survived in the nursery. In
spring 2005, three clones (17, 22 and 29) produced
new cones with about 300 g of seeds, of which
half were sown and the rest dried and conserved

THE STATE OF THE WORLD’S FOREST GENETIC RESOURCES – THEMATIC STUDY
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230
at −5°C. About 150 seedlings were obtained. In
spring 2007, two clones (1 and 6) produced cones,
and in 2010 about 110 new seedlings were ob-
tained from six ortets (1, 6, 13, 17, 18 and 24).
Numbers of seedlings produced seem to be low,
as a result of the young age of the orchards and
the low rates of germination. A cut test showed
an average of about 20 percent empty seeds, and
germination rates were extremely low, 1–39per-
cent, very similar to the original population.
Nevertheless, each year more maternal genotypes
are producing seeds and open-pollinated off-
spring (Figure 15.3).
The conservation programme was continued in
2007 with the establishment of two permanent
experimental plots for dynamic ex situ conserva-
tion of A. nebrodensis. In this model, the gene
pool of the original population is introduced and
tested within a new ecological context. The effects
of different patterns of driving forces (e.g.more
light, more drought or more continental climate)
are assumed to increase the possibility of conserv-
ing different alleles in different combinations.
The ex situ conservation plots are located in
the northern Apennines (1000 m above sea level,
northern aspect) and were established in col-
laboration with the State Forest Service, Forest
Biodiversity Bureau of Pieve S. Stefano (Arezzo).
The plots differ in their ecological contexts. One
of the plots is under the canopy cover of sweet
chestnut (Castanea sativa Mill.) and other noble
hardwoods at very low tree density. The other
plot was planted in an abandoned field area sur-
rounded by a forest consisting of European beech
(Fagus sylvatica L.), Turkey oak (Quercus cerris L.)
and hop hornbeam (Ostrya carpinifolia Scop.)
(Figure 15.4).
The new ex situ populations are “open,” in
that they were established using material from
more than one seed collection (i.e. from succes-
sive years and from different genotypes). The
idea is to plant the offspring produced in the two
seed orchards year after year. The establishment
of grafted seed orchards allows the replication
of the source genotypes several times. This way
all the adult trees of the population have been
Figure 15.2.
The natural range of A. alba (light green) and of A. nebrodensis (orange cross) in Sicily, and the
location of the clonal orchards and ex situ dynamic conservation test in Tuscany (orange circles)

GENET IC CONS IDERATI ONS IN ECO SYSTEM RE STORATI ON USING N ATIVE TRE E SPECIE S
231
rescued from possible random genetic losses, and
the homozygote excess will also be reduced ef-
ficiently.
Both plots are planted according to a random
design in order to create maximum opportunities
for outcrossing within the population. Together
with the randomized design, having such differ-
ent microenvironmental contexts contributes to
maintaining allelic richness and the variability of
genes or gene patterns responsible for adapta-
tion. Moreover, offspring and consequently the
gene pool of the population are being intro-
duced in new ecological contexts in order to start
new dynamics that may help increase heterozy-
gosity. Nevertheless, the evolutionary forces in
these new contexts may be different from those
acting in situ and consequently the populations
may adapt and evolve differently. Monitoring the
genetic dynamics in ex situ conservation and the
genetic variation in offspring is one of the aims
of the experiment, in order to have a reference
model of the possible genetic implications of such
conservation approaches.
Next steps in the ex situ conservation pro-
gramme will be artificial flower induction in the
orchards, and the use of genetic markers to mon-
itor the genetic structure of the new generations
produced. This will allow the orchards to supply
in situ nurseries with reproductive materials that
have been monitored and checked for their ge-
netic structure.
Experiences gathered during these research
activities have allowed local Sicilian bodies to
develop strategies for dynamic in situ conserva-
tion of the relict A. nebrodensis population. In
the 2000s the Ente Parco delle Madonie, the re-
gional entity responsible of the management of
the Regional Park of the Madonie Range, where
the Sicilian fir is protected in situ, established a
strategy for developing in situ conservation initia-
tives. A Life Project on “Conservation in situ and
ex situ of Abies nebrodensis (Lojac.) Mattei” was
approved and funded in 2000.29 A second project,
“Conserving Abies nebrodensis and restoring the
bogs of Geraci Siculo,” was approved and funded
in 2004.30 These projects focus on preserving the
individuals still surviving in situ. As in our project,
they grafted and established a clonal orchard to
preserve the original gene pool in situ, with the
aim of increasing the species’ distribution and
starting future reforestation programmes. So,
two parallel strategies were developed, one in
situ and another ex situ.
29 http://www.parcodellemadonie.it/doc/Relazione_progetto_di_
recupero_Abies_nebrodensis.pdf
30 http://www.parcodellemadonie.it/conservazione-di-abies-
nebrodensis-e-ripristino-delle-torbiere-di-geraci-siculo.html
Figure 15.3.
A grafted Abies nebrodensis tree producing
cones in one of the Tuscan orchards

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232
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Relationships between Abies nebrodensis (Lojac.)
Mattei and other Mediterranean firs. Ann. Ist. Sper.
Selv., Arezzo, 31: 73–93.
Gramuglio, G. 1960. Appunti sulla distribuzione geogra-
fica dell’Abies nebrodensis (Lojac.) Mattei. Acc. naz.
Lincei, 29: 106–114.
Gramuglio, G. 1962. Sexual awakening of Abies nebro-
densis. Giorn. Bot. Ital., 69(1/3): 207–210.
Figure 15.4.
Ex situ conservation plots of Abies nebrodensis (a) under the cover of chestnut and maple trees and
(b) in an open field near a Fagus sylvatica–Quercus cerris–Ostrya carpinifolia forest
(a) (b)

GENET IC CONS IDERATI ONS IN ECO SYSTEM RE STORATI ON USING N ATIVE TRE E SPECIE S
233
IUCN. 2007. A preliminary world list of threatened coni-
fer taxa. Biodivers. Conserv., 2: 304–326.
Lojacono Pojero, M. 1907. Flora sicula. Palermo, Italy.
Mattei, G.E. 1908. L’abete dei nebrodi. Boll. R. Orto.
Bot. Giard. Col., Palermo, 7: 56–59.
Morandini, R. 1969. Abies nebrodensis (Lojac) Mattei,
Inventario 1968. Pubblicazione n° 18. Arezzo, Italy,
Istituto Sperimentale per la Selvicoltura.
Morandini, R. 1986. Abies nebrodensis (Lojac.) Mattei.
In FAO. Databook of endangered tree and shrub
species and provenances. FAO Forestry Paper 77:
11–20.
Morandini, R., Ducci, F. & Menguzzato, G. 1994.
Abies nebrodensis (Lojac) Mattei – Inventario 1992.
Ann. Ist. Sper. Selv., Arezzo, 22: 5–51.
Parducci, L., Szmidt, A.E., Villani, F., Wang, X.R.
& Cherubini, M. 1996. Genetic variation of
Abies alba in Italy. Hereditas, 125: 11–18. doi:
10.1111/j.1601-5223.1996.00011.x
Parducci, L., Szmidt, A.E., Ribeiro, M.M.A. & Drouzas,
D. 2001. Taxonomic position and origin of the
endemic Sicilian Fir Abies nebrodensis (Lojac.) Mattei
based on allozme analysis. Forest Genet., 8(2):
119–127.
Raimondo, F.M., Giancuzzi, L. & Schicchi, R. 1990.
Carta della vegetazione del massiccio carbonatico
delle Madonie (Sicilia centro - settentrionale). Quad.
Bot. Ambient., Palermo, 3: 23–40.
Vendramin, G.G. 1997. Abies nebrodensis (Lojac.)
Mattei, a relevant example of a relic and highly
endangered species. Bocconea, 7: 383–388.
Vicario, F., Vendramin, G.G., Rossi, P., Liò, P. &
Giannini, R. 1995. Allozyme, chloroplast DNA and
RAPD markers for determining genetic relation-
ships between Abies alba and the relic popula-
tion of A. nebrodensis. Theor. Appl. Genet., 90:
1012–1018.
Virgilio, F., Schicchi, R., La Mela Veca, D. 2000.
Aggiornamento dell’inventario della popolazione
relitta di Abies nebrodensis (Lojac.). Naturalista
Siciliano, 24(1–2): 13–54.
Thomas, P. 2011. Abies nebrodensis. In IUCN. IUCN red
list of threatened species. Version 2011.2 (avail-
able at http://www.iucnredlist.org/apps/redlist/
details/30478/0).
15.2. Restoration and
afforestation with Populus
nigra in Hungary
Sándor Bordács and István Bach
Central Agricultural Office, Hungary
In the 1990s, European black poplar (Populus
nigra L.) was declared an endangered species in
Europe because of widespread planting of hybrid
poplar and genetic hybridization and introgres-
sion by pollen flow from hybrid poplar clones
(mainly P. × euramericana).
Since 1997, reforestation or afforestation on
protected areas in Hungary has been carried out
using only autochthonous species, as mandated
by the National Law of Nature Protection. This
regulation has increased demand for reproduc-
tive material of European black poplar for affor-
estation on protected sites, mostly in the former
willow–poplar riparian forests (Salicetum albae-
fragilis) near riverbanks. In central Europe the riv-
er bank forests are composed of black and white
poplar (P. nigra and P. alba L.), silver willow (Salix
albaL.), and on less wet sites (higher elevation)
of narrow- leaved ash (Fraxinus angustifolia Vahl)
and pedunculate oak (Quercus roburL.).

THE STATE OF THE WORLD’S FOREST GENETIC RESOURCES – THEMATIC STUDY
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A national programme was started in Hungary
in 1997 to restore riparian forests. As black poplar
was a fundamental species for riparian forests and
there was no autochthonous reproductive mate-
rial available, genetic monitoring had to be carried
out. First, registered occurrences in the national
forest inventory were surveyed. The first survey
listed 4390 ha of forest in Hungary, mostly mixed
forest stands of P. nigra. Thousands of hectares
were surveyed to develop a distribution map of
existing populations, forest stands, small patches
and even solitary trees (Borovics et al., 1999). Since
1998, about 4000 mature trees have been select-
ed across the country for the gene conservation
and restoration programme. All trees have been
characterized morphological descriptors pub-
lished by the European Forest Genetic Resources
Programme (EUFORGEN) to exclude artificial or
natural hybrids (P. × euramericana). Propagation
material, usually stem cuttings, was collected from
the trees to establish clonal gene banks at six lo-
cations in four geographical regions of Hungary.
Additionally, DNA of all sampled trees was tested
according to the method used by Heinze (1997) to
exclude introgressed and hybrid genotypes. Only
clones with proven taxonomic status were regis-
tered in the national database. All the gene bank
clones were tested for cultivation value (e.g. viabil-
ity, root formation, growing capacity). The most
appropriate clones were certified as basic material
in the National List of Basic Materials (NLBM) by
the national authority, and clonal mixtures (repro-
ductive material of clonal composition) have been
produced and marketed. The clonal mixtures have
usually been composed of 30–60 clones and are
recommended for use in the same region in which
the provenances originated. Since 1999, when the
first stock plantations were established, stem cut-
tings have been increasingly produced. Usually, the
stem cuttings are used in nurseries where rooted
cuttings (height 150–300 cm) are produced.
Some of the selected populations and stands
(according to EUFORGEN criteria) have been offi-
cially registered as seed stands in the NLBM. As a
minimum requirement, the stands must comprise
both female and male trees and be at least 1 km
from any hybrid poplar plantations or black pop-
lar stands with unknown origin. Since poplar seeds
are able to germinate for only one or two days,
the female (seed) trees must be felled to collect vi-
able seed. Seedlings are produced in forest nurs-
eries. The seedlings must also be tested for purity
of their taxonomic status. In late summer, a local
inspector collects leaves from the seedlings for
DNA testing. Nursery production is under official
control to fulfil the requirements of the certifica-
tion system for forest reproductive material (FRM).
Only DNA-tested seedling lots can be certified and
marketed.
The certified cuttings and seedlings are used for
forest restoration on protected sites, mainly for
riverbank forests, to recreate natural forest veg-
etation. The reconstruction of forest vegetation
has been carried out using reproductive material
of two to four autochthonous species planted at
Figure 15.5.
An old black poplar tree in the Danube valley
in Hungary

GENET IC CONS IDERATI ONS IN ECO SYSTEM RE STORATI ON USING N ATIVE TRE E SPECIE S
235
5000–8000 trees/ha. Based on the experiences of
local foresters, rooted cuttings can better survive
short-term high floods and seedlings can better
tolerate long-term lesser flooding with muddy
water. Between 2000 and 2010 a total of 120 000
to 1 060 000 black poplar seedlings and 55 000
to 350 000 rooted cuttings were certified as FRM
each year and marketed in Hungary (Figure15.6).
Each year, 30-120 ha of river bank forests are
being restored. Black poplar is also being plant-
ed in roadside plantations. Reforested areas are
under the official control of local inspectors of
the state forest service. The inspectors check cer-
tificates (supplier’s documents) of FRM used for
planting and survey the plantations regularly.
The basic data and ecological information on the
reforestation area must be documented in the
forest management plan, including species com-
position, total area, origin of FRM, and soil type.
Management interventions such as nursing and
tending may be required. The restoration forests
are supposed to be managed less intensively than
hybrid poplar plantations, with the aim of estab-
lishing forest that is as close to natural as possible.
Information on the black poplar restoration pro-
gramme and its results have reported and pub-
lished for the public, especially for local commu-
nities and nature conservation organizations. The
first positive results from the black poplar restora-
tion programme have stimulated work on gene
conservation programmes for other endangered
tree species, such as Sorbus and Pyrus species.
References
Heinze, B. 1997. A PCR marker for a Populus deltoides
allele and its use in studying introgression with
native European Populus nigra. Belg. J. Bot., 129:
123–130.
Borovics, A., Gergácz, J., Bordács, S., Bach, I.,
Bagaméry, G. & Gabnai, E. 1999. A fekete nyár
génmeg˝orzésben elért eredmények [Recent results
on genetic conservation of black poplar]. Erdészeti
Kutatások 89: 135–148.
Figure 15.6.
Certified reproductive material of black poplar (Populus nigra L.) produced in Hungary from 2000
to 2010

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15.3. Restoration of threatened
Pinus radiata on Mexico’s
Guadalupe Island
J. Jesús Vargas-Hernández,1 Deborah L.
Rogers2 and Valerie Hipkins3
1 Programa Forestal, Colegio de
Postgraduados, Montecillo, México
2 Center for Natural Lands
Management, Temecula, California,
United States
3 National Forest Genetics Laboratory,
US Forest Service, Placerville, California,
United States
Monterey pine (Pinus radiata D. Don) is a forest
tree species of great economic importance world-
wide, although its native range is restricted to
the coastal zone of central California and north-
ern Baja California (Figure 15.7). The species is
grown in exotic commercial plantations on over
4 million hectares, but in its countries of origin
it faces heightened conservation issues; the spe-
cies has lost over 50 percent of its natural habi-
tat and is threatened by various human-related
disturbances. Monterey pine is on the IUCN (In-
ternational Union of Conservation of Nature) Red
List of threatened species, and the FAO Panel of
Experts on Forest Gene Resources has identified
it as a species with high global, regional and/or
national priority for genetic conservation.
The island of Guadalupe in the Pacific Ocean,
250 km off the coast of Baja California, Mexico,
hosts one of the five remnant natural popula-
tions of Monterey pine. Environmental conditions
on the island are harsh, with annual rainfall av-
eraging less than 200 mm. Although dense fogs
are common in winter, especially at higher el-
evations, they are less frequent in summer. The
volcanic-origin island has thin, rocky soils with
little organic substrate. The Monterey pine pop-
ulation on this island has evolved isolated from
the other island and mainland populations, so it
has become genetically differentiated from them,
showing distinctive morphological and adaptive
traits as well as genetic diversity measured with
molecular markers. Some authors recognize this
population with the varietal name of P. radiata
var. binata. The original pine population once oc-
cupied an extensive area on the northern end of
the island. However, even though the island has
not been permanently inhabited by humans, the
pine population shrank dramatically in the last
two centuries because of goats that were intro-
duced in the mid-nineteenth century, preventing
successful regeneration of the pines. The current
population is down to about 220 adult, over-
mature trees (2001 census), growing isolated or
in small patches (Figure 15.8), in an environmen-
tal context hostile to recruitment of seedlings.
The drastic reduction in population size led to
the opinion that this population was headed to-
wards extinction. In 1981, the Guadalupe Island
population of Monterey pine was declared “en-
dangered” by the FAO Panel of Experts on Forest
Gene Resources largely because of the grazing
pressure from introduced goats.
In 2001, a multinational team completed an
expedition to Guadalupe Island to make seed
collections of Monterey pine for conservation,
restoration and research purposes. In addition to
collecting seed from individual trees, the team
described the status of the pines, evaluated risks
and threats, and gathered information on pine
ecology to assist in restoration efforts. For exam-
ple, before the expedition it had been speculated
that microsite conditions may have deteriorated
to a state that would no longer support seed ger-
mination or seedling growth. However, soil and
moisture conditions, at least within the canopy
and fog-drip zone of living trees, appeared to

GENET IC CONS IDERATI ONS IN ECO SYSTEM RE STORATI ON USING N ATIVE TRE E SPECIE S
237
Figure 15.7.
Location of natural Pinus radiata populations along the west coast of North America. Mainland
populations occur at Año Nuevo, Monterey and Cambria. Island populations occur on the Mexican
islands of Guadalupe and Cedros
Figure 15.8.
Approximate location of remnant trees of Pinus radiata at the northern tip of Guadalupe Island (see
also Figures 15.9 and 15.10)

THE STATE OF THE WORLD’S FOREST GENETIC RESOURCES – THEMATIC STUDY
PART 3
238
be sufficient to allow at least initial seedling es-
tablishment. The discovery of a few small pine
seedlings supported the hypothesis that natural
recovery may be possible if grazing pressure were
reduced or removed. At the same time, a multi-in-
stitutional project, led by the Grupo de Ecología y
Conservación de Islas A.C. (GECI, a binational non-
governmental organization), in collaboration
with several institutions from the federal govern-
ment, was initiated to eradicate the resident goat
population. The eradication project started in
2000, when Mexican ranchers from Sonora, assist-
ed by GECI, local fishermen and the Marines sta-
tioned on the island, began trapping and remov-
ing goats for direct sale and use as breeding stock
on the Mexican mainland. To test the response
to release from grazing pressure, GECI installed
exclusionary fencing in three areas in the pine
population in summer 2001. Several years later
the potential for natural recruitment of seed-
lings in the grazing-excluded areas was evident,
as was the necessity of complete removal of the
goats (Figure 15.9). The goat eradication project
accomplished its goal on 2007, with Guadalupe
Island being officially declared free of goats.
Environmental conditions conducive to natu-
ral recruitment are one important component of
the restoration process for this pine population.
However, the question remains as to whether
levels of genetic diversity in the population are
sufficient. Because of the rapid and presumably
massive loss of pines, causing fragmentation and
drastic reduction in population size, several ge-
netic impacts may have already occurred, includ-
ing reduction of genetic diversity and increased
inbreeding. Thus, to ensure full recovery of the
population it was important to evaluate the need
for genetic intervention during the restoration
process. For instance, if genetic diversity had been
drastically reduced in the population, it might be
important to reintroduce genetic material from
ex situ collections. Similarly, if inbreeding were
Figure 15.9.
One of the southernmost isolated, over-mature, Pinus radiata trees remaining on Guadalupe Island
with the natural recruitment of seedlings moving out of the protection zone under the tree canopy,
May 2006

GENET IC CONS IDERATI ONS IN ECO SYSTEM RE STORATI ON USING N ATIVE TRE E SPECIE S
239
an issue, actions might be necessary to promote
cross-pollination and seed dispersal between
patches to reduce relatedness among parental
trees in the next generation. Based on a spatially
representative sample of germplasm from about
35 percent of the current population, genetic
diversity, its spatial structure and inbreeding
level were analysed using microsatellite mark-
ers. The sampling structure allowed comparison
of the genetic diversity and inbreeding level in
the progeny (seed) to that in the maternal gen-
eration (remnant trees). Results showed that, de-
spite the drastic reduction in population size, the
level of genetic diversity – both in the parental
trees and their open-pollinated progeny – has
not been greatly reduced. The data also indi-
cated a minimum of 45 percent cross-pollination
in the population. Thus, the genetic information
obtained so far does not support the need for
genetic intervention to restore this population
other than to move seed among resident trees
to increase dispersion distance and accelerate
connectivity between patches (Figure 15.10).
Although the population is still far from restored,
the outlook is much more promising now than it
was ten years ago.
Figure 15.10.
Panoramic view of the northern patches of remnant Pinus radiata trees on Guadalupe Island with the
natural recruitment of seedlings appearing across the landscape and helping to connect the patches

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15.4. A genetic assessment
of ecological restoration
success in Banksia
attenuata
Alison Ritchie
School of Plant Biology, University of
Western Australia, Australia
Despite the importance of genetic management
within ecological restoration for the long-term
persistence, functionality and self-sustainability
of populations, genetic assessments of the success
of ecological restoration remain rare. If a found-
ing population is sourced from a limited genetic
pool, genetic bottlenecking and increased in-
breeding could potentially occur, reducing the
resilience and adaptability of the future popula-
tion to a changing climate. If the population is
established with seeds of non-local provenance,
issues such as outbreeding depression could arise
and lack of genetic integration with surrounding
populations could transpire, influencing the over-
all success of restoration.
Microsatellite markers were used to assess the
genetic success of ecological restoration, asking
two main questions: (i) Is sufficient genetic diver-
sity maintained within adult and offspring indi-
viduals in restored populations? (ii) Is there ge-
netic connectivity between restored and adjacent
natural, undisturbed populations? (Ritchie and
Krauss, 2011). Our focal species was the wood-
land tree Banksia attenuata R.Br. (Proteaceae),
which is wide ly distributed across the south-west
of Western Australia and is a keystone species
used in ecological restoration within this region.
The species displays some of the highest outcross-
ing rates seen among plants (Scott, 1980) and
has mixed generalist pollinator mutualisms with
nectar-feeding birds (predominantly honeyeaters
in the family Meliphagidae), native bees, wasps
and introduced honeybees. Both the restored
and natural undisturbed populations of B. at-
tenuata were located at Gnangara (31°47’09’’S,
115°56’32’’E), 40km north-northeast of Perth in
Western Australia.
The restoration site was established 14 years
ago within a 100-ha leasehold of Rocla Quarry
Products (Rocla), one of five sand extraction quar-
ries surrounding the Perth Metropolitan area. The
restoration efforts at Rocla satisfied the environ-
mental completion criteria of the time, aimed
at restoring species richness, plant density and
cover to the conditions prior to sand extraction.
Restoration of the site focused on two main re-
search areas: (i) seedling recruitment and plant
survival and (ii) plant growth and developmental
responses to a reconstructed soil environment.
Seed pretreatments to enhance germination
(e.g. smoke), greenstock-enabling treatments
(e.g. tree guards, antitranspirants) and various
soil treatments (e.g. mulching, irrigation, ripping
and application of soil stabilizers) were used to
increase plant survival. As a result, by the third
year plant density was 59 plants/5 m2 and the
number of species was 14/5 m2, compared with
78 plants/5m2 and 8 species/5m2, respectively, on
undisturbed adjacent woodland plots, surpass-
ing previous restoration figures (Stevens et al., in
press). Details of the restoration techniques em-
ployed can be found in Rokich and Dixon (2007).
The restored population of B. attenuata com-
prised approximately 200 mature adult trees. The
natural undisturbed population was located ad-
jacent to the Rocla mine and comprised approxi-
mately 350 mature trees within naturally occur-
ring Banksia woodland that is up to 300 years old.
Microsatellite analysis revealed that similarly
high levels of genetic diversity (heterozygosity
and allelic diversity) were maintained within
both restored and natural populations, and were
high for both adult trees and their offspring

GENET IC CONS IDERATI ONS IN ECO SYSTEM RE STORATI ON USING N ATIVE TRE E SPECIE S
241
(Figure15.11). Genetic diversity was also similar
to that of a large, naturally occurring reference
population (Kings Park). There was very weak
population divergence between the restored and
natural populations, signifying sourcing of seed
from local provenances within the restoration
project. Genetic structure was undetectable with-
in the restored population; in contrast the natural
population showed significant structuring up to
30 m. This reflects the pattern of natural seedling
recruitment, as opposed to man-made mixing and
broadcast seeding used in the establishment of
the restored population.
We observed complete outcrossing, very low
biparental inbreeding and low correlated pater-
nity in both restored and natural populations, in-
dicating similar patterns of contemporary mating
at both sites. Paternity analysis revealed extensive
pollen dispersal within and among the two popu-
lations (Figure 15.12). Pollination events were
recorded at distances of 2 m to 324 m between
paternal plants, and more than 50 percent of
paternity was assigned to sires beyond the local
populations (Figure 15.12).
In conclusion, we discovered that the restored
population was successfully integrated with the
natural population, likely because of the initial
sourcing of genetically diverse seed from local
genetic provenances. We also observed the deliv-
ery of robust pollinator services to the restored
population. This study has revealed that, with the
application of best practices for seed collection
and restoration, the desired genetic goals for suc-
cessful genetic management of a keystone species
within a post-mining restoration site have been
achieved.
Figure 15.11.
Genetic diversity measures in populations of Banksia attenuata. Line indicates heterozygosity

THE STATE OF THE WORLD’S FOREST GENETIC RESOURCES – THEMATIC STUDY
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242
References
Ritchie, A.L. & Krauss, S.L. 2011. A genetic assessment
of ecological restoration success in Banksia attenu-
ata. Restor. Ecol., article first published online, 28
June 2011.
Rokich, D.P. & Dixon, K.W. 2007. Recent advances in
restoration ecology, with a focus on the Banksia
woodland and the smoke germination tool. Aust. J.
Bot., 55: 375–389.
Figure 15.12.
Network of inferred pollen dispersal events using a paternity maximum likelihood exclusion analysis.
Dark grey area is the natural, undisturbed population and lighter grey areas are the restored
populations of Banksia attenuata. Each circle indicates the position of a sampled tree, with open
circles indicating maternal trees that were also sampled for seed.
Scott, J.K. 1980. Estimation of the outcrossing rate for
Banksia attenuata R.Br. and Banksia menziesii R.Br.
(Proteaceae). Aust. J. Bot., 28: 53–59.
Stevens, J.C., Dixon, K.W., Newton, V. & Barrett, R.L.
2013. Restoration of Banksia woodlands. Crawley,
Western Australia, University of Western Australia
Publishing.

Part 4
ANALYSIS

GENET IC CONS IDERATI ONS IN ECO SYSTEM RE STORATI ON USING N ATIVE TRE E SPECIE S
245
This chapter discusses genetic aspects of current
practice for ecosystem restoration using native
tree species. Practical implications for the viabil-
ity of the restored tree populations are analysed,
and options are presented for improving resto-
ration success by applying genetic principles. To
this end, the restoration methods and approaches
presented in Part 3 were analysed from a genetic
perspective, based on the theoretical and prac-
tical issues of ecosystem restoration introduced in
Part 2. Authors of the methods described in Part 3
were surveyed to collect additional information
about the genetic considerations of the restora-
tion methods presented. The survey included
questions about species composition, source of
propagation materials, and practical details of
the design and implementation of each method.31
In total, 23 survey responses were received.
Most of the respondents carry out applied re-
search, developing and testing a range of resto-
ration approaches and methods that use native
species. A rigorous quantitative analysis of the
31 The survey is available from https://www.bioversityinternational.
org/fileadmin/user_upload/SoW_FGR_RestorationSurvey.pdf.
survey results was not undertaken, as the meth-
ods and experiences included do not represent a
random sample of ecosystem restoration efforts
globally. Nevertheless, the responses can be con-
sidered indicative of general trends in ecosystem
restoration with respect to genetic aspects, and
provide useful information to guide the incor-
poration of genetic considerations in restoration
projects.
This chapter includes a brief summary of sur-
vey results on identified key areas, followed by a
broader literature-based discussion of the issues
and recommendations for research and action.
16.1. Appropriate sources of
forest reproductive material
Survey results: Only half of the restoration meth-
ods incorporated guidelines or recommendations
for the collection of forest reproductive material
(FRM). Such recommendations were clearly more
common for approaches aimed at conserving or
restoring populations of particular tree species
than for approaches that focused on restoring
Chapter 16
Analysis of genetic considerations
in restoration methods
Riina Jalonen,1 Evert Thomas,1 Stephen Cavers,3 Michele Bozzano,1 David Boshier,1,3 Sándor
Bordács,4 Leonardo Gallo,1,5 Paul Smith6 and Judy Loo1
1 Bioversity International, Italy
2 The Centre for Ecology and Hydrology, Natural Environment Research Council, United Kingdom
3 Department of Plant Sciences, University of Oxford, United Kingdom
4 Central Agricultural Office, Department of Forest and Biomass Reproductive Material, Hungary
5 Unidad de Genética Ecológica y Mejoramiento Forestal, INTA Bariloche, Argentina
6 Seed Conservation Department, Royal Botanic Gardens, Kew, United Kingdom

THE STATE OF THE WORLD’S FOREST GENETIC RESOURCES – THEMATIC STUDY
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246
habitats or ecosystems in general. Sourcing FRM
locally or from similar ecological conditions was
considered ideal for nearly all restoration ap-
proaches. For approaches focusing on ecosystem
restoration, distance of the source of FRM ranged
between a few hundred metres to approximately
100 km from the restoration site, but typically
FRM originated from within a few kilometres
of the restoration site. Half of the respondents
indicated that lack of populations of the target
species in the vicinity of the restoration area very
often limited the availability of FRM.
Ecosystem restoration efforts commonly source
FRM locally. An emphasis on the use of local
germplasm is likely linked to the assumption that
FRM will be well adapted to the ecological con-
ditions of the restoration site, although the rea-
soning is not always stated. In fact, the excessive
focus on “local” germplasm may obscure the fact
that geographical proximity to the restoration
site is not necessarily always the best indication
of the quality or suitability of FRM. The problem
is exacerbated where remaining forests near the
restoration area are fragmented, as trees may
suffer from inbreeding, low fitness of progeny
or other negative consequences of small popu-
lation size, and may not constitute good seed
sources (see Chapter 2; Lowe et al., 2005; Eckert
et al., 2010; Vranckx et al., 2012). These condi-
tions can be assumed to be common in most areas
where restoration efforts are underway, yet rec-
ommendations for the collection of FRM in such
situations are generally lacking. Environmental
changes may also already affect genetic quality
of tree populations as sources of FRM, although
the impacts are not well understood. Quality of
existing forest patches as sources of FRM must be
carefully evaluated in the light of past or ongo-
ing silvicultural management practices and other
forms of resource use or disturbance (Lowe et al.,
2005; Schaberg et al., 2008). For example, some
logging methods may modify the mating system
of the residual trees and result in increasingly in-
bred seeds through selfing or mating between
related individuals (Ghazoul, Liston and Boyle,
1998; Obayashi et al., 2002), compromising the
population as a quality seed source. In such cases,
sourcing FRM from further away, albeit from simi-
lar ecological conditions, may be a better option
than resorting to nearby fragmented or (geneti-
cally) degraded forests or isolated trees.
However, any introduction of genetically dis-
tinct FRM, even of native species, holds risks.
If FRM is not adapted to the conditions on the
restoration site, severe consequences may result,
such as low seed germination or mortality of the
plants before reproductive age. Alternatively, and
probably more typically, lack of or poor adapta-
tion to site conditions may be expressed more
gradually, for example through slower growth
or lower survival rates. The first generation of
trees plays a key role for subsequent natural re-
generation on site, and low genetic diversity in
this founder population may result in deterio-
rating genetic health and fitness over following
generations (Reed and Frankham, 2003; Rogers
and Montalvo, 2004; see also Insight 1: Examples
illustrating the importance of genetic considera-
tions in ecosystem restoration). If species being
introduced are the same as or closely related to
the species remaining on the restoration site but
from genetically distinct sources, there is an ad-
ditional risk of genetic contamination (Ellstrand
and Schierenbeck, 2000; Ellstrand, 2003; Rogers
and Montalvo, 2004). Gene flow between native
resident populations and non-local introduced
plants might lead to outbreeding depression. This
refers to a situation where, after repeated crosses
between local and introduced provenances, hy-
brid progeny show lower fitness than local prog-
eny because of the breakup of co-adapted gene
complexes by recombination. The phenomenon
of outbreeding depression is commonly discussed,
although there is still little hard evidence of its
effects in trees (Frankham et al., 2011). This might
be because its effects may emerge only after sev-
eral generations (Rogers and Montalvo, 2004)
or because many tree species have regular long-
distance dispersal events, resulting in sufficient
gene flow to avoid complete genetic isolation of
populations even when they are geographically
distant from each other (Ward et al., 2005; Dick

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et al., 2008). Therefore, outbreeding depression
seems most likely to be a risk only where FRM is
introduced from provenances very remote or iso-
lated from the local one.
Other than the geographical origin of FRM,
genetic considerations seem to receive relatively
little attention in restoration efforts, in spite of
their ecological and economic importance (in
terms of return on investment; Le et al., 2012).
Descriptions of restoration methods do not con-
sistently incorporate guidelines for the collection
of FRM, for example on the number of trees from
which FRM should be collected or the distance
between seed trees. Some of the guidelines that
do exist may not be adequate; for example, a rec-
ommendation to collect seed from (at least) five
trees that we came across during this study clearly
falls short. Moreover, even when adequate guide-
lines are available, they may not be followed for
practical or other reasons. Contributors to this
study indicated, for example, that good parent
trees may already have been removed from the
landscape or may be difficult to access, and that it
is often very difficult to get people to collect FRM
from more than one tree per species even when
forest areas remain near the restoration site.
Adherence to the guidelines may be even more
difficult to evaluate and ensure when collection
of FRM is outsourced.
Rules of thumb have been developed for how
many samples one should collect to capture at
least 95 percent of genetic variation (measured
as alleles) with the least amount of effort. Such
rules relate to many factors, such as breeding
system, pollination system, flowering and seed
characteristics (Dvorak, Hamrick and Hodge,
1999; Brown and Hardner, 2000). For example,
Brown and Hardner (2000) estimated that some
59 unrelated gametes are required to obtain
95percent of the alleles in a local population, but
twice as many gametes are needed if alleles at
different loci are represented at equal frequen-
cy. Translated into practical terms, this means
that in a completely outcrossing species at least
30–60 randomly selected trees should be sampled
(Rogers and Montalvo, 2004). When outcrossing
species are open-pollinated and many seeds per
plant are available (>30), seed should be collected
from at least 15 trees. If the number of seeds per
plant is restricted, or there is evidence of mating
between full siblings, then seed should be col-
lected from more trees. If there is evidence of
substantial self-pollination, a minimum sample
of 60 trees is recommended (Brown and Hardner,
2000). Smaller samples are very likely to result in
genetic erosion, whereas collecting more than
the minimum sample size is recommended when
the aim is to maintain genetic diversity (Rogers
and Montalvo, 2004). In general, a few seeds
from many trees is genetically a more efficient
sample of the diversity within a population than
many seeds from a few trees (Brown and Hardner,
2000). A number of general guidelines already ex-
ist and could be adopted for collecting FRM, such
as those published by The Australian Network for
Plant Conservation Inc. (Vallee et al., 2004), the
University of California (Rogers and Montalvo,
2004) and the World Agroforestry Centre32
(ICRAF; Kindt et al., 2006). Broad-scale guidelines
for the collection of FRM are generally widely ap-
plicable and need to be better communicated to
restoration practitioners. At the same time, it is
important to recognize that the extent and dis-
tribution of genetic diversity varies widely among
tree species (see Chapter 7), meaning there is also
a need for more ad hoc guidelines.
If properly designed, individual restoration ef-
forts could also contribute to higher-level goals
– in particular to the provision of FRM for future
restoration efforts, and the conservation of native
tree species and their genetic variation. Such out-
comes merit further consideration by restoration
practitioners and researchers. Restored forests
may later become seed sources for further res-
toration, both in the landscape through natural
seed dispersal and through collection of FRM. This
aspect should be taken into consideration when
planning restoration, especially for rare, endemic
32 Also see http://www.worldagroforestry.org/resources/
databases/tree-seeds-for-farmers for additional manuals and
guidelines.

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or endangered species for which the availability
of appropriate FRM is often already very limited.
Maintaining records on the sources of FRM is es-
sential, as it will inform decisions about future
collection sites for restoration materials. Such re-
cords will also allow lessons to be learned as the
restored forests mature and will permit evalua-
tion of the fitness of the populations (Rogers and
Montalvo, 2004). The objective of establishing a
future seed source provides further impetus for
ensuring a sound genetic basis of initial FRM.
Efforts should be made to avoid the conventional
risks of “seed collection chains” (e.g. Lengkeek,
Jaenicke and Dawson, 2004; Pakkad et al., 2008)
whereby successive use for seed collection of
planted stands with low genetic diversity exacer-
bates the effects of a narrow genetic base in sub-
sequent populations. Increased use of certifica-
tion schemes, such as that of the Organisation for
Economic Co-operation and Development (OECD)
scheme for certification of FRM (OECD, 2011), and
the associated guidelines and protocols should be
promoted in this respect, because they not only
ensure systematic record-keeping but also trace-
ability of germplasm movement.
Examples included in this study illustrate how
conservation of genetic variation is often well
integrated in restoration efforts that focus on
particular species (see, for example, sections 15.1
and 15.2). In such cases, restoration activities are
based on analysis of the genetic diversity of a spe-
cies within the country or across its distribution
range in order to identify distinctive populations
and optimal seed sources for each restoration
site. Such approaches are particularly valuable for
conserving and multiplying the genetic resources
of adaptively or historically distinct populations
and thus conserving adaptive capacity within a
species – even for widespread species that are not
threatened. However, lack of knowledge about
the extent and distribution of genetic variation in
all but a few widely planted, mainly commercial,
tree species, especially in the tropics, currently
constrains adequate planning of targeted species
restoration.
On the bright side, it is becoming feasible to
conduct genetic analyses for increasing numbers
of species as costs decline and new techniques
proliferate. As availability of genetic data for a
larger number of species increases, it will become
possible to design restoration efforts in such a
way that they also contribute to conservation of
genetic variation in target species. However, it
would be unrealistic to assume that genetic data
will become available soon for all relevant native
species in restoration projects. In the meantime,
guidelines about how to safely extrapolate knowl-
edge about the genetic diversity of well-studied
species to broader groups of plants with compara-
ble characteristics would be very helpful to guide
decision-making processes in restoration pro-
jects. Certain characteristics, such as reproduction
mode, breeding systems, and means of seed and
pollen dispersal (called life-history traits; Hamrick
and Godt, 1990, 1996), have been shown to cor-
relate with patterns of genetic diversity. In the
absence of direct genetic information, such traits
may provide some guidance about the genetic
structure of species, especially in species-rich tropi-
cal forests (Rogers and Montalvo, 2004; Vranckx
et al., 2012). As such, genetic patterns recorded
for well-studied species could be generalized to
other species with similar life-history traits to for-
mulate recommendations, for example on the col-
lection of FRM for capturing adequate divers ity,
risks of fragmentation for the quality of natural
seed sources or required population densities in
mixed stands for groups of species. Restoration
efforts and research are recommended to strive
to understand and account for differences in pat-
terns of genetic variation between species groups
in order to more effectively capture adequate di-
versity for establishing viable tree populations for
a full range of native species. However, care has
to be taken not to over-extrapolate observations
made in one particular geographical area or spe-
cies (Duminil et al., 2007). For example, in some
cases, genetically highly diverse tree populations
identified in single-species studies were found
to coincide with areas that also had the highest
species and ecosystem diversity (e.g. Gallo et al.,

GENET IC CONS IDERATI ONS IN ECO SYSTEM RE STORATI ON USING N ATIVE TRE E SPECIE S
249
2009), whereas in other cases quite the opposite
was observed, with genetic diversity in wide-
spread species not being congruent with species
richness (e.g. Taberlet etal., 2012).
16.1.1. Needs for research, policy
and action
• Quantify the risks associated with genetic
mismatching resulting from the use
of narrow or exotic genetic diversity,
including long-term studies. Identify the
critical thresholds for genetic diversity in
restoration material and the key variables
for well-matched sources of FRM. Studies
should be initiated in systems that are
simple in terms of species and structural
diversity to facilitate understanding of
genetic and ecological interactions.
• Develop and promote decision-support tools
for collecting germplasm for restoration that
consider the variation in genetic patterns
among tree species and ways of predicting
it for lesser-known species based on life-
history traits of species. Such tools should
allow determination of whether remaining
populations of a species in the landscape
are likely to contain adequate diversity
for sourcing good-quality FRM, and how
to identify alternative or complementary
sources of FRM when necessary.
• Create wider awareness among restoration
practitioners about the risks of using FRM
with a narrow genetic base. Promote the
adoption of national or international
certification schemes, standards and
guidelines for collecting FRM and
documenting its origin.
• Promote awareness of the potential
of individual restoration projects to
contribute to species conservation and
serve as future seed sources, especially
for rare, endemic and endangered tree
species. Develop approaches and tools for
planning, coordination and communication
of restoration activities that support these
objectives.
16.2. Species selection and
availability
Survey results: Lack of FRM of native tree species
was the most common constraint to the wider
application of the various restoration methods.
Availability of FRM was limited, above all, by lack
of knowledge of species biology (e.g. phenology
or propagation methods) and lack of populations
of the target species in the vicinity of the restora-
tion area. Availability of FRM and ease of propa-
gation and cultivation were the most important
reasons for the choice of species after the suc-
cessional characteristics of the species, and were
considered more decisive than, for example, func-
tional characteristics or conservation status of
the species. Most respondents implied that FRM
was collected and nursery seedlings were raised
as part of the restoration effort. One out of four
respondents reported that exotic species were
regularly used. The most common reasons for the
use of exotic species were their functional charac-
teristics or product preferences.
In spite of the growing recognition that native
species are best for ecosystem restoration, their
wider use often seems to be constrained by the
lack of knowledge about their biology, such as
phenology or propagation techniques, and diffi-
culties in sourcing FRM. Limitations in knowledge
may be so severe that they compromise the opti-
mal selection of species for restoration and result
in including exotic species because native alterna-
tives are not known. Yet the Leguminosae fam-
ily, for example, is known to comprise more than
23 000 species,33 many of which are nitrogen-fix-
ing, which implies wide-ranging possibilities of
using native legume species for site amelioration
in virtually any area targeted for restoration.
While large gaps remain in knowledge about
the biology and ecology of most native tree spe-
cies, especially in the tropics, it is noteworthy
that a considerable amount of useful informa-
tion about many species has already been col-
lected over the years in various studies and field
33 http://www.theplantlist.org

THE STATE OF THE WORLD’S FOREST GENETIC RESOURCES – THEMATIC STUDY
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visits. Much of this information is hidden in grey
literature, and often written in language that is
inaccessible to those working directly in the field
(Boshier et al., 2009). Similarly, several contribu-
tors to this study indicated that when they began
to develop their respective restoration methods,
there was a preference for species that were com-
monly available and for which some information
already existed. As knowledge on (other) native
species in the area improved, it became possible
to select species more in accordance with their
functional characteristics, competitive abilities
or other desirable properties. Existing local and
traditional knowledge about species, their propa-
gation and management can be an important
information source and should be better docu-
mented and integrated in restoration efforts
(Douterlungne et al., 2010).
The limited number of native species for which
information is available from long-term studies
and adaptive management, as well as the difficul-
ty of access to that information, forces many res-
toration practitioners to compromise species se-
lection or conduct their own ad hoc experiments.
Enhancing the public availability of information
on species with potential in restoration efforts
would considerably benefit the community of
restoration practitioners and researchers. To have
the greatest impact, information should be made
freely accessible and easily searchable. Translation
into non-specialist or local languages should be
considered, at least for the most important species
in each area. Some publicly available databases
and tools with information relevant to restoration
already exist. The Agroforestree database34 and
the Useful Tree Species for Africa,35 both by the
World Agroforestry Centre, include information
on propagation and distribution of hundreds of
tree species. The Tropical Restoration Information
Clearinghouse (Environmental Leadership and
34 http://www.worldagroforestry.org/resources/databases/
agroforestree
35 www.worldagroforestrycentre.org/our_products/databases/
useful-tree-species-africa
Training Initiative, Yale University)36 has devel-
oped annotated literature lists relevant to res-
toration, including grey literature, as well as
information on dozens of restoration projects.
The Seed Information Database37 of the Royal
Botanic Gardens, Kew, includes information on
optimal germination protocols and other traits
such as seed storage behaviour on more than
11 000tree and shrub species (July 2012), and the
UK Germination Toolbox,38 also from Kew, pro-
vides detailed information on germination for
native species in United Kingdom.
Exotic species are recognized by many resto-
ration practitioners as important under certain
circumstances (Newton, 2011; Alexander et al.,
2011a; Lamb, 2012). Many exotic species are
known to grow well under harsh conditions and
to improve site conditions, for example through
biological nitrogen fixation. Therefore, they are
often included as nurse crops in the choice of ini-
tial species to ameliorate the microenvironment
on very degraded sites and to facilitate the later
introduction of native species that may be less tol-
erant (see, for example, sections 13.1 and 14.6).
In particular, late-successional species may need
the protection of nurse plants against drought,
direct solar radiation or drying winds, and may be
unable to establish on sites where soil structure,
chemistry or hydrology differ considerably from
natural forest ecosystems. In some cases, planting
a non-invasive, non-persistent exotic species may
be a more ecologically compatible solution than
choosing an ill-adapted population as a source
of a native species that may genetically contami-
nate resident populations (Rogers and Montalvo,
2004). Exotic species may later be intentionally re-
moved (e.g. before they reach reproductive age),
or may be outcompeted by developing vegetation
once the native vegetation is well established.
In restoration or rehabilitation projects that
include production objectives, one frequent mo-
36 http://reforestation.elti.org/
37 http://data.kew.org/sid/
38 http://www.kew.org/science-research-data/databases-
publications/uk-germination-tool-box/

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tivation for using exotic species is preference for
specific products, for example valuable timber or
fruits. Common perceptions of superior growth
or functional characteristics of exotics compared
with native species are at least in part related
to better knowledge of some traditional socio-
economically important species than of native
species. However, according to the contributors to
this study, the use of exotic species was commonly
motivated by their specific products or services,
and seldom only by a lack of information on or
availability of native species. When native species
were highly valued in restoration efforts, lack of
information or FRM on those species may have re-
sulted in their use but in lower numbers, or in use
of native species less suited to the site conditions
or restoration objectives, rather than resorting to
exotic species.
Introduction of exotics in restoration projects
should be carefully planned and based on knowl-
edge of the species and their characteristics. This
is particularly important in the light of climate
change, since changes in habitat conditions may
alter species interactions, for example worsening
problems related to invasiveness (Hulme, 2012;
Lamb, 2012). Research is needed to better under-
stand the ecological and socioeconomic trade-offs
between exotics and native species in a variety of
contexts, and more specifically the factors that
currently limit the use of native species, includ-
ing lack of knowledge on propagation methods,
availability of FRM and limits imposed by people’s
prejudices.
By far the most common planting material in
restoration projects seems to be nursery-grown
seedlings, and the possibility of using optimal
species combinations and FRM that are both
adapted to site conditions and genetically di-
verse appears often limited by what is available
in nurseries. Nursery operations range from very
small-scale roadside enterprises to large, efficient,
commercial suppliers, and may be commercially
driven or directly linked to projects, local com-
munities, institutions or the state. In commercial
nurseries the range of available species, particu-
larly native species, is generally low, largely be-
cause the production of species is often guided
by general demand, supply and ease of propaga-
tion (see Insight 6: Seed availability: a case study).
Nurseries, private or public, have limited resourc-
es and logically try to minimize costs associated
with collecting FRM of species that are not nor-
mally produced or are from remote areas. Storing
and growing FRM with a wide genetic base for a
large number of non-commercial species can be
extremely expensive. As it can take several years
from collection until the material is ready for
transplanting, nursery operations need to know
what future demand will be, or to predict future
demand, which involves considerable risk.
Often, little information is available on the
origin of germplasm in commercial nurseries
and how it was collected (see Chapter 3 and
Chapter9).
Uncertainty about the identity of FRM may in-
crease the risk of genetic mismatching, as the ma-
terial may represent too narrow a genetic base or
be genetically too distant, increasing the risks of
maladaption to the target site and genetic con-
tamination (Rogers and Montalvo, 2004). Where
data are available, collection practice often tar-
gets the “lowest hanging fruit” of easily accessed
trees, which may result in genetic bottlenecks
(Lengkeek, Jaenicke and Dawson, 2005; see also
Lee, 2000; Hai et al., 2008) and associated reduced
fitness in planted trees.
Any restoration project that uses nursery seed-
lings should, where possible, include a nursery
strategy from the outset and integrate the costs
and time required for the development of FRM.
This would help to avoid dependency on com-
monly available FRM, allow collection standards
on genetic diversity to be met and provide suf-
ficient time for propagating the germplasm us-
ing methods tailored to the project. The selection
of species can then be better guided by analysis
of the natural vegetation and the current and
future habitat conditions rather than being sub-
ject to the vagaries and practicalities of supply.
Enhancing the establishment and use of commu-
nity nurseries could contribute both to the use of
an increased number of local tree species and to

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an improved sense of ownership of restoration
activities among local people. Many rural people
have a deep knowledge about their environment,
including habitat and growth requirements, phe-
nology and usefulness of locally available species.
Current initiatives to develop and test community
nursery guidelines and protocols through partici-
patory research with local practitioners should be
strengthened and intensified to ensure that FRM
produced in local nurseries is of good quality, ap-
propriate to the target site and originates from
a sufficient number of genetically diverse par-
ent trees. Certification systems such as the OECD
scheme for certification of FRM (OECD, 2011) ex-
ist to ensure the quality of FRM and reduce the
risk of its uncertain origin, and their use should be
promoted in ecosystem restoration.
In addition to improving nursery strategies,
considering a wider selection of different restora-
tion approaches and types of FRM could help to
include a more diverse set of species in restora-
tion efforts. For example, seed banks store large
amounts of seed of many species and sources, and
many of them also supply seed for restoration or
afforestation purposes. Some larger restoration
projects have established their own seed banks,
to compile and securely store diverse germplasm
until required (see Insight 7: The role of seed
banks in habitat restoration).
Creating demand for good-quality germplasm
of native tree species through political commit-
ments, supportive and regulatory frameworks is
necessary for effecting large-scale changes in the
production and supply chains of FRM for restora-
tion. Such frameworks should explicitly address
species and germplasm selection and the role of
native tree species in ecosystem restoration. For
example, development of zoning systems for
sourcing FRM and mechanisms for their imple-
mentation could result in more consistent use of
appropriate germplasm in restoration projects.
When developing such supportive and regulatory
frameworks, appropriate financing mechanisms
should also be identified. For example, in many
countries large-scale afforestation projects re-
ceive subsidies from the government. It would be
helpful to analyse the implications of creating or
extending such subsidies to ecosystem restoration
and conditioning them to the use of adequate
germplasm.
16.2.1. Needs for research, policy
and action
• Conduct applied research to understand
the potential of native species to achieve
various restoration objectives in assorted
states of site degradation and ecological
and socioeconomic contexts. Analyse the
ecological and socioeconomic trade-offs
related to the use of exotic versus native
species, and the factors that currently
constrain wider use of native species.
Develop knowledge-based decision-support
tools for identifying the conditions under
which the use of exotic species in ecosystem
restoration can be considered beneficial and
justified, or risky and best avoided.
• Improve access to information that is
relevant for the restoration community,
particularly data on the biology and ecology
of native species. Encourage restoration
researchers and practitioners to share
information and contribute results to
publicly available databases, and develop
new decision-support tools for facilitating
the selection of species and restoration
methods. Ensure access to information
by local restoration practitioners, farmers
and other stakeholders by developing and
promoting appropriate communication
technologies and products and provision of
information in locally relevant languages
that uses easily understandable terminology
and accessible formats.
• Raise awareness among restoration
practitioners of the need for early planning
of appropriate and adequate germplasm
supplies of desired species, including
the associated time and costs. Envisage
best ways to embed collection of FRM
and nursery production in projects from
the outset. Improve documentation of

GENET IC CONS IDERATI ONS IN ECO SYSTEM RE STORATI ON USING N ATIVE TRE E SPECIE S
253
collection and propagation of FRM as well
as communication channels, cooperation
and feedback loops between seed suppliers,
nurseries and restoration projects.
• Analyse the needs and options for support
and regulatory frameworks tailored to
the restoration of forested ecosystems
and production and supply of FRM. Such
frameworks should explicitly address the
role and use of native species and the
minimum set of genetic considerations that
should be taken into account, the role and
knowledge of local communities and other
stakeholders, and capacity strengthening
of local nurseries and seed companies,
as appropriate. Develop and implement
the frameworks based on needs analysis,
ensuring adequate financing for the
activities.
16.3. Choice of restoration and
propagation methods
Survey results: Nursery seedlings were by far
the most common FRM used across the range of
restoration methods, followed by wildings and
seeds. With few exceptions, the respondents indi-
cated that nursery seedlings were very often used
as FRM in the restoration method they presented.
Wildings were also used as FRM in most of the
cases, and in half of the cases direct seeding was
sometimes applied. Although vegetative propa-
gation was mentioned it was not often used.
There was a reliance on natural regeneration
when constraints to regeneration (e.g. excessive
grazing) were solved.
Nursery seedlings are very commonly used
in restoration efforts, and may currently be the
most feasible propagation method for many
species. However, if direct seeding or vegetative
propagation can be used, these techniques are
less expensive and generally require less labour
and care than production and planting of nurs-
ery seedlings (see Chapter 8). The nursery phase
generally seems to receive much attention in the
production of seedlings, while crucial aspects of
identifying the preferred origin or provenance of
germplasm, collection methods and tending of
seedlings on site after planting are often over-
looked. Selection pressures on seedlings in nurs-
eries are quite different from those in a degrad-
ed forest. Many more seedlings can reach older
life stages in nurseries than would in the forest,
where selective pressures are often more severe.
Selection in the nursery that mimics natural selec-
tion in degraded ecosystems can filter out inbred
individuals and produce healthy, vigorous seed-
lings. On the other hand, some of the selection
exercised at nurseries may unnecessarily (and un-
intentionally) narrow down or cause directional
changes in the genetic diversity among seedlings
(Gillet, Gömöry and Paule, 2005). It is not well
understood to what extent characteristics such
as slow shoot growth at seedling stage relate to
genetic diversity and viability, and what the ge-
netic consequences may be of systematic removal
of certain phenotypes from the nursery seedling
pool. Increasing the genetic diversity of planting
stock used in nurseries can help ensure that seed-
lings are fit to survive and establish in the degrad-
ed ecosystem in sufficient numbers. Special care
should be given to avoid unintentional selection
of traits during harvest of seed for propagation.
For example, seed dormancy and seed shattering,
which can be important adaptive traits in plants,
are often selected against and lost unintention-
ally under standard seed harvesting and propaga-
tion practices (Cai and Morishima, 2002). Growth
rate and timing of flowering and fruiting are
other traits that are subject to unintentional se-
lection. Harvesting seed in a narrow time window
can reduce genetic variation in terms of timing of
flowering if there is wide genetically controlled
variation in flowering and maturation of seed in
the parental population. Harvesting seed towards
the beginning or end of seed maturity may simi-
larly result in genetic shifts in the trait (Rogers
and Montalvo, 2004).
Restoration researchers consistently highlight
the need for careful assessment of site condi-
tions and factors that restrict regeneration before

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choosing an appropriate restoration method and
set of species (e.g. FORRU, 2006; Holl and Aide,
2011; section 14.1). In sites with low to interme-
diate levels of degradation, where soils are largely
intact and there are sufficient genetic sources for
the next generation (e.g. parent plants or soil
seed bank), natural regeneration may be the best
choice (Chazdon, 2008). If natural regeneration
potential is good, it may suffice to tend natural
seedlings already occurring on the site and to con-
trol suppressing factors such as competing vegeta-
tion or grazing by animals (see sections 12.3, 12.4,
14.1, and 15.3). Seedling recruitment can be pro-
moted by planting fast-growing species to attract
natural seed dispersers to the site and to provide
shelter for other species that require some degree
of shade (FORRU, 2006; section 12.2). Research has
shown the effectiveness of this method in recruit-
ing a large number of seedlings and species to
the site (Sinhaseni, 2008). In such conditions, in-
vesting in large-scale production of nursery seed-
lings and planting may not be an efficient use of
resources. Natural regeneration bypasses the risks
associated with introducing FRM and hence pro-
motes maintenance of genetic integrity (Rogers
and Montalvo, 2004). Additionally, this approach
constantly allows the best-adapted seedlings to
be recruited and hence builds in some capability
to cope with changing environmental conditions.
However, natural regeneration may be susceptible
to genetic bottlenecks where seed trees arefew.
Natural seed sources, whether on site or from
nearby forest areas, must be sufficiently large and
diverse to provide a sound genetic basis for the
restored tree populations. Seed produced by trees
in small fragmented forest patches in a non-forest
landscape matrix may be of poor quality because
of inbreeding and genetic drift (Lowe etal., 2005;
Eckert et al., 2010; but see Chapter 2). Much de-
pends on the pollination and mating system of the
tree species in question. For example, the genetic
consequences of fragmentation may differ be-
tween wind-, animal- or insect-pollinated species
and between self-incompatible and self-compati-
ble species (Vranckx et al., 2012; Chapter 4). So far,
few studies have examined the genetic composi-
tion of tree populations restored through natural
or artificial recruitment processes (Pakkad et al.,
2004, 2008; Burgarella etal., 2007; Navascues and
Emerson, 2007; Liu et al., 2008; Broadhurst, 2011;
Ritchie and Krauss, 2012; s ection 15.4). Research
is needed on this topic for landscapes in different
states of degradation and for different species to
provide guidance for restoring viable, genetically
diverse populations through natural regenera-
tion and seed dispersal.
In sites where diverse natural seed sources are
lacking, or seed sources suffer from genetic ero-
sion, introducing additional FRM (and genetic
diversity) may either be advantageous (Rogers
and Montalvo, 2004) or simply the only solution,
at least in the short term. In such areas, more re-
search is needed on alternatives to raising nurs-
ery seedlings, as well as on using a mixture of
different FRM (Chapter 8). Little is known about
the most suitable propagation methods for the
majority of tropical tree species. While it is es-
timated that 80–90 percent of all plant species
produce orthodox seed, the proportion is clearly
lower among tree species in the humid tropics,
where only 50 percent of plant species may have
orthodox seed (Tweddle et al., 2003; Kettle, 2012)
and are thus more difficult to propagate from
seed. Future research should seek to determine
which types of species are suitable for restoration
through direct seeding under what conditions
(FORRU, 2006: 62) and how the effectiveness of
direct seeding can be improved. Seed banks usu-
ally carry out research for developing optimal
germination protocols, and may be both valuable
sources of information and excellent research
partners on species propagation for restoration
purposes (Hardwick et al., 2011; see Insight 7: The
role of seed banks in habitat restoration).
Vegetative material can be a valid alternative
or complement to seedlings for species that are
easily propagated vegetatively, especially when
FRM is otherwise scarce or lacking (Chapter 8), as
may be the case for rare and endangered species.
However, suitability for vegetative propagation,
for example, the rooting ability of cuttings, may
vary between genotypes. There does not appear to

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be any evidence that a species’ ability to reproduce
vegetatively is associated with lower levels of ge-
netic diversity, and clonal plants seem to display in-
termediate levels of diversity (in the overall range
of genetic diversity for plant species) (Ellstrand
and Roose, 1987; Rogers and Montalvo, 2004).
However, at population level, and especially for
planted material such as living fencerows, there is
evidence of reduced genetic diversity. It is impor-
tant to ensure that the source material for vegeta-
tive propagation is of adequate genetic diversity
to avoid inbreeding depression in subsequent gen-
erations and ensure future viability and resilience
of the stand (e.g. reduce disease susceptibility).
Ultimately, FRM must be chosen to match the
environmental conditions, including the level of
degradation of the restoration site. Germination
and initial establishment of seedlings are the
most sensitive phases of the whole development
of the restored forest ecosystem. In some cases,
changing climatic conditions and moderate to se-
vere site degradation could modify the environ-
mental conditions of a restoration site in such a
way that only nursery-grown seedlings or saplings
that have already survived through some of the
early critical phases can ensure regeneration.
16.3.1. Needs for research, policy
and action
• Carry out research to develop and test
suitable restoration and propagation
methods and decision-making tools for a
variety of native tree species and states of
site degradation. Research should include
analysing the genetic composition of tree
populations restored through different
approaches and comparisons with existing
tree populations in the surrounding
landscape or in similar conditions further
away to help ensure the genetic integrity
of the restored plant communities. Phased
analysis to track the development of genetic
diversity at restored sites over multiple
generations would help to refine guidance
for good nursery practice, for example, by
assessing the long-term fitness consequences
for different species of using seedlings
produced under the relaxed selective
environment of the nursery.
• Create awareness of the importance of
carefully evaluating site conditions as a basis
for choosing the restoration approach that
best addresses the causes of degradation
and the types of FRM most likely to ensure
successful establishment of viable tree
populations and most efficient in terms of
use of resources.
16.4. Restoring species
associations
Survey results: Only a third of the respondents
indicated that the restoration methods they had
used deliberately considered restoration of spe-
cies associations or symbiotic relationships.
Most restoration efforts appear to focus ex-
plicitly on restoration of the tree component in
forest ecosystems. Such focus may be based on
the general perception that trees are commonly
foundation species in the ecosystems where they
occur, facilitating the occurrence and evolution
of other less prominent organisms (Lamit et al.,
2011). However, this overlooks the fact that dur-
ing their growth and development trees them-
selves interact with and depend on many other
species, including pollinators, seed dispersers,
herbivores and symbiotic organisms such as my-
corrhizal fungi or nitrogen-fixing bacteria. There
is increasing awareness that genetic variation in
one species affects that in another species, result-
ing in complex co-evolutionary processes within
entire ecosystems (community genetics; Whitham
et al., 2003, 2006). In forest ecosystems such rela-
tions may arise, for example, when bird or mycor-
rhizal fungal species preferentially associate with
particular genotypes of a tree species. In some
cases, species and genotype relationships may
have significant impacts on successful establish-
ment of a population (Ingleby et al., 2007) or may
ameliorate negative impacts of abiotic or biotic
stresses such as herbivory (Jactel and Brockerhoff,

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Mycorrhizal and rhizobial symbioses are among the
most obvious beneficial associations to consider
for successful restoration. For example, mycorrhizal
symbiosis is known to improve the vitality of plants
under various stresses (Van Tichelen, Colpaert and
Vangronsveld, 2002; Domínguez-Núñez
et al.
, 2006),
and growth capacity of various tree species planted
in truffle plantations in Central Europe was correlated
with mycorrhizal colonization (Bratek, 2008).
Efficiency of the symbiotic associations is affected
by the identity of both the host plants and microbial
symbionts. Nutrient acquisition or nitrogen fixation
may vary significantly according to the plant–microbe
combinations (e.g. Lesueur
et al.
, 2001). Moreover,
there is increasing evidence that mycorrhizal
symbioses are capable of altering competitive
relationships between plants and consequently
composition of entire plant communities (van der
Heijden
et al.
, 1998; Hart, Reader and Klironomos,
2003). Although natural inocula of mycorrhizal fungi
or rhizobia can usually be found in soils on restoration
sites, their populations may have changed because
of changes in environmental conditions. Mycorrhizal
fungal species found in closed forests may differ
from those in areas under anthropogenic influence
(Öpik
et al.
, 2006). Inoculation of propagules in the
nursery with appropriate mycorrhizal fungi or rhizobia
or by seed treatment can facilitate and accelerate
seedling establishment by increasing the availability
of nutrients and water. Importantly, it can assist
seedlings in the crucial early years of restoration, for
example in overcoming competition with undergrowth
or in tolerating the more extreme weather conditions
and drought found in degraded open areas. There is
clear evidence that inoculation can increase success
rates in plantings (Lesueur
et al.
, 2001; Ingleby
et al.
,
2007). Mycorrhizal colonization can be facilitated by
the choice of potting medium and nursery practices,
such as establishing forest nurseries and storing
seedlings near the ground where they can be more
easily infected by native soil fungi (Nandakwang
et al.
, 2008).
References
Bratek, Z. 2008. Mycorrhizal research applied to experiences
in plantations of mycorrhizal mushrooms, especially in
Central Europe.
In
J.I. Lelley & J.A. Buswell, eds.
Mushroom
biology and mushroom products. Proceedings of the
Sixth International Conference on Mushroom Biology and
Mushroom Products,
pp. 272–286. Bonn, Germany, World
Society for Mushroom Biology and Mushroom Products.
Domínguez-Núñez, J.A., Selva-Serrano, J., Rodríguez-
Barreal, J.A. & Saiz de Omeñaca-González, J.A. 2006.
The influence of mycorrhization with
Tuber melanosporum
in the afforestation of a Mediterranean site with
Quercus
ilex
and
Quercus faginea
.
Forest Ecol. Manag.
, 231:
226–233.
Hart, M.M., Reader, R.J. & Klironomos, J.N. 2003. Plant
coexistence mediated by arbuscularmycorrhizal fungi.
Trends Ecol. Evol.
, 18: 418–423.
Ingleby, K., Wilson, J., Munro, R.C. & Cavers, S. 2007.
Mycorrhizas in agroforestry: spread and sharing of
arbuscular mycorrhizal fungi between trees and crops:
complementary use of molecular and microscopic
approaches.
Plant Soil
, 294: 125–136.
Lesueur, D., Ingleby, K., Odee, D., Chamberlain, J.,
Wilson, J., Manga, T.T., Sarrailh J.-M. & Pottinger, A.
2001. Improvement of forage production in
Calliandra
calothyrsus
: methodology for the identification of an
effective inoculum containing
Rhizobium
strains and
arbuscular mycorrhizal isolates.
J. Biotechnol.,
91:
269–282.
Nandakwang, P., Elliott, S., Youpensuk, S., Dell, B.,
Teaumroon, N. & Lumyong, S. 2008. Arbuscular
mycorrhizal status of indigenous tree species used to
restore seasonally dry tropical forest in northern Thailand.
Res. J. Microbiol.
, 3(2): 51–61.
Öpik, M., Moora, M., Liira, J. & Zobel, M. 2006.
Composition of root-colonizing arbuscular mycorrhizal
fungal communities in different ecosystems around the
globe.
J. Ecol.,
94: 778–790.
van der Heijden, M.G.A., Klironomos, J.N., Ursic, M.,
Moutoglis, P., Streitwolf-Engel, R., Boller, T., Wiemken,
A. & Sanders, I.R. 1998. Mycorrhizal fungal diversity
determines plant biodiversity, ecosystem variability and
productivity.
Nature,
396: 69–72.
Van Tichelen, K.K., Colpaert, J.V. & Vangronsveld, J. 2002.
Ectomycorrhizal protection of
Pinus sylvestris
against
copper toxicity.
New Phytol.
, 150: 203–213.
Box 16.1.
Mycorrhizal and rhizobial symbioses

GENET IC CONS IDERATI ONS IN ECO SYSTEM RE STORATI ON USING N ATIVE TRE E SPECIE S
257
2007). Overall, higher species and genetic diver-
sity improves the stability and resilience of entire
ecosystems and ecosystem recovery to climate ex-
tremes, which is of increasing importance under
environmental change (Gregorius, 1996; Elmqvist
et al., 2003; Muller-Starck, Ziehe and Schubert,
2005; Reusch et al., 2005; Paquette and Messier,
2010; Thompson et al., 2010; Alexander et al.,
2011a; Gerber, 2011; Isbell et al., 2011; Sgro, Lowe
and Hoffmann, 2011), although the relationship
between diversity and associated species can be
complex (Castagneyrol et al., 2012).
The genetic diversity of FRM from genetically
distant sources might have consequences for the
species associations that they are used to restore,
in some cases leading to cascading effects through-
out the ecological community. For example, when
plant populations are introduced with earlier or
later flowering and seed-setting times than the
resident populations, this might have consequenc-
es for associated animal species such as pollinators
or seed dispersers (Rogers and Montalvo, 2004).
Analysis of the dynamics of the main energy chains
within the ecosystem can help to identify the inter-
dependencies of species and traits.
Restoration projects should, as far as pos-
sible, create appropriate conditions to foster
restoration of the interactions and associations
between species and genotypes. This should
both improve success rates for restoration and
promote the biodiversity benefits of restora-
tion projects. Most restoration methods rely at
least partly on natural recruitment of seedlings
to the restoration site, and their success, there-
fore, crucially depends on the restoration of pol-
lination and seed dispersal within the ecosystem
and landscape. Enhancing diverse species asso-
ciations also promotes optimal use of growth
resources at restoration sites. Using pioneer spe-
cies that form effective symbiotic relationships
with mycorrhizal fungi, nitrogen-fixing bacteria
or both could contribute to improving site condi-
tions, provide inocula to assist establishment of
other species and, as such, help ensure successful
restoration. Additionally, other fungi and bacte-
ria that are not necessarily symbiotic can be used
in inocula to trigger bioactivation and recoloni-
zation of soil life, and to promote decomposi-
tion of organic material, thus increasing nutrient
availability. This approach has been successfully
applied on highly degraded soils, such as gold
mine spoils (section 14.6).
Analysis of site conditions as a basis for iden-
tifying appropriate restoration methods should
include the associated species present or required
in the system (e.g. symbiotic or herbivorous spe-
cies or species that compete for habitat), their life
cycles and probable effects on restoration pro-
cesses. This also requires study of the approaches
required for promoting their establishment and
survival and restoring and managing their func-
tions. It is noteworthy that exotics can also con-
tribute to species associations, although there
is a lot still to learn about the implications and
risks related to the introduction of exotics (weeds,
herbivores, mycorrhizal and other fungi, bacteria
and insects) in specific systems. As a general rule,
it is wise to avoid introducing FRM of uncertain
origin and on which information is scarce. Lastly,
it would be particularly valuable to consider inter-
actions of multiple species at the landscape scale,
bearing in mind the potential costs and benefits
to other land uses, including other forest types
(Carnus et al., 2006).
16.4.1. Needs for research, policy
and action
• Analyse the importance and strength of
relationships among foundation species,
associated organisms and their genotypes
and the implications of the relationships
for successful establishment of diverse and
viable tree populations. Identify success
factors and develop practical approaches
and guidelines for restoring species
associations using different restoration
methods and in different ecosystem and
landscape contexts. Develop and test
models for predicting the likely benefits
of restoration to plant-community
relationships, biodiversity conservation, and
ecosystem function and resilience.

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• Raise awareness of the importance of
species associations for the successful
restoration of ecosystem functions and
promote the consideration of species
associations in the planning and design of
restoration projects.
16.5. Integrating restoration
initiatives in human
landscape mosaics
Survey results. Area of application varied widely
within and among methods, but the most typical
size reported for restoration actions was about
10 ha. Two out of three respondents indicated
that landscape connectivity needs to be considered
when applying their restoration method. Land-
scape considerations were most commonly associ-
ated with seed dispersal distances from surround-
ing forests to the restoration site. The majority of
respondents considered carbon sequestration and
restoration of habitats for flora and fauna as the
most common benefits expected from restored
forests, while production of timber, fodder or
fuelwood were considered important by only half
of the respondents. Half of respondents reported
that the restoration methods they used could in
some cases be applied to agroforestry or other
land-use types, integrating livelihood aspects.
Very little is known about minimum viable
population or effective breeding unit sizes of
tree species, especially of those that occur in
species-rich tropical forests where population
densities are often low. However, from theory,
it has been suggested that at least 50 unrelated
reproductive trees are needed to form a viable
population and avoid inbreeding (Brown and
Hardner, 2000; Frankham, Ballou and Briscoe,
2002). Estimates of the typical area of applica-
tion of restoration methods suggest that many
sites for ecosystem restoration may be too small
to sustain viable populations of tree species on
their own. Therefore, it is important to design
restoration projects in a way that connects them
to existing tree populations in the landscape or
to other restoration efforts. Ensuring ecologically
and genetically effective connectivity requires
that mating systems, pollen- and seed-dispersal
distances or landscape permeability to gene flow
are taken into account from the outset of resto-
ration projects. For tree stands to contribute to
gene flow it is important to make sure that FRM
is genetically matched to the remaining (frag-
mented) populations of the same species (Rogers
and Montalvo, 2004). Hence, if FRM is obtained
from outside the target area, the risk of hybridi-
zation with existing populations needs to be con-
sidered (Chapter2 and Chapter 6).
Species pollinated by generalist pollinators are
generally more readily connected within a land-
scape than species with specialized or low-energy
pollinators (Vranckx et al., 2012). It is important
to ensure favourable conditions for pollinator
survival and mobility, especially for the latter
type of pollinators. Connectivity and gene flow
are important for both self-compatible and self-
incompatible species: lack of cross-pollination can
result in increased selfing and inbreeding depres-
sion in the former and in reduced seed set in the
latter. Site conditions should also be attractive to
seed dispersers that promote natural dispersal
and recruitment (Markl et al., 2012). Although
in many cases knowledge of species biology and
ecology may not be readily available, it is impor-
tant to learn as much as possible before initiat-
ing a restoration project. Interaction with indig-
enous and other local human communities can
be very useful and rewarding as they often hold
extensive knowledge of the ecology of their lo-
cal plants and associated fauna. When planning
restoration activities, it may in some cases also be
useful to use historically reconstructed reference
land scapes and ecosystems based on historical
ecology techniques (Egan and Howell, 2005). This
allows restoration trajectories to be anchored in
historical time (Chapter 10).
In many tropical areas, the success of resto-
ration practices depends on the engagement
of local communities (Newton, 2011), not as a
workforce but as true participants and direct
beneficiaries of restoration projects. People will

GENET IC CONS IDERATI ONS IN ECO SYSTEM RE STORATI ON USING N ATIVE TRE E SPECIE S
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invest more and care more for species and systems
that correspond to their own needs and values. It
must be recognized that in some cases local peo-
ple may prefer species for production purposes
rather than ecosystem restoration, may not value
non-timber forest products as highly as is often
assumed, or may prefer exotic species for their
marketability, ease of production or other charac-
teristics. It is noteworthy that half of the respond-
ents of the survey conducted for the preparation
of this study did not consider the restored ecosys-
tems as potentially important sources of timber,
fuelwood or fodder, all of which are products that
may often be of particularly high importance for
local livelihoods. This may be indicative of typi-
cal ecology-oriented objectives in restoration re-
search, whereas restoration practitioners may fo-
cus more on the utilitarian value of the restored
ecosystems. Nevertheless, there is increasing
recognition of the potential of managed ecosys-
tems to provide important ecosystem goods and
services, including carbon sequestration, nutrient
and water cycling and biodiversity conservation
(Thompson et al., 2010). One contributor to this
thematic study pointed out that, in many parts of
the world, local people have long-standing expe-
rience with production systems that incorporate
trees and agricultural plants and in many cases
are undoubtedly more knowledgeable about the
species and appropriate propagation and plant-
ing methods than researchers. He noted that
what may often be missing is the promotion of
the concept that such production systems also
have high potential for ecosystem restoration and
conservation (see Chapter 4).
To improve the feasibility and the socioeco-
nomic value of ecosystem restoration projects,
efforts are needed to better understand and
incorporate local people’s preferences for spe-
cies, land uses and management options, and to
identify how outside interests can contribute to
or interfere with local objectives. Participation of
local authorities is important for continuity and
coherence of restoration projects at the land-
scape scale. In addition, education and training
curricula need to be broadened to raise aware-
ness among conservation biologists and rural
development practitioners about the potential
role of on-farm conservation of biodiversity.
Overall, optimal allocation of restoration efforts
at the landscape level requires close collaboration
and coordination among the various landowners
and users (see Chapter 11). Research is needed on
the best means to ensure that individual restora-
tion projects add value to the landscape in terms
of connectivity between populations and habi-
tats as well as complementarity of land uses and
livelihood strategies of local people. Examples of
potentially useful approaches include the analysis
of the viability of tree populations in providing
products and services in diverse land-use mosaics;
development of distribution maps and analysis
of necessary landscape linkages (e.g. stepping
stones, corridors or favourable land-use matri-
ces) and gaps for target species; development
of species-specific action plans for conservation
through restoration; and regional-scale habitat
corridor initiatives. Finally, landscape restoration
depends on national public policies and politics.
Decisions leading to large-scale forest conversion
are taken by national and state governments, and
these bodies are also able to take decisions that
will reverse the trend. In spite of recent initia-
tives, landscape-scale restoration efforts are still
few and the area of restored forested ecosystems
remains small in comparison with the 13 million
hectares of forest converted to other land uses
each year (FAO, 2006).
16.5.1. Needs for research, policy
and action
• Consistently plan restoration efforts at a
landscape scale and seek to integrate them
into the surrounding land-use matrix or
existing networks of habitat corridors. The
presence of existing tree populations of
target species needs to be explicitly taken
into account to facilitate establishment and
maintenance of viable tree populations.
Develop and promote tools and
opportunities for learning, coordination,
communication and joint decision-making

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among landowners and users on the
allocation of restoration efforts.
• Carry out research on the best approaches
for ensuring that individual restoration
projects benefit from the landscape
context and add value to it in terms of
ecological and genetic connectivity, land
uses and livelihood strategies. Transform
the main findings into practical decision-
support tools for landscape planning.
Researchers should seek to consolidate
the role of production systems and on-
farm conservation in providing ecosystem
goods and services while contributing
to landscape connectivity, and should
analyse the genetic impacts of different
management practices and land-use
patterns on tree populations.
• Advocate among politicians for policy
measures and decisions in favour of
landscape-scale restoration of degraded
forest ecosystems.
16.6. Climate change
Survey results: Two out of five respondents indi-
cated that the restoration methods they use con-
sider effects of climate change at least to some
extent. At the same time, only two of the 23 re-
spondents provided explicit approaches for antici-
pating climate impacts. Climate change was most
commonly related to changes in species compo-
sition, with only two respondents explicitly men-
tioning intraspecific effects.
Climate change is expected to change habitat
and growth conditions rapidly and profoundly
in most regions of the world. Some current com-
binations of climatic and edaphic conditions will
disappear or be strongly reduced in size, while
other entirely new combinations are expected to
emerge. Climate change will have a strong impact
on most restoration activities, yet at present few
restoration practitioners appear to purposefully
consider climate predictions in the design of res-
toration activities.
Several approaches have been suggested to
build resilience to climate change in forest man-
agement and restoration initiatives. Given the
uncertainty of future climatic conditions and lack
of knowledge of the nature and distribution of
adaptive traits of species, a prudent adaptation
strategy at this time may be to increase popula-
tion sizes, enhance species and genetic diversity,
and ensure genetic and geographic connectivity
between different biotic elements of both natu-
ral and cultural landscapes (Ledig and Kitzmiller,
1992). The underlying assumption of this approach
is that high levels of genetic and species diversity
and gene flow, along with large population sizes,
will allow natural selection to shift fitness- related
traits so that populations can adapt to changing
environmental conditions (Thompson et al., 2010).
Larger population sizes also buffer against the risk
of population extinction resulting from extreme
events, such as drought, storms or fire. Many tree
species exhibit a high degree of plasticity in fit-
ness-related traits, which gives a population time
to adapt to changes (O’Neill, Hamann and Wang,
2008; Thompson et al., 2010; Mata, Voltas and Zas,
2012). Tree species generally also have relatively
high genetic variation in adaptive traits, consti-
tuting latent adaptive potential that is expressed
only when conditions change (Doi, Takahashi and
Katano, 2009; Thompson et al., 2010). High di-
versity in FRM could be combined with increased
planting densities or fostering natural recruitment
to increase the absolute amount of diversity in the
seedlings (Ledig and Kitzmiller, 1992; Guariguata
et al., 2008; Chmura et al., 2011) and to anticipate
relatively high mortality rates. The role of gen-
erational turnover is key to the capability of tree
populations to adapt through shifts in standing
genetic variation, and methods to accelerate it –
such as gap creation – may have to be considered.
Although many tree species have high capabil-
ity for gene flow among populations (Ward etal.,
2005; Dick et al., 2008) this may vary in accordance
with species life-history traits such as pollination
patterns (Vranckx et al., 2012). Thus, appropriate
(species-specific) conditions should be created to
ensure genetic connectivity between existing and

GENET IC CONS IDERATI ONS IN ECO SYSTEM RE STORATI ON USING N ATIVE TRE E SPECIE S
261
restored populations to increase the selection
pool and allow better-adapted recruits (contain-
ing new genetic variation from gene flow or new
combinations of existing variation) to enter the
population (Newton, 2011). In a similar fashion,
conditions should be created to promote the mo-
bility and migration of plant and animal species,
most notably pollinators and seed dispersers, to
habitats or microhabitats within or near to resto-
ration sites where environmental conditions best
match their requirements for survival, growth
and reproduction. This can be achieved, for ex-
ample, by facilitating movement across hard
edges such as human infrastructure (e.g. “biod-
ucts” over highways), or by taking advantage of
the topographic heterogeneity of a site being
restored. Connectivity is also key in this context,
particularly in terms of providing migration path-
ways (Rogers and Montalvo, 2004; Newton, 2011).
Degraded forest sites that require restora-
tion typically constitute tough environments for
seed ling establishment and growth. When the
climate simultaneously becomes harsher, natural
or planted propagules experience high selection
pressure. Hence, it is even more important than
before to collect FRM from a large number of
parent trees to maximize genetic diversity (see
Chapter 2). Especially where climate change is
already evident and suitable seed source popula-
tions are available, FRM should be collected from
a range of environmental conditions in the same
or neighbouring seed zone to increase the varia-
bility in adaptive traits and thus enhance adaptive
capacity of the next generations of established
tree populations. This approach relies on the as-
sumption that the species have relatively high ge-
netic diversity in or near the target area; it may
therefore be less relevant for species whose popu-
lations have been severely reduced or for species
with naturally low genetic diversity.
Additionally, the expected direction of selec-
tion could be taken into account, for example by
including FRM from warmer rather than cooler en-
vironments or from drier or wetter environments,
depending on general climate predictions for the
region. Topographic variation in the area could
be taken advantage of when sourcing more di-
verse FRM that may represent adaptations to
varying microclimates. Furthermore, drought
gradients are present at large landscape scales
all over the world, and remnant trees within the
more arid areas could be potential seed suppli-
ers for restoration of the same species in neigh-
bouring areas that are currently still more humid.
Similarly, the gene pool of remaining trees on
sites that are already affected by climate change
can provide a useful seed source for sites with
conditions that are currently less extreme but still
nearing the edge of the species’ tolerance. This is
because such residual trees are survivors and may
be better adapted to the extreme conditions. The
quality of residual tree populations as seed sourc-
es, including the risk of inbreeding and the risks
associated with transfer of provenance, should,
however, be evaluated in all cases.
If provenance trials have been established for
species of interest, it is possible to select prov-
enances that are adapted to the expected cli-
matic conditions of a restoration site (Ledig and
Kitzmiller, 1992; O’Neill, Hamann and Wang,
2008; Wang, O’Neill and Aitken, 2010; O’Neill and
Nigh, 2011). However, provenance and progeny
trials that can provide knowledge about adaptive
traits and climatic or environmental tolerance
within species and populations exist mainly for
introduced, commercially valuable species. They
cover only a small proportion of species of po-
tential interest for restoration, and most of those
trials that do exist do not sample the full species’
ranges. New trials should be established along
gradients of relevant variables, such as eleva-
tion, latitude and aridity, both within and outside
natural distribution areas of a species, and use
various approaches, such as reciprocal transplant-
ing. The main purpose of such trials would be to
understand the extent and patterns of adaptive
variation, rather than just choosing the best-per-
forming provenance. Similarly, interpretation of
results from existing provenance trials originally
established for production purposes should be
studied in the context of ecosystem restoration
and adaptation.

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Importantly, research is needed to understand
climatic tolerance of species and their genetic vari-
ants during the critical phases of tree life cycles,
such as germination and seedling establishment. If
mortality risks during such phases are over looked,
the appropriateness of germplasm to given site
conditions may be misjudged and the success of
restoration compromised. As mentioned before,
an important concern for FRM grown in nurseries
is that it may be exempted from normal selection
pressures during germination and initial establish-
ment. These are precisely the stages during which
trees typically experience the strongest selection
pressures of their life cycle under normal field
conditions. Exempting plants from such selection
pressures may result in poorly adapted seeds and
seedlings in future generations. Tree breeders also
can contribute to enhancing survival chances of
planting stock under changing environmental con-
ditions, for example, by crossing provenances dis-
playing adaptive traits of interest. Possible breed-
ing strategies include: (i)selecting and breeding,
for specialists, varieties that perform particularly
well in specific, defined conditions, and then using
them on the appropriate site types; and (ii)select-
ing and breeding, for generalists, varieties that
perform at least moderately well in a broad range
of environments (Ledig and Kitzmiller, 1992; see
also Chapter2 for other strategies). Given the un-
certainties of expected climate change, breeding
for generalists would be the better choice (Ledig
and Kitzmiller, 1992).
In some cases habitat conditions will be al-
tered by climate change to such an extent that
the classical preferential use of local germplasm
may no longer be valid. If climate conditions in
the area are expected to change substantially, de-
liberate moving of FRM along climate gradients
may need to be considered (Ledig and Kitzmiller,
1992). Examples of such strategies include assist-
ed migration based on predictive provenancing
(Chmura et al., 2011; Chapter 2). Such approaches
can target both provenances and species, and in-
volve either matching of germplasm to a given
restoration site or suitable sites to germplasm of
interest. In mountainous areas planting stock is
typically moved upwards, anticipating that future
climatic conditions at higher altitudes will be simi-
lar to those presently occurring at lower altitudes.
In general, FRM could be moved along latitudi-
nal gradients to help plant communities follow
changing temperature patterns. These selection
strategies need to be combined with a study of
the expected future environmental and climatic
changes and predicted rate of change for the
restoration site to ensure that FRM selected has
the most potential to be resilient to future condi-
tions. As stated previously, care must be taken to
evaluate the risks of provenance transfer. Species
associations such as mycorrhizal symbionts or spe-
cific pollinators should be considered when plan-
ning provenance transfer.
Spatial species distribution models can be useful
tools to identify sources of FRM with potential ad-
aptation to extreme habitat conditions and to pre-
dict the suitability of future climatic conditions for
given species on particular sites (O’Neill, Hamann
and Wang, 2008; Wang, O’Neill and Aitken, 2010;
Sáenz-Romero et al., 2010). Spatial models are al-
ready used to inform approaches to forest land-
scape restoration by indicating those locations
within a landscape where particular restoration
approaches would most likely be successful based
on local environmental conditions (Newton, 2011).
In a similar way, spatial models could be employed
to identify areas and cases where climate change
considerations in restoration may be most impor-
tant, for example, at the retreating edge of a spe-
cies range. The modelling approach allows projec-
tions of regeneration and spread of native forest
under various anthropogenic disturbance regimes,
providing insights into the potential for passive
restoration approaches (Newton, 2011). However,
spatial distribution models should be used with
caution. Their predictions are directly related to
the quality of available data on species distribu-
tion. Lack of distribution data, especially for many
tropical species, and alterations to natural species
distributions caused by human influence limit the
application of such models. In addition, it should
be noted that species distribution models are usu-
ally based on climatic factors only and often do

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263
not take into account other factors affecting spe-
cies distribution, such as soil types, population ge-
netics or species plasticity and adaptive potential
(O’Neill, Hamann and Wang, 2008). Such models
typically do not treat species as dynamic entities
and fail to build in evolutionary processes. Hence,
the potential for in situ adaptive change is not ac-
counted for and predictions of changes in distri-
bution may be over-simplified. Therefore, results
of species distribution models should be ground-
truthed as much as possible before using them in
the design of restoration projects. In sum, distri-
bution models can be useful for obtaining first
approximations of expected species distributions
or site conditions to inform restoration research,
but should not be used as self-standing tools for
decision-making.
As noted above, several precautions should be
borne in mind with respect to moving germplasm
along environmental and climatic gradients. First,
there are clearly risks in the level of confidence
with which predictions of future climates can be
made. Even if the predictions per se proved to be
correct, it is likely that in many cases novel envi-
ronments will emerge that are currently not part
of species ranges, and suitable FRM may, there-
fore, be difficult to identify. Second, in addition
to climate, local adaptation has been shaped by
numerous factors, such as photoperiod, soil con-
ditions, the local biotic environment, and com-
petitive and symbiotic relationships with other
species and their variants, including pest and
diseases (see section 16.4). Therefore, transfer-
ring germplasm on the basis of climatic gradients
alone is usually not to be recommended as it risks
exacerbating fitness deficits, potentially divert-
ing attention and resources from vital efforts to
bolster local population sizes through restoration
and, at worst, causing genetic contamination or
introducing new pests or diseases to existing pop-
ulations. In spite of these precautionary observa-
tions, it should be noted that there are very dif-
ferent perspectives on the benefits of germplasm
movement for matching expected climatic condi-
tions, both among scientists and policy-makers
(Seddon, 2010). In western Canada, for example,
a forest regulation has already been changed to
accommodate new seed transfer rules to better
match seedlings to expected future conditions.
Finally, planting stock that is best adapted to
changing environmental conditions and thus
most suitable for use in restoration projects may
not always be available in the country of imple-
mentation. Hence, the need for cross-border
movement of germplasm will likely have to in-
crease if ecosystem restoration projects are to be
designed to respond most effectively to climate
change. In light of this, there is an urgent need
for countries to re-examine regulatory norms that
currently impede or excessively regulate germ-
plasm movement across political borders (Koskela
et al., 2010).
16.6.1. Needs for research, policy
and action
• Given the uncertainty of future climate
predictions, the most prudent approach
to preparing for climate change for most
restoration efforts is to use as much as
possible of the genetic and species diversity
available near the restoration site or in
sites with similar (macro)environmental
conditions, which will allow natural
selection to take its course and move
the restored population in the required
direction. Restoration projects should
collect forest reproductive material from
a large number of parent trees and from
as many sites as possible with locally
varying (microenvironmental) habitat
conditions. Such approaches should be
used in combination with planning and
management strategies explicitly designed
to promote gene flow and facilitate species
migration. In cases where genetic diversity
is lacking and where impacts of climate
change are already stressing the ecosystem,
assisted migration may be necessary,
taking precautions to match changing
environmental conditions as closely as
possible and to avoid possible associated
risks to local biodiversity in target areas.

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• Conduct research on the extent and
distribution of plasticity and adaptive
capacity in native tree species, particularly
in areas that are especially vulnerable
to climate change, in order to identify
appropriate FRM for restoration that
also maximizes resilience. Develop and
test practical approaches and decision-
support tools for improving ecosystem
resilience through restoration. Establish
provenance trials using seed sources
collected from a stratified sample across
the species’ distribution range, on sites
across environmental gradients within
and beyond current species distributions.
Research should also be designed to test the
feasibility of assisted migration. Modelling
approaches that take into account genetic
diversity and selection seem a promising
approach to yield timely and relevant
results.
16.7. Measuring success
Evaluating the success of any action requires a
clear definition of objectives or baselines against
which performance can be judged. Possible ob-
jectives of ecosystem restoration are at least as
diverse as its definitions, ranging from restoring
tree cover, original vegetation structure and bio-
diversity, to ecosystem functions, services or pro-
vision of livelihoods. One of the proposed, more
holistic goals for restoration is restoring ecologi-
cal integrity, defined as “maintaining the diver-
sity and quality of ecosystems, and enhancing
their capacity to adapt to change and provide for
the needs of future generations” (Mansourian,
2005). Another, probably more dynamic defini-
tion by lead members of the International Society
of Ecological Restoration emphasizes “reinstat-
ing autogenic ecological processes by which spe-
cies populations can self-organize into functional
and resilient communities that adapt to changing
conditions while at the same time delivering vital
ecosystem services” (Alexander et al., 2011b).
Despite an accumulation of experience on eco-
system restoration over the past decades, it is still
quite common to measure the success of refor-
estation and restoration efforts mainly in terms
of number of seedlings planted or their survival
in the short term (Le et al., 2012). Such measures
ignore the importance of using good-quality FRM
that is capable of establishing on the site and cre-
ating functional ecosystems over time and do not
help evaluate achievement of the actual restora-
tion objectives. If the focus on planting targets is
too strong, it may divert attention from the actual
objectives (the establishment of resilient plant
communities) and factors critical to success, result-
ing in inefficient use of resources and wasting of
time. Restoration success needs to be evaluated in
a more holistic way, not only by restoration prac-
titioners but also by government institutions, civil
society organizations, the private sector and, im-
portantly, funding agencies. The fact that genetic
factors are still missing from the recent conceptual
models and otherwise extensive lists of success in-
dicators and drivers (Le et al., 2012) is illustrative
of the scale at which awareness needs to be raised
about the importance of genetics in reforestation
and restoration projects. Genetic variation itself is
an indicator of functional and resilient ecosystems
and hence also the success of restoration activities
(Thompson et al., 2010).
Successful re-establishment of functional eco-
systems can only be truly evaluated in the long
term by covering all the main stages in restora-
tion projects (including forest establishment,
growth and maturation; Le et al., 2012). The
problem is that such assessments extend sub-
stantially outside the time span of most restora-
tion projects. Nevertheless, a plan or strategy
for continuous monitoring of progress towards
set objectives should be an integral part of any
restoration effort to allow for steering and cor-
rective management practices where necessary
throughout the different stages of vegetation
development. Effective monitoring requires the
establishment of a baseline and a set of indicators
that relate to the specific objectives of restora-
tion. Ideally, especially for research purposes, the

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baseline for genetic monitoring should include
the genetic structure of: (i) the remnant trees
of the degraded populations in the landscape,
(ii) their naturally regenerated saplings, (iii) the
source populations of germplasm used, (iv) the
progeny (seedlings) of this germplasm grown un-
der nursery conditions and (v) the mating pattern
in undisturbed and disturbed populations. This
information would allow assessment and a bet-
ter understanding of the changes in the genetic
structure of species throughout the restoration
process, evaluation of the genetic viability of the
progeny and, eventually, assessment of the suc-
cess of restoration on time scales at which fitness
of species can be judged.
The genetic diversity profile of one or more
healthy reference populations occurring natu-
rally (as far as possible) in the same seed zone or
ecological niche should ideally be known for as-
sessing success in restoring genetic diversity. This
could then be compared with the genetic diver-
sity of the developing tree populations under
restoration at various stages in time throughout
the restoration process. Such references for tar-
get levels of genetic diversity may already exist
for some species and areas, but are lacking for
the large majority of species and contexts. Use
of similar or standardized molecular techniques
to assess diversity of restored populations would
facilitate comparability and wider applicability of
the findings, although this is probably not realistic
for the majority of species, as techniques are con-
stantly changing and being improved. In the long
term, databases could be established contain-
ing reference levels of genetic diversity per spe-
cies and for different target areas of restoration.
Genetic assessments of the success of restoration
projects could then be limited to measuring the
genetic diversity of the restored tree populations
and comparing these values with reference values
from the databases. Ideally, such genetic assess-
ments could also be extended to populations of
other, naturally establishing (plant and animal)
species that are representative of certain groups
of species with similar functional, structural or
life-history traits. In some cases it may be difficult
to determine the reference target level of genetic
diversity for species used in restoration activities,
for example when natural populations have been
nearly or completely eliminated. In such cases it
may be necessary to define a baseline rather than
target to allow assessment of the success of resto-
ration activities. In addition to comparing levels of
genetic diversity between restored plant popula-
tions and their natural analogues, it would also
be important to assess the genetic connectivity be-
tween restored and adjacent natural, undisturbed
populations.
Examples of possible indicators that could be use-
ful for evaluating genetic composition of restored
populations include genetic structure and genetic
diversity for forest structure demographic charac-
teristics, and gene flow and inbreeding for forest
function (Newton, 2011). Early detection of genetic
bottlenecks, which may go undetected in traditional
demographic monitoring, is important to avoid po-
tentially harmful effects and relatively easily done,
for example by comparing neutral allele frequency
between different generations (Luikart etal., 1998;
Rogers and Montalvo, 2004; Kettle, 2012). It is im-
portant to note that monitoring changes in genetic
diversity must be framed in a biologically meaning-
ful context so as to be able to interpret whether any
observed changes are within a normal or desirable
range, or whether they might signal some serious
loss that could have negative repercussions (Rogers
and Montalvo, 2004). For example, the loss of se-
lectively neutral traits measured using molecular
markers does not necessarily translate into loss of
adaptive traits (Holderegger, Kamm and Gugerli,
2006). After going through an extended genetic
bottleneck that dramatically reduces population
size and genetic diversity, genetic variation in selec-
tively neutral traits may require many thousands of
generations to recover, whereas recovery of varia-
tion for adaptive traits may require only hundreds
of generations (Milligan, Leebens-Mack and Strand,
1994; Rogers and Montalvo, 2004).
There is emerging consensus that a combina-
tion of ecological and genetic indicators would
provide the best results in genetic monitoring of
forested ecosystems (reviewed in Aravanopoulos,

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2011). However, most restoration efforts can-
not be realistically expected, at least in the short
term, to include molecular studies to assess lev-
els of genetic diversity. Moreover, most native
tree species in the most biodiverse areas of the
world have not been subject to molecular analy-
ses, and most restoration practitioners are not
equipped to carry out molecular analyses, even
for those species that have been relatively well
characterized. Two sets of indicators to evaluate
genetic composition of restored tree populations
are therefore needed: one for situations where
molecular studies are feasible and more detailed
information can be obtained, and another for
situations where such studies are not feasible
and information must be obtained more indi-
rectly. Developing effective surrogate indicators
for genetic diversity for wider application first re-
quires a good understanding of various genetic,
biological, ecological and management processes
and how they may affect genetic diversity during
restoration (Graudal et al., 2013). Priority species
for which to develop surrogate indicators might
include those for which some baseline genetic
data exist, and that are known or suspected to be
particularly sensitive to human influences or envi-
ronmental change (e.g. based on their life-history
traits; Vranckx et al., 2012; see also Jennings et al.,
2001; Rogers and Montalvo, 2004). Studies should
be initiated in ecosystems that are simple in terms
of structure and species composition to facilitate
understanding of the interactions between eco-
logical and genetic processes and consequences
for genetic diversity.
Knowledge from existing studies that looked
at the genetic structure, diversity and connectivity
of restored tree populations in combination with
ecological observations can be used to inform
the development of monitoring guidelines and
choice of surrogate indicators for practical res-
toration. The case study on Pinus radiata D.Don
(section 15.3) illustrates how measuring genetic
diversity of a target species at a degraded site, as
a baseline, can be highly informative for select-
ing optimal approaches for restoration interven-
tions or providing support to chosen methods.
An initial assessment of genetic diversity in the
seem ingly highly degraded P.radiata population
on the island of Guadalupe showed still relatively
high genetic diversity and low levels of inbreed-
ing, and led to limited intervention, simply elimi-
nating the main factor that prevented natural
regeneration (i.e. grazing by goats). Similarly,
a more frequent application of genetic assess-
ments of the success of restoration projects, such
as that performed for Banksia attenuata (see
section15.4), would permit testing and compari-
son of the performance of different restoration
methods for different species combinations and
site contexts. Unfortunately, given the limited
attention to genetic aspects in ecosystem resto-
ration to date, little information is available on
factors related to success and fail ure in restoring
the genetic diversity and adaptive capacity of tree
populations under different contexts and using
different restoration methods.
While the type and amount of genetically rel-
evant information that can be collected in practi-
cal restoration projects may (still) be limited, such
efforts would be important in building a critical
knowledge base on ecosystem restoration that
can, in turn, support restoration research and help
improve restoration guidelines. Global initiatives
could be designed to collect important data from
different restoration sites that could subsequently
be used to better understand the genetic dimen-
sion of ecosystem restoration, conduct meta-anal-
yses across sites, and synthesize general approach-
es for restoration that incorporate genetic criteria
and conservation. Lessons could be learned from
restoration practices across different habitats.
For example, considerable effort has gone into
developing standards and criteria for measuring
the success of restoration of fresh water ecosys-
tems (Palmer et al., 2005) that could be relevant
for restoration activities in other habitats. Among
the data to be recorded at different restoration
sites would be: the location of the source popula-
tion of FRM; environmental description; number
of trees from which FRM was collected; distances
between them; amount of seed per tree; whether
seed was mixed among the trees; year of seed col-

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267
lection and associated climatic conditions; nursery
conditions; size of seedlings; whether any selec-
tion was performed in the nursery; acclimatization
method prior to planting out; and whether natu-
ral regeneration also occurred on site. Enhanced
collaboration between restoration practitioners
and researchers would contribute to better un-
derstanding of genetic diversity and genetic pro-
cesses important in restoring functional and resil-
ient populations of native tree species, and the
associated ecosystem and evolutionary processes.
Lastly, while there is a dire need for better ways
to synthesize and distribute knowledge from suc-
cessful projects for the definition of best practices
in ecosystem restoration, it is also important that
failures in restoration projects are reported more
systematically to help improve future strategies.
16.7.1. Needs for research, policy
and action
• Conduct research for different combinations
of native species, degradation states and
restoration methods to understand how
various biological, genetic, ecological and
management processes interact and affect
ecosystem functions, and the resilience of
genetic diversity during restoration. Develop
protocols for collecting related baseline
information that are widely applicable to
different species and contexts, as well as
sets of genetic and surrogate indicators
that allow assessment of the viability and
resilience of restored tree populations.
• Through the collaboration of researchers and
policy-makers, compile compelling evidence
and advocate for the need to measure
success in restoration projects in ways that
reflect ecosystem functioning and long-term
resilience. Foster collaboration between
restoration researchers and practitioners to
compile information, conduct meta-analyses
and generalize good practices for ensuring
viability of restored tree populations for
functional and resilient ecosystems.
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Part 5
CONCLUSIONS AND
RECOMMENDATIONS

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Chapter 17
Conclusions
Evert Thomas,1 Riina Jalonen,1 Judy Loo,1 Stephen Cavers,2 Leonardo Gallo,1,3 David Boshier,1,4
Paul Smith,5 Sándor Bordács6 and Michele Bozzano1
1 Bioversity International, Italy
2 Centre for Ecology and Hydrology, Natural Environment Research Council, United Kingdom
3 Unidad de Genética Ecológica y Mejoramiento Forestal, INTA Bariloche, Argentina
4 Department of Plant Sciences, University of Oxford, United Kingdom
5 Seed Conservation Department, Royal Botanic Gardens, Kew, United Kingdom
6 Central Agricultural Office, Department of Forest and Biomass Reproductive Material, Hungary
In a world characterized by unprecedented rates
of biodiversity loss, ecosystem degradation and
environmental change, ecosystem restoration is
more important than ever. Considerable efforts
have already been made to restore degraded for-
ested ecosystems globally, supported by experi-
mental and anecdotal research and rapidly ma-
turing new scientific disciplines such as restora-
tion ecology. However, there is a need to further
upscale and mainstream activities. Policy-makers
are increasingly recognizing the potential of eco-
system restoration for mitigating and reversing
a wide range of environmental problems and
associated opportunities for socioeconomic ben-
efits. The most important future challenge will
be to translate the knowledge generated from
research into widespread sustainable practice. It
is imperative that restoration practice develops
strong multidisciplinary approaches that include
a stronger focus on important but previously
neglected factors. The genetic composition (di-
versity and adaptedness) of tree populations in
restored ecosystems is often overlooked despite
its fundamental importance for the success of res-
toration in both the short and the long term. The
objective of this thematic study is to highlight the
breadth and depth of genetic aspects that need
to be considered in ecosystem restoration using
native tree species, and to propose recommenda-
tions for researchers, policy-makers and restora-
tion practitioners to better address the deficien-
cies that may currently compromise the success of
some restoration efforts.
There is a dire need to develop elemental,
practical and convincing recommendations and
guidelines for selecting, collecting and propagat-
ing genetically diverse and appropriately adapted
planting material that is specifically tailored to
ecosystem restoration. General guidelines for se-
lecting and collecting planting material for res-
toration are largely compatible with those tra-
ditionally used in forestry and agroforestry and
should build on them. However, it is clear that
some issues need more emphasis in restoration. In
particular, the traditional rule of thumb that local
seed is generally the best choice when sourcing
planting material for restoration may no longer
be universally applicable. Although local seeds
may still be the preferred option in the absence
of knowledge about the qualities of germplasm
sources, scientific evidence is increasingly showing
that local tree populations may not be adapted
to environmental conditions that are already de-
graded or those that are expected in the future.
Even if local populations are adapted to future
conditions, in many areas they may already be
too degraded or fragmented to constitute good
sources of seed for restoring viable and resilient

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self-sustaining populations. Such fragments may
no longer have the adaptive capacity required
to cope with environmental changes, although
responses may be species-specific. More research
and decision- support tools are needed to help
evaluate the quality of local germplasm as seed
sources. In some cases it may be appropriate to
source seed from more remote, larger populations
growing under ecological conditions similar to the
planting site, but perhaps with climatic conditions
more similar to those expected in the near future.
However, special care should be taken to avoid po-
tentially harmful effects from genetic contamina-
tion of remaining resident populations.
Given the uncertainty of predictions of future
climate and the limitations of current knowledge
about the vast majority of native tree species, for
most restoration efforts the most prudent ap-
proach for anticipating the impacts of environ-
mental change would be to maximize genetic and
species diversity that is well-matched to the tar-
get sites, while at the same time creating favour-
able conditions for connectivity (i.e. for gene flow
and species migration) and regeneration. There is
still considerable controversy about the necessity,
utility, feasibility and risk of deliberately moving
germplasm over long distances to help tree spe-
cies and populations track changing environmen-
tal conditions. The most informative and safest
way to tackle this uncertainty would be to ex-
pand the use of provenance trials to cover a wider
range of environments, more native species and
a wider range of traits (e.g. the consequences of
germplasm transfer for associated organisms are
rarely considered). There is considerable potential
for learning about the performance and adaptive
potential of propagation material from known
sources in already established and planned resto-
ration sites. In addition, increased use of ecologi-
cal zones or other proxies for adaptive variation,
such as environmental gradients, would help res-
toration practitioners chose the best sources of
site-matched planting material.
Another major constraint on diversification of
restoration efforts and optimization of different
approaches to local conditions is the availability
of adequate planting material. At present, the
choice of species and propagation material of
many, if not most, restoration projects is heavily
influenced by what is available in (commercial)
nurseries or that can be easily collected. This may
lead to a mismatch between planting stock and
the conditions at the restoration site, or may even
result in the selection of suboptimal restoration
approaches. First and foremost, adequate politi-
cal incentives and public policies are key for mo-
tivating nurseries, especially commercial opera-
tions, to widen the choice of planting material.
A second priority is to improve restoration practi-
tioners’ and nursery managers’ access to existing
and new knowledge about the genetics, biology
and management of native species. This implies
the development of better tools and communica-
tion strategies to improve information sharing.
In particular, knowledge has to move beyond the
scientific restoration community and academic
journals to reach the much broader community of
restoration practitioners, nursery managers and
seed suppliers. This may require translation of in-
formation into different languages and accessible
styles, increased efforts to raise awareness and
strengthening of capacity specific to the cultural,
socioeconomic and gender context of the target
audience.
A third priority for widening the choice of
propagation materials is to improve communi-
cation channels, cooperation and feedback be-
tween seed suppliers, nurseries and restoration
projects. During the planning stages, restoration
practitioners should inform nursery managers or
seed suppliers about the planting material they
want to use and help them to identify potential
seed sources. Restoration practitioners, if not
linked to the nursery where their planting mate-
rial is produced, should provide feedback to nurs-
ery managers about problems experienced with
their planting material in the field. To compen-
sate for the limited choice of planting material
in existing (commercial) nurseries or seed banks,
or to complement the germplasm that is avail-
able, restoration projects could set up their own
(project or community) nurseries or seed banks to

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produce the desired plant material. In either case,
adequate time needs to be reserved for sourcing
the germplasm, including identifying source pop-
ulations and flowering and fruiting seasons of the
different species.
Record-keeping is essential for good genetic
management in individual restoration projects, as
well as at nurseries and seed banks. Adoption of a
certification scheme such as the OECD Forest Seed
and Plant Scheme would ensure not only system-
atic record-keeping, but also traceability of germ-
plasm movement. Looking to the future, records
of the genetic base and source of materials used
in restoration projects will also help to evaluate
the qualities germplasm and inform decisions
about where to make future collections of plant-
ing material in restored vegetation sites.
Political commitment and supportive regula-
tory frameworks can help promote demand for
and supply of good-quality germplasm of na-
tive tree species. Governments must support and
guide restoration initiatives through carefully
defined policies and financial support. Currently,
few countries have regulatory frameworks and
funding schemes for ecosystem restoration in
place. Strengthening national capacities to re-
store genetically diverse, healthy and resilient
forested ecosystems is also important and should
be supported by integrating genetic aspects into
curricula on biodiversity conservation and resto-
ration, and the provision of appropriate training
to restoration practitioners.
The actions proposed above can be expected
to contribute to increased popularity and use
of native tree species in restoration projects.
Nonetheless, the use of exotics may sometimes
be justified, particularly as nurse plants to facili-
tate subsequent establishment of native species
by improving (micro)environmental conditions in
severely degraded environments, or when native
species with comparable properties have not yet
been identified. However, a preference for exotics
even when native alternatives are available sug-
gests that there are other factors that constrain
the use of natives besides availability and knowl-
edge of their biology. More research is, therefore,
necessary to understand underlying factors that
may affect species selection and the trade-offs
related to the use of exotic versus native species
in different contexts, such as behavioural patterns
or local perceptions of the ecological and socio-
economic value of species.
In addition to their contribution to mitigating
and countering ecosystem degradation, restora-
tion projects also hold great potential for contrib-
uting to biodiversity conservation. Restoration
practitioners should plan to integrate their ac-
tivities in the wider landscape from the outset,
and consider how they can both benefit from
the landscape context and in turn bolster biodi-
versity conservation. A stronger focus on species
associations, such as microbial symbionts and
pollinators, can enhance the chances of survival
of planted or recruited trees, optimize the use
of available resources on site, and increase the
resilience of the plant community to biotic and
abiotic stresses. Genetic connectivity contributes
to the maintenance of population-level diversity
through pollen flow, which reduces the likelihood
of inbreeding and results in new genetic combi-
nations that may be better adapted to changing
environmental conditions. Genetic connectivity
also facilitates migration of populations to more
suitable habitats through seed dispersal. On the
other hand, restoration efforts have considerable
potential to support conservation of threatened
or endangered native tree species, their genetic
diversity and associated species, for example, by
restoring and maintaining genetic diversity of
species across a range of ecological conditions
and establishing genetically diverse sources of
propagation material.
New approaches and tools are needed to en-
able communication and coordination among
restoration practitioners in order to realize the
biodiversity potential of restoration projects.
Policy-makers can play a facilitating role in bring-
ing together stakeholders involved in landscape
planning. This could lead to development of
practical guidelines on how to best organize eco-
logical and genetic connectivity and conservation
at landscape level, which could serve as policy

THE STATE OF THE WORLD’S FOREST GENETIC RESOURCES – THEMATIC STUDY
PART 5
280
support tools. Lessons can be learned from re-
constructing the historical ecologies of particular
target areas, which could guide the design of con-
nectivity across different landscape scenarios. In
this context, clarity of the objectives and species
for which connectivity is sought is key.
Finally, in monitoring and assessing the suc-
cess of restoration it is important that restoration
practitioners systematically integrate genetic as-
pects in ways that reflect ecosystem functionality
and resilience in the long term. Reference levels
against which changes in genetic diversity and
structure of plant populations can be assessed in-
clude baseline genetic diversity at the start and
the target genetic diversity (e.g. by comparison
with known natural populations). There is an ur-
gent need to develop indicators that allow meas-
urement of the success of restoration efforts to
(re-)establish self-sustaining plant communities,
including the likelihood of long-term population
viability of the tree species. Such indicators may
require direct monitoring of genetic parameters
in restoration sites or monitoring of other (less-
costly) variables as proxies for genetic diversity.
Ideally, indicators should be few and designed in
such a fashion that they are based on ecological
or biological measures and can be widely applied
across restoration projects. In general, restoration
interventions should incorporate a monitoring
(and evaluation) strategy that extends beyond the
establishment of seedlings and, ideally, continues
long enough to assess the reproductive success
of species in restored ecosystems. Comparative
research should be carried out to analyse the suc-
cess of various restoration approaches, including
– critically – recording and reviewing failures as
well as promoting methods that have successfully
restored viable populations across varying states
of degradation and landscapes. This will allow
identification of good practices and potential
problems associated with genetic quality under
various restoration approaches and contexts, and
facilitate formulation of solutions to overcome
them, thus contributing to more successful (eco-
logical and socioeconomic) restoration of forest-
ed ecosystems.
17.1. Recommendations arising
from the thematic study
17.1.1. Recommendations for
research
• Evaluate the impact of different restoration
methods on the genetic diversity of restored
tree populations.
• Expand knowledge on native species,
particularly with respect to their ecological
and livelihood importance, propagation
methods and genetic variability, and identify
ways to overcome constraints that limit their
use in restoration.
• Develop, make available and support the
adoption of decision-support tools for:
(i) collecting and propagating germplasm
in a way that ensures a broad genetic base
of restored tree populations; (ii) matching
of species and provenances to restoration
sites based on (current and future)
site conditions, predicted or known
patterns of variation in adaptive traits,
and availability of seed sources; and (iii)
landscape-level planning in restoration
projects.
• Develop protocols and practical indicators to
monitor and evaluate the genetic diversity
of tree populations in restoration efforts as
an indicator of the viability and resilience of
ecosystems.
• Intensify research on the ecology of
mycorrhizal and bacterial symbiotic systems,
focusing on the most commonly used tree
species and their symbiotic partners to
increase the resilience of plant associations
in restoration against biotic and abiotic
stresses.
17.1.2. Recommendations for
restoration practice
• Give priority to the use of native tree species
in restoration projects.
• Strive to use propagation material that
is well matched to the environmental

281
GENET IC CONS IDERATI ONS IN ECO SYSTEM RE STORATI ON USING N ATIVE TRE E SPECIE S
conditions of the restoration site and
represents a broad genetic base.
• Given the uncertainty of predictions of
future climate, aim to promote resilience
by maximizing species and genetic diversity
from sources that are similar to the site
conditions, encouraging gene flow and
generational turnover, and facilitating
species migration to allow natural selection
to take place.
• Plan for the sourcing of propagation
material of desired species and associated
information well before the intended
planting or seeding time to ensure
that optimal material for the site and
restoration objectives can be identified
and produced.
• Consistently plan restoration efforts in the
landscape context and seek to integrate
them into the surrounding landscape matrix.
17.1.3. Recommendations for policy
• Create an enabling national policy
environment that fosters long-term,
ecologically based forest management
that explicitly favours the use of native
species in ecosystem restoration and genetic
conservation and provides adequate
financial support.
• Put in place supportive regulatory
frameworks that guide the production and
supply of propagation material of native
tree species and the use of adequately
diverse material of appropriate origin in
restoration efforts.
• Broaden education and training curricula to
promote understanding of the importance
of using native species and genetically
diverse and appropriate propagation
material, as well as appropriate approaches,
in restoration projects.

THE STATE
OF THE WORLD’S
FOREST GENETIC RESOURCES
THEMATIC STUDY
GENETIC
CONSIDERATIONS
IN ECOSYSTEM RESTORATION
USING NATIVE TREE SPECIES
THE STATE OF THE WORLD’S FOREST GENETIC RESOURCES – THEMATIC STUDY
GENETIC CONSIDERATIONS IN ECOSYSTEM RESTORATION USING NATIVE TREE SPECIES
There is renewed interest in the use of native tree species in ecosystem restoration
for their biodiversity benefits. Growing native tree species in production systems
(e.g. plantation forests and subsistence agriculture) can also ensure landscape
functionality and support for human livelihoods.
Achieving full benefits, however, requires consideration of genetic aspects that are
often neglected, such as suitability of germplasm to the site, quality and quantity of
the genetic pool used and regeneration potential. Understanding the extent and
nature of gene flow across fragmented agro-ecosystems is also crucial to successful
ecosystem restoration.
This study, prepared within the ambit of
The State of the World’s Forest Genetic Resources
,
reviews the role of genetic considerations in a wide range of ecosystem restoration
activities involving trees. It evaluates how different approaches take, or could take,
genetic aspects into account, thereby leading to the identification and selection of
the most appropriate methods.
The publication includes a review and syntheses of experience and results; an
analysis of successes and failures in various systems; and definitions of best
practices including genetic aspects. It also identifies knowledge gaps and needs for
further research and development efforts. Its findings, drawn from a range of
approaches, help to clarify the role of genetic diversity and will contribute to future
developments.
I3938E/1/07.14
ISBN 978-92-5-108469-4
9789251 0 8 4694