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FOREST INVASIONS
Ecology of invasive forest pathogens
Luisa Ghelardini .Nicola Luchi .Francesco Pecori .Alessia L. Pepori .
Roberto Danti .Gianni Della Rocca .Paolo Capretti .Panaghiotis Tsopelas .
Alberto Santini
Received: 22 October 2016 / Accepted: 18 June 2017
ÓSpringer International Publishing AG 2017
Abstract Invasive forest pathogens are a major
threat to forests worldwide, causing increasing dam-
age. The knowledge of both the specific traits under-
lying the capacity of a pathogen to become invasive,
and the attributes predisposing an environment to
invasion are to be thoroughly understood in order to
deal with forest invasions. This paper summarizes the
historical knowledge on this subject. Many aspects of
the ecological processes underlying alien forest
pathogens invasions are still unknown, which raises
several scientific issues that need further study. The
introduction of invasive forest pathogens to areas
where naı
¨ve hosts are found, is mainly due to global
plant trade. Rapid transportation and reduced delivery
times increase the chances of survival of pathogen
propagules and of their successful establishment in
new environments. In forest pathogens, the reproduc-
tion mode seems not to be a crucial determinant of
invasiveness, as highly destructive pathogens have a
variety of reproductive strategies. The most important
drivers of forest pathogen invasions appear to be
(a) great adaptability to new environmental condi-
tions; (b) efficient dispersal over long and short
distances, possibly assisted by the capacity to form
novel associations with endemic and/or alien insect
vectors; (c) the ability to exchange genetic material or
hybridize with resident or alien species. Moreover,
these features interact with some key traits of the
invaded environment, e.g. environmental variability
and biodiversity richness. Host resistance and natural
enemies may occur as a result of rapid selection/
adaptation after the epidemic phase of invasion.
Keywords Disease spread Economic impact
Invasibility Invasion pathways Invasiveness Novel
insect-fungus associations Pathogen hybridization
Introduction
Currently, emerging infective diseases of plants (plant
EIDs) are one among the greatest threats to
Guest Editors: Andrew Liebhold, Eckehard Brockerhoff and
Martin Nun
˜ez / Special issue on Biological Invasions in Forests
prepared by a task force of the International Union of Forest
Research Organizations (IUFRO).
L. Ghelardini N. Luchi F. Pecori
A. L. Pepori R. Danti G. Della Rocca
P. Capretti A. Santini (&)
Istituto per la Protezione Sostenibile delle Piante IPSP,
Consiglio Nazionale delle Ricerche CNR, Via Madonna
del Piano 10, 50019 Sesto Fiorentino, Italy
e-mail: alberto.santini@cnr.it
L. Ghelardini P. Capretti
Dipartimento di Scienze delle Produzioni Agroalimentari
e dell’Ambiente DiSPAA, Universita
`di Firenze, Piazzale
delle Cascine 18, 50144 Florence, Italy
P. Tsopelas
Hellenic Agricultural Organization ‘Demeter’ Institute of
Mediterranean Forest Ecosystems, Terma Alkmanos,
Ilisia, 11528 Athens, Greece
123
Biol Invasions
DOI 10.1007/s10530-017-1487-0
agriculture, forestry and biodiversity conservation
(Anderson et al. 2004; Misra and Chaturvedi 2015).
In the past century, plant EIDs mostly resulted from
the accidental introduction of invasive alien pathogens
to new geographic areas by global trade and transport
(Bandyopadhyay and Frederiksen 1999; Brasier 2008;
Liebhold et al. 2012; Roy et al. 2014; Santini et al.
2013; Xu et al. 2006). Well known cases of introduc-
tion of invasive forest pathogens are the entrance of
Cronartium ribicola, the agent of white pine blister
rust, from Europe to North America along with plant
material; the import of Cryphonectria parasitica, the
agent of chestnut blight, from Asia to North America
through plant propagation material; and the arrival of
the subspecies americana of the Dutch elm disease
fungus Ophiostoma novo-ulmi, from North America to
Europe with rock elm logs (Brasier and Gibbs 1973).
The human-mediated extension of the distribution
range of microbes also enabled the emergence of new
pathogens through hybridization or horizontal gene
transfer (HGT) between previously isolated species
(Brasier 2000; Gluck-Thaler et al. 2015). In some
cases, the epidemic emergence of native or introduced
pathogens was caused by the establishment of novel
associations with introduced or native arthropod
vectors (Wingfield et al. 2016).
Quantifying the damage by invasive pathogens is
complex and only a few studies have calculated the
cost of multiple alien diseases either at regional or
global scale. Pimentel et al. (2001) estimated that
invasive diseases and pathogens caused a worldwide
loss of US$ 426 billion in 1998. In the US, the losses
and control costs due to plant diseases introduced from
abroad annually reach about US$ 21 billion (Brownlie
et al. 2006).
With regard to forest trees, the few available
quantitative estimates indicate large economic impact
by alien pathogens (Sache et al. 2011; Lovett et al.
2016). Pimentel et al. (2005) reported that approxi-
mately US$ 2.1 billion in forest products are lost each
year to alien forest pathogens in the US. In Canada,
past introductions of harmful invasive plant pests on
agricultural crops and forestry cost US$5.7 billion per
year (Environment Canada 2004).
Apart from production loss, the full economic costs
of invasions include negative side effects on trade of
forest products and plants, control expenses due to
inspections, monitoring, prevention and response, and
ecological and environmental impacts on ecosystems
(Morse 2005). Recently, Bradshaw et al. (2016)
performed a critical synthesis of the available infor-
mation to estimate the global goods and services costs
of invasive insects. The ten most harmful insects to
agriculture and forestry cost US$35 billion annually
(2014 values), a massive yet largely underestimated
value due to the shortage of quality data at the regional
or local scale, as highlighted by Bradshaw et al.
(2016). Invasive pathogens have ecological and envi-
ronmental effects at all organization levels, from
genes to the whole ecosystem (Morse 2005; Loo
2009). They affect the gene level by, for instance,
hybridizing with native microorganisms, a phe-
nomenon that drives the fast emergence of new
pathogens able to infect new host species (Brasier
et al. 1999; Stukenbrock 2016). Epidemic outbreaks of
alien disease agents may modify diversity, richness,
composition and abundance of host and non-host
species, and in turn affect ecosystem processes and
biogeochemical cycles (Morse 2005; Lovett et al.
2016).
To identify both the specific traits that underlie
invasiveness, i.e. the capacity of any introduced
organism to become a serious pest, and the attributes
predisposing an ecosystem to invasion, i.e. determin-
ing its invasibility, is crucial for understanding the
dynamics of invasion by forest pathogens (Santini
et al. 2013; Desprez-Loustau et al. 2016). Recently,
Garcia-Guzman and Heil (2014) reviewed the studies
that have identified life history traits of plants and of
their pathogens predictive for emergence of new
diseases in the tropics. They concluded that the only
widely accepted predictive quality in plant-pathogen
interactions is ‘host conservatism’, i.e. the probability
that a pathogen with a compatible interaction with a
given host will be able to infect another species
decreases with the phylogenetic distance among the
hosts (Gilbert and Webb 2007; Schulze-Lefert and
Panstruga 2011). These authors also showed that plant
life-history, in addition to determining genetic struc-
ture and evolution of parasite populations (Barrett
et al. 2008; Giraud et al. 2010), influences disease
incidence and the relative frequency of plant infection
by fungal pathogens differing in life-history traits. As
for other invasive organisms, fungal pathogens may
become successful alien invaders thanks to pre-
introduction traits conferring them the ability for
easy/fast adaptation, such as relatively high pheno-
typic plasticity, i.e. the ability of individual genotypes
L. Ghelardini et al.
123
to produce different phenotypes when exposed to
different environmental conditions (Pigliucci et al.
2006), and a broad ecological niche; or to post-
introduction changes in key life-history traits, which
may favour host jump (Giraud et al. 2010; Gladieux
et al. 2011,2015). Changes may consist in increased
aggressiveness, increased numbers of virulence traits,
and the loss of sexual reproduction, which most
probably results from adaptation than from genetic
drift (Ali et al. 2010). Finally, the ability to adapt to
novel abiotic stress factors is a crucial determinant of
invasion success (Gladieux et al. 2015).
A comparative analysis of the differences in many
phenotypic traits between source and invasive popu-
lations of Seiridium cardinale, the fungal agent of
cypress canker, showed that phenotypic plasticity
increases at first during invasion, possibly helping
survival in novel habitats, and it decreases afterwards
(Garbelotto et al. 2015). Moreover, in the same study,
selection proved to be stronger on traits related to
dispersal than on traits related to virulence.
The concepts of species invasiveness and commu-
nity invasibility, and their relations have been exten-
sively studied in the ecology of plant invasions
(Richardson and Pysek 2006; Hui et al. 2016).
In forest trees, the primary cause of EIDs is the
invasion by alien fungal and fungal-like pathogens
(Anderson et al. 2004; Ghelardini et al. 2016). Fungal-
like organisms (FLOs) are species that were recently
moved from Fungi to the kingdom Chromista, Phy-
tophthora species among them. Many pathogenic
fungi and FLOs share typical traits (for instance, high
virulence; high environmental persistence in the
absence of the host due to sapronotic potential or
durable inocula; large host range (generalist patho-
gens); highly dynamic genomes (high rate of macro-
mutations); broad environmental envelope, i.e. the set
of environments within which the species can persist,
that may represent key features in invasiveness
(Desprez-Loustau et al. 2007; Fisher et al. 2012).
Traits linked to invasiveness and invasibility in fungal
pathogens of temperate forests were investigated by
several studies (Desprez-Loustau et al. 2010; Vacher
et al. 2010; Philibert et al. 2011; Santini et al. 2013).
However, many aspects of the ecological processes
leading to invasion of forest ecosystems by introduced
pathogenic fungi remain unknown. This paper sum-
marizes the historical knowledge on this subject,
examining each stage of the invasion process:
transport, arrival, establishment, and spread as defined
in Roy et al. (2016), but many issues remain open and
need to be further investigated.
Transport and arrival
Critical events in human history and the technology
progress applied to intercontinental transport had a
huge impact on the spread of invasive forest patho-
gens. Until the end of World War I, the main pathway
of emergence of new forest diseases in Europe was the
transport of commodities among European countries.
Afterwards, Europe became the first market of US
products, and North America, as a consequence, the
main source of new forest pathogens (Santini et al.
2013). Following the increase of commercial
exchanges with the US, temperate regions of North
America became the main source of alien pathogens
also for China (Xu et al. 2006).
The number of introductions of invasive forest
pathogens rapidly increased after World War II, when
standardized containerization revolutionized interna-
tional trade, dramatically reducing transport costs and
shortening shipping time. Since the mid 1950s, when
the first commercially successful container ships
carried a few dozen containers between US destina-
tions, the number of containers transported by a single
ship has increased up to more than 19,000 units
transported by the China Shipping Container Lines
Globe, the world’s largest container ship in 2014.
Rapid transportation and reduced delivery times
increase the survival of pathogen propagules, and
increase their chances of successful establishment in
the environment of arrival (Hulme 2009). Three
among the most infamous alien pathogens of Euro-
pean trees, namely S. cardinale,Ceratocystis platani,
and Heterobasidion irregulare, the agents of cypress
canker, plane stain canker and Heterobasidion root rot,
respectively, were introduced to Europe by the US
army during WWII (Fig. 1). The fall of the Iron
Curtain opened the doors to further globalization,
leaving a world without strict borders and few, if any,
areas sheltered from invasive species migrations.
From the 1990s onwards Asia became, mainly because
of cheaper labour costs, one of the world’s most
important producers of plants for planting and,
therefore, a new source of invasive fungal pathogens
of trees for Europe (Fig. 1).
Ecology of invasive forest pathogens
123
Establishment
Pathogen traits driving invasiveness
Life history traits
Reproduction in fungi is complex, and a species’ life
cycle may include asexual, sexual and/or parasexual
stages (in the latter two cases novel genetic variation is
produced by recombination). Fungi exhibit three
different sexual compatibility systems: (1) hermaph-
roditic, i.e. a mycelium produces both female and male
organs (self-fertile or not); (2) dioecious, i.e. a certain
mycelium may only bear female or male organs
(obligatory outcrossing); (3) sexually undifferenti-
ated, i.e. sexual structures may act as male or female
(Nieuwenhuis and Aanen 2012).
The reproductive strategy does not appear to be
among the main determinants of invasiveness in forest
pathogens, although it may be important for invasion
management. In fact, the most destructive epidemics
of forest trees were caused by invasive fungi with very
different reproduction modes. For instance, S. cardi-
nale,Diplodia sapinea and C. platani (for more details
see ‘‘Box 1’’, ‘‘Box 2’’ and ‘‘Box 3’’ sections,
respectively) spread clonally in the invaded range;
the first two species reproduce by agamic and para-
sexual systems, while the third species reproduces
agamically and by sexual mechanisms, being her-
maphroditic and self-fertile. Cronartium ribicola,a
rust fungus colonizing white pines, has a complex life
cycle during which it produces 5 different spore types,
requiring an alternative host (Ribes spp.) to produce
basidiospores, i.e. sexual spores. The fungus is native
to central Siberia (Leppik 1970) and became invasive
in the mid 1800s on American white pine species in
Europe, being accidentally introduced to North Amer-
ica and Asia after a few decades through infected
nursery stock (Hummer 2000). In every part of the
invaded range the fungus spread very rapidly, jumping
to new hosts (Pedicularis and Castilleja spp.)
(McDonald et al. 2006).
Ophiostoma novo-ulmi, the causal agent of Dutch
elm disease and one of the most destructive forest
pathogens ever known, reproduces both clonally and
sexually. Once introduced into a new area, this fungus
generally spreads at the epidemic front, with a single
mating type exhibiting only one or a few vegetative
compatibility groups. A few years after establishment,
these genetically uniform invasive populations sud-
denly increase in genetic variability due to the
appearance of the other mating type (Brasier 1988;
Brasier et al. 2004).
Host specificity
In ecology, species confined to particular habitats
because of their narrow environmental tolerances are
defined specialist species, while species with a broader
environmental tolerance enabling them to live in many
Fig. 1 Temporal changes
in the origin of forest
pathogens established in
Europe (updated from
Santini et al. 2013)
L. Ghelardini et al.
123
and diverse habitats are defined as generalist species.
Generalist parasites, as opposed to specialist parasites,
are able to infect a wide range of hosts with varying
severities (Holmes 1979). Observed host ranges of
parasites often include only part of the potential hosts
because of limitations to dispersal (Poulin 2011). In
Europe, generalist invasive pathogens of trees tended
to spread over a larger area than specialist pathogens
(Santini et al. 2013). The ability of generalist species
to become successful invaders is probably due to their
capacity to infect novel hosts, persist in a wide range
of environmental conditions, and to spread over longer
distances (Brown and Hovmøller 2002; Evangelista
et al. 2008). The unspecialized pathogen Phytoph-
thora ramorum probably became a successful invader
in the UK, causing huge damages to the Japanese larch
Larix kaempferi (Lamb.) Carr., by virtue of the
pathogen’s capacity to quickly adapt to a naı
¨ve host,
grown in uniform plantations and highly susceptible
through lack of coevolution (Brasier and Webber
2010). In generalist pathogens, the ability to live on
many host species probably helps the maintenance of
high within-population genetic variation, which
increases their evolutionary potential (Kassen 2002).
Nevertheless, among forest pathogens there are also
striking examples of successful invaders with high
host specificity, as Hymenoscyphus fraxineus, the
agent of European ash dieback (Kra
¨utler and Kirisits
2012), and obligate pathogens, for instance rusts
(Prospero and Cleary 2017). However, highly special-
ized pathogens frequently experienced local extinc-
tion, especially when the host population’s range was
discontinuous (Barrett et al. 2008). Other specialized
pathogens, as O. ulmi s.l. and C. parasitica, which
have nearly destroyed their hosts in two continents,
underwent less extreme boom and bust population
cycles possibly because they share with other fungal
pathogens the capacity to survive as a saprophyte on
plant debris between successive epidemic outbreaks
(Webber et al. 1987; Baird 1991; Prospero et al. 2006;
Rigling and Prospero 2017).
Hybridization
The 1990s and early 2000s were characterized by the
discovery of an exceptional number of hybrid forest
pathogens (reviewed in Brasier 2000 and Ghelardini
et al. 2016). Again, the cause of this phenomenon is
probably the accidental transfer of micro-organisms
through global transport and trade, which increased
the chances of contact between species that were
previously isolated by distance. Once established in a
new area, invasive pathogens come into contact with
native and possibly related pathogens, some of which
may have similar hosts or vectors. Thus it is more
likely that hybridization will occur between immigrant
species and an endemic species belonging to the same
genus because allopatric species are less likely to have
evolved strong barriers to hybridization (Brasier
2000). Other factors influencing the probability of
hybridization are the likelihood of niche contact
between the two species, and the fitness of any
recombinant hybrids relative to that of the parent
species (Brasier 2000). However, there is sound
evidence that hybrid plant pathogens may display
increased virulence or enlarged host range when
compared with their parental lineages, which provides
them with a great potential for invasion and epidemic
outbreak (Depotter et al. 2016; Stukenbrock 2016).
With regard to forest pathogens, Paoletti et al. (2006)
demonstrated selective acquisition of mating type
(Mat) and vegetative incompatibility (vic) genes by O.
novo-ulmi from O. ulmi. Brasier and Kirk (2010)
provided evidence for emergence of a hybrid swarm of
the O. novo ulmi subspecies across Europe with an
increased level of pathogenic fitness. Dhillon et al.
(2015) provided evidence that Mycosphaerella popu-
lorum, the Septoria canker of poplars, acquired the
capacity to infect, colonize, and cause mortality on
poplar woody stems through horizontal transfer of the
necessary gene battery from ascomycete fungi asso-
ciated with wood decay and from prokaryotes. On
hybridization see also Wingfield et al. (2017a).
Environmental and host species features driving
invasibility
Climatic variability, biodiversity and human activity
The pattern of invasion by forest pathogens depends
on the environmental variability and biodiversity of
the invaded region, in addition to the magnitude of
human activities and commercial exchanges (Santini
et al. 2013). Consistently with the ‘environmental
heterogeneity hypothesis of invasions’ (Melbourne
et al. 2007), a high environmental diversity allows the
establishment of many forest pathogens with different
ecological niches. Regions with high diversity may
Ecology of invasive forest pathogens
123
display environmental conditions suitable for numer-
ous alien species and function as ‘invasion hotspots’
(O’Donnell et al. 2012). For example, Italy and
France, having high environmental and biological
diversity among European countries, suffered histor-
ically the highest number of invasions by forest
pathogens in Europe (Santini et al. 2013; Roy et al.
2014). However, part of the observed differences in
the frequency of introductions/invasions is probably
due to non-uniform biosecurity regulation or ineffec-
tive application of common rules (Eschen et al. 2015;
Klapwijk et al. 2016). In addition, human activities,
which can be represented by various economical
indices, as the volume of imports (Desprez-Loustau
et al. 2010) or the gross domestic product (GDP)
(Santini et al. 2013; Roy et al. 2014), or by human
presence as measured by total population (Desprez-
Loustau et al. 2010) or population density (Santini
et al. 2013), influence the frequency and pattern of
invasion. For instance, the high incidence of forest
invasions in Italy probably depends on the synergistic
effect of biogeographical diversity, concentrated
human activities, and intense prolonged commerce
(Santini et al. 2013). This is consistent with the
‘species-energy theory’, which posits that the stability
of a host-pathogen system and the ecosystem’s
pathogen-carrying capacity are higher in regions with
a higher biomass productivity (Wright 1983).
In the near future, the countries at the highest risk of
invasion by forest pathogens are probably those that
have been commercially isolated in the recent past,
and those with a low Human Development Index
(HDI), previously limited and currently expanding
international trade/GDP, and fragile economy with no
resources to invest in prevention/control of alien
invasions (Early et al. 2016; Roy et al. 2014).
Biodiversity hotspots in their territories will probably
be vulnerable to relatively low-impact invasive
(pathogen) species (Early et al. 2016; Li et al. 2016).
Europe is the continent with the highest number of
invasive forest pathogen species, i.e. there are about
four and five times as many reports as in China and
North America, respectively (Fig. 2). Since plant trade
is the major source of plant pathogens’ introduction, a
possible explanation is that Europe has one of the most
open phytosanitary systems among highly developed
countries (Brasier 2008; Eschen et al. 2015), i.e. any
plant that is not specifically regulated can be imported.
Brasier (2008) highlights in particular (1) that many of
the introduced threat organisms were previously
unknown to science; (2) the European reliance on
named organisms; (3) risk of rapid evolution via
hybridization; (4) the danger of the ‘open border’ basis
of plant biosecurity protocol within Europe. Plant
trade is currently a huge business in Europe and North
America, therefore the risk of further invasion remains
very high even in regions that suffered many invasions
in the past.
Climate change may alter ecosystem functioning
and change species’ richness and abundance, enhance
the fitness of pathogens and drive their range-expan-
sion, weaken host plants, i.e. predispose them to
infection, eventually promoting the establishment of
new invasive alien plant, animal and pathogen species
(Scherm and Coakley 2003; Theoharides and Dukes
2007; Eastburn et al. 2011). According to Early et al.
(2016) the continuous climate change disturbance
expected for the near future will increase the invasion
threat in Eastern North America, Northern Europe,
Central and Southern Asia, and Northern Australia.
Spatial structure of host populations
Besides the host’s life-style and mode of reproduction,
abundance and spatial distribution of the host’s
populations also significantly influence a pathogen’s
population dynamics, invasion success, and disease
impact (Real and Biek 2007). Therefore, the capacity
of a tree species to form pure forests or on the contrary,
its tendency to grow scattered in mixed forests,
together with the host’s tendency to reproduce sexu-
ally or clonally, might favor a pathogen or another
depending on their requirements and features. As an
extreme example, genetically uniform populations on
large areas, such as planted forests, represent a
Fig. 2 Number of invasive forest pathogens (data modified and
updated from Xu et al. 2012; Liebhold et al. 2012; Santini et al.
2013)
L. Ghelardini et al.
123
uniform and continuous carbon source, and an empty
niche, readily available for pathogens, whether they be
introduced species, against which there is usually a
lack of natural resistance, or native species, which
may, in such conditions, be subjected to fast adaptive
selection. In this case, a pathogen’s population may
undergo massive size growth, and only afterwards,
when the population size is stable again, new genetic
variation usually appears (Barrett et al. 2008). In wild
tree populations of smaller size, patchily distributed or
isolated, the populations of invasive pathogens remain
usually smaller, have limited genetic variability, and
are commonly subjected to near-extinction events, in
which drift and gene flow play a significant role
(Barrett et al. 2008).
Host resistance
The interactions between host plants and pathogens
enhanced ecosystem diversity during evolution
(Thompson 1998). Every ecosystem is shaped by
different interactions between plants and herbivores
including pests and pathogens, which result in the
selection of useful genes both in prey and predator,
enabling a long-standing association between the
involved species. Flor’s gene-for-gene hypothesis
(Flor 1971) and the ‘‘zig-zag model’’ by Jones and
Dangl (2006) best explain parasite-host interactions in
obligate pathogens. They proposed that for each gene
that conditions reaction in the host (i.e., a resistance-
or R-gene) there is a corresponding gene in the parasite
(i.e., an avirulence- or Avr-gene) that conditions
pathogenicity. According to this model plants detect
microbial/pathogen–associated molecular patterns
(MAMPs, PAMPs) via receptors triggering an immu-
nity response. Pathogens may evolve new effectors
that enable them to overcome plant immunity. Plants
may in turn evolve new receptors able to recognize
those new effectors, triggering immunity again.
In this way, host plants and their pathogens
establish a dynamic equilibrium in a continuous
process of co-evolution. The assemblages of plants
and pathogens have mostly evolved in a condition of
genetic isolation because of geographical barriers.
When these barriers are broken, most frequently
because of human activities, and exotic pathogens
are introduced into new environments, they may find
naı
¨ve hosts lacking specific resistance genes targeting
them, and a favorable environment, i.e. conditions that
increase pathogen aggressiveness and possibly result
in disease outbreak (Brasier 2001)
Host phenology and disease avoidance
Avoidance can be defined as a mean of disease
resistance, encompassing mechanisms under the
genetic control of the host, taking place whenever
susceptible plants do not become infected because the
factors necessary for disease development (susceptible
host, virulent pathogen, and favorable environment)
do not coincide and interact at the proper time or for a
sufficient period of time (Ghelardini and Santini
2009). The variation in the timing of the host’s life
stages and phenological phases characterized by
different degrees of susceptibility, i.e. the transitory
presence of plant tissues or organs susceptible to
infection, change host availability, making it possible
for a tree to resist infection, therefore altering disease
epidemiology (Dantec et al. 2015). For example, in the
European field elm (Ulmus minor Mill.) plants that
open their vegetative buds earlier in spring are less
susceptible to Dutch elm disease than late flushing
plants (Santini et al. 2005). Early flushing may
function as a mechanism of disease avoidance inas-
much it causes a temporal mismatch between the time
of maximum susceptibility, which is linked to the
host’s phase of growth, and the appearance of the
Scolytus insects vectoring the infection. Rhythm of
leaf expansion, rate of height and diameter growth,
seasonal changes in wood anatomy, and seasonal
variation in secondary metabolism are likely involved
in the seasonal variation of susceptibility in elms and
in disease avoidance by early flushing genotypes
(Solla et al. 2005; Ghelardini and Santini 2009). The
variation of host susceptibility in the course of time
may however exert a strong selective pressure, which
might be able to promote fast adaptation in pathogens
or vectors.
Enemy release
The traditional explanation for the success of alien
species as invaders in new regions is that they do not
have enemies in the non-native range (Wolfe 2002)
and/or they attack non-coevolved hosts, which lack
specific defenses against them (Morrison and Hay
2011). However, at least in the case of invasive plants,
the enemy release hypothesis has been questioned
Ecology of invasive forest pathogens
123
(Heard et al. 2006; Liu and Stiling 2006; Parker and
Gilbert 2007). As regards invasive forest pathogens,
the effect of a temporary lapse of enemies in the new
environment has never been tested. However, there is
circumstantial evidence for either delayed adaptation
or acquisition of natural enemies in the invaded range
at least for two alien invasive pathogens in Europe, i.e.
the appearance of virus mediated hypovirulence in
Cryphonectria parasitica (Grente 1965) and of dele-
terious d-factor viruses in Ophiostoma ulmi s.l.
(Brasier 1983; Buck et al. 2002).
Novel associations
Once established in a new area, invasive alien pests or
pathogens may replace endemic species in local
insect-fungus associations. The establishment of a
novel association may increase dispersal and/or
transmission efficiency of the involved alien pests or
diseases resulting in epidemic outbreaks (Wingfield
et al. 2016).
An illustrative case is represented by the replace-
ment of the non-aggressive fungus Pestalotiopsis
funerea (Desm.) Stey, which is endemic in Europe,
by the aggressive alien invasive pathogen Seiridium
cardinale (Wag.) Sutton and Gibson, the causal agent
of cypress canker, in the ancient symbiotic association
with the cypress seed bug Orsillus maculatus in the
Mediterranean Basin. P. funerea and O. maculatus had
co-evolved with the Italian cypress (Cupressus sem-
pervirens L.), and only caused negligible damage in
the native range (Santini and Di Lonardo 2000). Once
introduced into Europe and North Africa, S. cardinale
was rapidly spread by the bug with disastrous conse-
quences for the survival of C. sempervirens (Battisti
et al. 1999).
Another illustrative example is the association
between the introduced O. ulmi and O. novo-ulmi
(collectively Ophiostoma ulmi s.l.) and endemic elm
bark beetles in the subfamily Scolytinae in the invaded
range. The spread of the two fungi is mainly operated
by these beetles (Fransen and Buisman 1935; Webber
and Brasier 1984), and disease transmission to
suitable hosts is effective since the life cycles of the
host-trees, the pathogen(s) and the beetles are syn-
chronized (Webber 2004; Ghelardini 2007; Ghelardini
and Santini 2009). Before DED, elm bark beetles in
Europe probably had a mutualistic ectosymbiosis with
the indigenous saprophytic fungus O. quercus
(previously known as the hardwood biological species
group, or OPH group, of O. piceae) (Brasier 1990). In
the past century, the introduction of O. ulmi s.l., a
species with similar niche requirements to O. quercus
but pathogenic on Ulmus spp., has resulted in the
replacement of this endemic congeneric species on
elm. The new bark beetle/fungus association resulted
in a highly effective transmission pathway with
devastating consequences for the survival of elms
(Brasier 1982; Santini and Faccoli 2015).
The Western conifer seed bug Leptoglossus occi-
dentalis Heidemann (Heteroptera,Coreidae), which
seriously impairs seed production of conifer species in
Europe (Roversi et al. 2011; Tamburini et al. 2012;
Lesieur et al. 2014), was introduced in the late 1990s
and in just a decade invaded large parts of the
continent (Taylor et al. 2001; Fent and Kment 2011).
In Italy, L. occidentalis is the vector of Diplodia
sapinea (Fr.) Fuckel Nassau (Luchi et al. 2012), a
cosmopolitan fungal pathogen of conifers, probably
native to Southern Europe. As a result of this new
association, D. sapinea is today dispersed much faster
and reaches a higher number of susceptible trees of the
genera Cedrus,Juniperus,Picea and Pseudotsuga,
which rarely were attacked by the fungus in natural
conditions before (Stanosz et al. 1999). In addition, in
the last 25 years, milder winters have allowed D.
sapinea to expand its distribution range towards
Northern Europe, threatening an even greater number
of conifer species (Hanso and Drenkhan 2009; Oliva
et al. 2013).
Several new interesting associations between
ophiostomatoid fungi, insects and tree hosts are also
described by Wingfield et al. (2017b in this issue) and
the possible reasons of a recent increase in emergence
of the disease due to these associations were also
discussed.
Spread
Domestic invasion pathways
Pathogens may be transported by human travel and
trade as hitchhikers or contaminants on different
goods or in soil. The intercontinental trade of live
plants for planting is the major pathway for the spread
of non-native plant pathogens globally. The commerce
of live plants also plays a major role in the dispersal of
L. Ghelardini et al.
123
invasive forest pathogens at the regional and local
scale (Jones and Baker 2007), confirming the impor-
tance of human-mediated jump-dispersal in determin-
ing invasion dynamics subsequent to introduction or
establishment (Suarez et al. 2001).
In addition, pathogens may be dispersed through
wind-, water-, and animal-borne propagules. Dispersal
at the continental scale by means of wind-borne
structures has long been documented in species
infecting crop plants (Brown and Hovmøller 2002)
as in forest tree pathogens (Close et al. 1978). When a
pathogen is introduced by humans to non-native areas
populated by naı
¨ve hosts, a largely wind-borne
epidemic is possible, as it happened with chestnut
blight in North America (also assisted by occasional
dispersal by insects and animals) or, more recently, in
the case of ash dieback in Europe (Anagnostakis 2012;
Gross et al. 2014). A recent analysis of a newly
compiled global database of invasive tree pathogens
suggests that wind-dispersed fungi may be overrepre-
sented among invasive tree pathogens worldwide
(Ghelardini et al. unpublished).
Medium- to short-distance dispersal is likely to be
crucial for the successful establishment of a pathogen
in the invasion range, greatly affecting an epidemic’s
development, and disease dynamics and persistence.
At the local scale, rain splash-dispersed pathogens are
those with the shortest spreading distance, while air-
borne pathogens have the widest spreading area. In
between these two extremes there are vector-borne
pathogens. Very efficient incidental vectors of inva-
sive forest pathogens can be animals, and humans
among them, with or without fitness benefit, as
respectively symbiotic insects, e.g. beetles and
ambrosia fungi, and squirrels (Panconesi 1999; Lieu-
tier et al. 2004). Ceratocystis platani (see ‘‘Box 3’’
section) and Cryphonectria parasitica provide good
examples of lethal wound pathogens possessing
multiple short-distance dispersal means, i.e. wind,
vertebrate and invertebrate animals, and water. How-
ever in some pathogens, as in C. platani, the most
important dispersal route to short- and medium-
distance is represented by pruning tools and terracing
machinery, just showing how crucial human activities
can be in spreading alien pests. Root anastomosis is an
important mean of spread invasive pathogens attack-
ing trees grown in lines as in city avenues, or in
monospecific plantations, as shown by the examples of
O. ulmi s.l. (Gibbs 1978), Heterobasidion irregulare
(Garbelotto and Gonthier 2013) and C. platani
(Tsopelas et al. 2017).
Impact
The economic and environmental impacts of alien
forest pathogens on natural ecosystems have only
recently, and in a limited number of studies, been
assessed in-depth (Lovett et al. 2006; Pimentel et al.
2000). The removal costs of infected trees were used
as an estimate of the economic damage caused by
invasive forest pathogens (Pimentel et al. 2000; Haight
et al. 2011). For instance, the removal cost was
estimated at about US$100 million per year for the
elms killed by DED in North America (Campbell and
Schlarbaum 1994) and at roughly US$250 per tree
killed by oak wilt (Haight et al. 2011). Pimentel et al.
(2000) estimated the loss of forest products caused by
plant pathogens in about 9% of the total production, or
US$7 billion, per year.
Forests provide qualitative and long-term services
which are difficult to monetize. The introduction and
spread of invasive pathogens may change tree species
abundance and diversity, altering the structure of
forest ecosystems, their productivity, nutrient uptake,
carbon and nitrogen cycles, and soil organic uptake
and turnover (Lovett et al. 2006,2010). The epidemic
of chestnut blight due to the introduction of C.
parasitica caused dramatic changes in the forests of
eastern North America. In 1934 Castanea dentata
(Marsh.) Borkh. was present in 98% of the forests in
the Coweeta Basin, in the Southern Appalachian
Mountains of North Carolina, contributing 22% of the
total density and 36% of the total basal area (Elliott
and Swank 2008). As the epidemic progressed, C.
dentata was replaced by different species along
environmental gradients, which produced a general
increase in species diversity (Elliott and Swank 2008).
Davis et al. (2014) showed that plant richness, cover at
all strata and flowering are reduced at sites infested by
Phytophthora cinnamomi in Western Australia, caus-
ing a drastic change in the bird community composi-
tion. In Oregon and California, Phytophthora
ramorum modified the structure of forests so that the
number of ecological niches suitable for two of the
vertebrates’ host of the Lyme disease pathogen and its
vector, the tick Ixodes pacificus, resulted increased
(Swei et al. 2011). The disturbance due to the invasion
Ecology of invasive forest pathogens
123
by P. ramorum indirectly increased the survival and
abundance of ticks at the nymph stage, which is most
likely to bite humans, consequently increasing the risk
of exposure to the Lyme disease.
The impact of S. cardinale on the Italian cypress,
the iconic landscape tree of Tuscany, was tentatively
assessed in the Project MED Operational Programme
‘‘CypFire’’ (see ‘‘Box 1’’ section). In order to evaluate
the loss of landscape/ornamental value caused by
cypress canker, about 600 tourists on holidays in
Tuscany were interviewed according to a previously
designed questionnaire. The questionnaire was based
on two assessment methods: (1) the ‘‘Landscape
components evaluation’’ (Tempesta 1997; Giau
1999), a non-monetary estimate of the importance of
cypress in the Tuscan landscape; and (2) the ‘‘Willing-
ness to pay’’, a contingent monetary method (Stellin
and Rosato 1998; Notaro et al. 2005) based on
people’s willingness to pay for a service, i.e. in this
case for preserving the Tuscan landscape from the
disappearance of cypresses.
The results showed that landscape is one of the
main attractions for tourists visiting Tuscany; that
cypress is considered a quintessential component of
the landscape of Tuscany; and that the disappearance
of cypress would render the Tuscan landscape far less
attractive and reduce the motivation for visiting the
region. The persons interviewed expressed their
willingness to pay a maximum of 11.2 Euros per
capita on the average for protecting cypress from
canker disease. Since each year between 460.000 and
550.000 tourists spend their holidays in farmhouses in
Tuscany, the estimated total willingness to pay
approaches 5–6 million Euros per year. This provides
a rough estimate of the yearly impact of cypress canker
on tourism, the economy and the landscape.
Conclusions
This paper summarizes some aspects of the ecological
process leading to invasion by forest pathogens.
Despite the worldwide importance and impact of
forest invasions, invasion ecology of forest pathogens
is a young research field, where much remains to be
understood (Lugo 2015). The introduction of alien
pathogens usually leads to novel host–pathogen or
pathogen–pathogen combinations with no previous
co-evolution history and complex outcome (Slippers
et al. 2005; Dunn and Hatcher 2015). Many introduced
pathogens responsible for severe disease epidemics of
trees were not considered serious pathogens or were
not even known before they attacked new hosts in the
invaded range (Brasier 2008). Microbial hitchhikers,
and latent or cryptic pathogens, living as endophytes,
epiphytes or weak pathogens on their native and co-
evolved host plants, once introduced into new regions
may undergo host shifts and move onto native plants,
changing their behaviour from mutualistic associates
to aggressive pathogens (Burgess et al. 2016; Roy
et al. 2016). The genomes of closely associated plant
microorganisms (including mutualists and pathogens),
recently referred to as the ‘second-genome’ (Grice and
Segre 2012; Turner et al. 2013; Berg et al. 2014), have
a crucial role in evolutionary mechanisms associated
to biological invasion (Blackburn et al. 2011; Zenni
et al. 2017). Because of post-invasion host jump and
limited knowledge of the many steps and different
components involved the invasion process, predicting
which species might become invasive forest pathogens
once introduced to new areas seems very difficult if
not unrealistic in many cases. The unpredictable nature
of biological invasion as well as the limitation posed to
current predictions by inadequate scientific knowl-
edge, has been recently analysed and discussed in
detail elsewhere (Roy et al. 2016). There is no doubt
that investment in research, possibly coordinated
within an interdisciplinary framework (Wingfield
et al. 2017a,b), and in-depth elaboration on the
ecology of forest invasions (Desprez-Loustau et al.
2016) is the only option to strengthen the holistic
interpretation of the phenomenon, develop well-
grounded predictive approaches and possibly conceive
the countermeasures for facing the constant attack
perpetrated by invasive forest pathogens.
Box 1
Seiridium cardinale
Seiridium canker, or bark canker of cypress, is a
pandemic disease due to three pathogenic fungi, i.e.
Seiridium cardinale (Wagener) Sutton and Gibson,
Seiridium cupressi (Guba) Boesew., and Seiridium
unicorne (Cooke and Ellis) Sutton, causing stem and
branch necrosis, dieback, and tree mortality in most
Cupressaceae species. Seiridium cardinale, the most
L. Ghelardini et al.
123
widespread and aggressive species, was first reported
in California in 1928, where in a few years it destroyed
the plantations of the inland districts (Wagener 1939).
During the following decades, the disease progres-
sively spread over the five continents and became
pandemic, reaching in succession New Zealand,
France, Chile, Italy, Argentina, Greece and other
Mediterranean countries, Central and Northern Eur-
ope, Canada, South Africa and Australia (Danti et al.
2013a). Seiridium cardinale infects various species of
Cupressus, Chamaecyparis, Cryptomeria, Juniperus,
Thuja and x Cupressocyparis leylandii (Graniti 1998).
Population genetic studies suggest the pathogen to be
native or long naturalized in California, which is the
source region of the Mediterranean population (Della
Rocca et al. 2011,2013). In the Mediterranean region,
the fungus reproduces only clonally, although varia-
tion is likely to be ensured by occurrence and
accumulation of mutations or mitotic recombinations,
as expected in a parasexual life cycle (Della Rocca
et al. 2011,2013). Movement of infected cypress
plants is thought to be responsible for the global spread
of the pathogen since identical genotypes of the fungus
were identified at large distances (Della Rocca et al.
2011,2013). In the Mediterranean basin, the extensive
presence of susceptible hosts, in combination with
climatic conditions conducive to the development of
the pathogen, has promoted its rapid spread. Cypress
canker reached an incidence as high as 70% in Central
Italy and Greece during the 1980s and severely
affected the local cypress and other introduced cypress
species (Panconesi 1990; Xenopoulos and Diamandis
1985). The epidemics killed millions of trees, spoiled
the landscape, and produced severe damage in the
woods, ornamental plantations, and nurseries (Graniti
1998; Panconesi 1990), with consequently high costs.
Cypress canker has gradually reached an endemic
phase in Central Italy, helped by extensive sanitation
measures and the breeding of resistant cypress culti-
vars (Danti et al. 2006,2013b).
Box 2
Diplodia sapinea
Diplodia sapinea (Sphaeropsis sapinea) is a haploid
fungus species with worldwide distribution as well as
an opportunistic plant pathogen of many conifers. The
pathogen is widespread in Europe, South Africa and
the United States, where it causes serious damage
especially to pine species both in natural forests and
commercial plantations (Stanosz et al. 1996; de Wet
et al. 2000; Luchi et al. 2014). The fungal spores are
dispersed by wind, water, animals, and human activ-
ities, and germinate on young elongating needles.
Penetration occurs through the stomata, or through
wounds caused by hail, insects, or pruning. Diplodia
sapinea is thought to reproduce only by asexual
mitospores, as no sexual stage has ever been observed
(Burgess et al. 2001).
The taxonomy of this species has been under
constant revision. The species included two morpho-
types, A and B, until the latter was elevated to species
level and named as Diplodia scrobiculata (J. de Wet,
B. Slippers and M.J. Wingfield) (de Wet et al. 2003).
These pathogens share the same habitat, but D.
sapinea is much more aggressive and produces more
pycnidia than D. scorbiculata (Luchi et al. 2007;
Blodgett and Bonello 2003). Therefore, D. sapinea has
spread faster and more efficiently in the northern
hemisphere than D. scrobiculata, which remained
confined to a smaller number of host species and
maintained a scattered distribution (Bihon et al. 2011;
Linaldeddu et al. 2010).
A key factor for understanding the invasion ecology
of these species is the latency period, during which the
pathogen lives asymptomatically in the host’s tissues
until a biotic or abiotic stress factors, for instance
water stress, modifies the equilibrium of the plant-
endophyte interaction, driving an epidemic outbreak
of the disease (Stanosz et al. 2001; Fabre et al. 2011).
In this respect, the current poleward spread of D.
sapinea is representative of the effects of climate
change-related stressors on disease emergence
(Adamson et al. 2015). Moreover, D. sapinea well
illustrates the case, probably quite common, in which
the presence of a possibly long latency period, when
apparently healthy plants carry the pathogen inside
their tissues, helps spreading a disease to new
geographic areas through commercial trade of asymp-
tomatic plants. D. sapinea and D. scrobiculata were
probably introduced to the southern hemisphere
through the import of asymptomatically infected pine
seedlings for plantations (Burgess et al. 2001). The
presence of a common allele between the populations
of D. scrobiculata from South Africa and California
confirms the introduction of the fungus from the US,
Ecology of invasive forest pathogens
123
probably through infected P. radiata seedlings (Bihon
et al. 2011).
Box 3
Ceratocystis platani
Ceratocystis platani (Walter) Engelbrecht et Harring-
ton causes mainly a vascular wilt disease of Platanus
spp. but it is also able to cause cankers on the bark,
which is the reason why the common and accepted
name of the disease, even if misleading, is canker stain
disease (CSD). The fungus, probably native to North
America (Walter et al. 1952; Engelbrecht et al. 2004),
was accidentally introduced during World War II into
Europe (Cristinzio et al. 1973; Panconesi 1999) where
it spread as a single clone (Santini and Capretti 2000;
Ocasio-Morales et al. 2007). The fungus is self-fertile
and reproduction is ensured by ascospores and by
three different forms of asexual conidia. C. platani
penetrates the host through wounds or other injuries
caused by biotic or abiotic agents in the branches, the
trunk, or the roots (Vigouroux and Stojadinovic 1990).
It is mainly spread by means of pruning tools,
terracing machinery and root anastomosis, and in the
second place by wind, animals, and water (Panconesi
1999; Luchi et al. 2013). All spore types of the fungus
produce infections on freshly wounded tissues. After
spore germination, the mycelium colonizes the
exposed tissues, advances into the xylem of the
underlying sapwood, and it develops both longitudi-
nally and tangentially (Tsopelas et al. 2017). CDS is a
devastating disease that may kill an adult tree in a
single growing season and a monumental tree in a just
few years (Tsopelas et al. 2017). Mortality rates due to
C. platani are extremely high. Many trees have been
killed throughout the invaded range in city parks and
avenues, including plants of great aesthetic value. The
spread of the disease out of the urban environment
threatens plane tree survival in natural forests.
Although often overlooked in forest pathology
reviews, in Southern Europe CSD has an impact
comparable to that of notorious tree diseases such as
DED, chestnut blight or ash dieback (Walter et al.
1952; Brasier and Kirk 2001; Kowalski 2006; Ocasio-
Morales et al. 2007; Anagnostakis 2012). The
pathogen, reported also in Switzerland and Spain
(Panconesi 1999; Vigouroux 2013; EPPO 2014),
caused widespread mortality of London plane (Pla-
tanus x acerifolia (Ait.) Willd) in Italy and France. But
the most dramatic impact of the disease concerns the
natural stands of oriental plane (Platanus orientalis L.)
in Greece (Ocasio-Morales et al. 2007; Tsopelas and
Soulioti 2011,2014) where CSD killed and is
currently killing each year thousands of trees that do
not exhibit any sign of resistance. More recently C.
platani was recorded for the first time in Albania
(Tsopelas et al. 2015). There are unconfirmed reports
for the presence of C. platani in Armenia (Simonian
and Mamikonyan 1982) and Iran (Salari et al. 2006),
and there is unpublished evidence that the fungus is
present in other parts of Western Asia (Dogmus
Lehtijarvi pers. comm.). Therefore, the pathogen
seems to continue spreading in the natural range of
P. orientalis in Western Asia with apparently no
constraints.
Acknowledgements The authors wish to thank the
anonymous reviewers and the editors for their helpful
comments and revision of the manuscript. Dr. Lorenzo Bonosi
is warmly acknowledged for revising the manuscript. The
International Union of Forest Research Organisations, Task
Force on Biological Invasions in Forests, and The Organisation
for Economic Co-operation and Development (OECD) are
acknowledged for financial support. Funding was provided by
OECD (Grant No. TAD/CRP JA87649).
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