Forest Ecology and Management 307 (2013) 165–177

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Forest Ecology and Management

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Review

Identifying and managing the ecological risks of using introduced treespecies in Sweden’s production forestry

0378-1127/$ - see front matter � 2013 Elsevier B.V. All rights reserved.http://dx.doi.org/10.1016/j.foreco.2013.06.059

⇑ Corresponding author. Tel.: +46 0730516180.E-mail addresses: adam.felton@slu.se (A. Felton), johanna.boberg@slu.se

(J. Boberg), Christer.Bjorkman@slu.se (C. Björkman), olof.widenfalk@skogforsk.se(O. Widenfalk).

Adam Felton a,⇑, Johanna Boberg b, Christer Björkman c, Olof Widenfalk d

a Southern Swedish Forest Research Centre, Swedish University of Agricultural Sciences, Box 49, Rörsjöv 1, 230 53 Alnarp, Swedenb Department of Forest Mycology and Pathology, Swedish University of Agricultural Sciences, Box 7026, Ulls v 26A, 750 07 Uppsala, Swedenc Department of Ecology, Swedish University of Agricultural Sciences, Box 7044, 750 07 Uppsala, Swedend Skogforsk, Forestry Research Institute of Sweden, Uppsala Science Park, SE-751 83 Uppsala, Sweden

a r t i c l e i n f o a b s t r a c t

Article history:Received 15 May 2013Received in revised form 26 June 2013Accepted 28 June 2013Available online 3 August 2013

Keywords:Introduced speciesEcological riskInvasiveBiodiversityPest

Introduced tree species are increasingly being considered for use in production forestry due to productiontargets, and demand for a diversity of wood products. However, prior to expanding their use, active con-sideration needs to be given to the breadth of potential ecological consequences associated with eachintroduced tree species. Ecological consequences include the invasion and modification of sensitive eco-systems, changes in habitat provision for native taxa, altered risk of pest and pathogen outbreaks, andhybridization with native con-generics. Here we review the scientific literature to assess the potentialecological consequences from expanding the use of introduced tree species within Swedish forestry.We use an interdisciplinary approach to evaluate ecological risks, and our assessment is based on the sce-nario that a proportion of Norway spruce (Picea abies) monocultures in southern Sweden will be replacedby monocultures of Sycamore maple (Acer pseudoplatanus), Douglas fir (Pseudotsuga menziesii), hybridaspen (Populus tremula tremuloides), or hybrid larch (Larix eurolepis/L. marschlinsii). Our results highlightthat univariate consideration of the ecological consequences of exotic tree species can be highly mislead-ing, due to the complex suite of costs, benefits, risks and uncertainties that each tree species brings to theregion of introduction. We discuss our results in relation to conflicting management goals, and the lack ofreversibility of some adverse ecological outcomes. We also highlight the need for assessments of ecolog-ical risk to facilitate evidence-based decision making by stakeholders. Our results provide a foundationfor adaptive management programs aiming to limit the extent to which introduced tree species usedin production forestry are accompanied by adverse ecological impacts.

� 2013 Elsevier B.V. All rights reserved.

Contents

1. Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1662. Methods . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1663. Results. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 167

3.1. Acer pseudoplatanus . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1673.2. Pseudotsuga menziesii . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1683.3. Populus tremula tremuloides . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1703.4. Larix eurolepis/L. marschlinsii. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 171

4. Discussion. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 172Acknowledgements . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 174References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 174

166 A. Felton et al. / Forest Ecology and Management 307 (2013) 165–177

1. Introduction

Approximately 25% of the world’s plantations consist of intro-duced tree species (FAO, 2010). This level of usage is driven pri-marily by the capacity of introduced tree species to achieveincreased levels of production, or to provide timbers with specifi-cally desired wood characteristics (Richardson, 1998; Dodet andCollet, 2012). Yet, the associated production benefits of using intro-duced tree species must be considered in light of the associatedecological risks. The use of introduced tree species in plantationshas exacerbated declines in stand-level biodiversity (Peterken,2001), contributed to pest and pathogen introductions (Wingfieldet al., 2001), has resulted in the invasion of sensitive ecosystems(Essl et al., 2010) and may also lead to the genetic dilution of nativecon-generics via hybridization (McKay et al., 2005; Goto et al.,2011). Solely with respect to invasiveness, approximately 60% ofinvasive tree species identified are used in forestry (Haysom andMurphy, 2003). Furthermore, assessments of the conifer familyPinaceae reveals that species used in commercial forestry are sig-nificantly more likely to escape, and become naturalized or inva-sive than those species not selected for use in forestry (Essl et al.,2010; see Castro-Díez et al., 2011, for relevant discussion of Austra-lian Acacias). Countries are therefore faced with the difficulttrade-off between using introduced tree species to achieve desiredproduction gains and avoiding adverse ecological impacts associ-ated with the extensive usage of introduced tree species.

Sweden’s forests are extensive, covering seventy percent of thecountry’s total land area (SFA, 2010); but the vast majority bearonly superficial resemblance to unmanaged forest ecosystems.Over 90% of Sweden’s productive forest land is use for forestry, withthe majority of production stands subjected to rotational clear-cut-ting, often in combination with extensive thinning and soil scarifi-cation. As a result, forest production outcomes are exceptional, asthe country can provide 10% of the world’s sawn timber, pulp andpaper using slightly less than 1% of the world’s commercial forestarea (Lundgren and Ingemarson, 2009). In 2003 the annual cut inSweden reached a level that approximated the sustainable maxi-mum annual cut (Nilsson et al., 2011). Whereas this could haveresulted in stabilization of forest productivity, it has insteadprompted discussion regarding how forest yield could be further in-creased. Recent analyses have highlighted the potential for usingintroduced tree species to achieve these increases (Nilsson et al.,2011), and as a means of diversifying timber production as part ofclimate change adaptation efforts. The potential for the expandedusage of introduced tree species is large, as only a small proportionof Sweden’s forest area is currently comprised of non-endemic treespecies (�2%, Forest Europe, 2011). As such, increased reliance onintroduced tree species could cause a substantial shift in the char-acter of Sweden’s forests. Furthermore, as the vast majority of forestarea in Sweden is allocated for production and lies outside of anysecure protective framework (Gustafsson and Perhans, 2010; CBD,2011), changes to these production forests has the potential to fur-ther affect the conservation status of many of Sweden’s forest-dependent animal and plant species.

Here we assess the potential ecological costs and benefits from asuggested increased reliance on four tree species introduced tosouthern Sweden for use in production forestry. Our assessment isbased on the scenario of a proportion of Norway spruce (Picea abies)monocultures in southern Sweden being replaced by monoculturesof either Sycamore maple (Acer pseudoplatanus), Douglas fir(Pseudotsuga menziesii), hybrid aspen (Populus tremula tremuloides),or hybrid larch (Larix eurolepis/L. marschlinsii). We conduct a reviewof the available evidence for the risk of these tree species to; (1) alterthe composition of forest-dependent taxa in production stands (forbetter or worse), (2) become invasive, (3) hybridize with endemic

tree species, and to facilitate outbreaks of (4) pathogens and (5)pests. We use our results to highlight the necessity for species-spe-cific and comprehensive considerations of ecological risks whendeveloping management strategies for introduced tree species.We discuss the complicating issues of conflicting managementgoals, the lack of reversibility of some adverse ecological outcomes,and highlight the need for assessments of ecological risk to facilitateevidence-based decision making by stakeholders. We also discusshow uncertainty may be reduced by targeted research and adaptivemanagement programs, and the overriding benefits of shifting to-wards more resilient ecosystems. We see this exercise as an essen-tial step to fostering discussion among interested parties regardingthe alternative forest futures available, and their associated implica-tions for forest ecology.

2. Methods

We searched electronic databases using different combinationsof Boolean search terms. The databases used were Google (http://www.google.se/), Google Scholar (http://scholar.google.se/), andWeb of Science (http://www.isiwebofknowledge.com/). We usedthe following search terms: ‘‘Acer pseudoplatanus’’, ‘‘Pseudotsugamenziesii’’, ‘‘Populus tremula’’, ‘‘Populus tremuloides’’, ‘‘Larix eurol-epis’’, ‘‘Larix marschlinsii’’, ‘‘Larix decidua’’, ‘‘Larix kaempferi’’, ‘‘Syca-more maple’’, ‘‘Douglas fir’’, ‘‘hybrid aspen’’, ‘‘hybrid larch’’,‘‘invasiv*’’, ‘‘hybrid*’’, ‘‘pest*’’, ‘‘patho*’’, ‘‘disease*’’, ‘‘fung*’’, ‘‘rot*’’,‘‘decay*’’, ‘‘biodivers*’’, and ‘‘conserv*’’ (hybrid larch is referred toas both Larix marschlinsii and Larix eurolepis). Search terms wererun in separate or limited combinations depending on the require-ments or limitations of the database used. We also obtained papersfrom colleagues and through reference lists from published studiesincluding major review articles and books. Furthermore, we ob-tained information from government studies, authorities and re-ports (see Table 1).

We define ‘‘risk’’ as the capacity for a chosen action to result inan undesirable outcome, and assess these risks by estimating theirlikelihood. We define the terms ‘‘introduced’’, ‘‘naturalized’’, and‘‘invasive’’ based on the criteria of Richardson and Rejmanek(2004). We define ‘‘introduced’’ taxa as a species which occurs out-side of its natural range. We define a species as ‘‘naturalized’’ if it isable to independently reproduce and sustain populations over sev-eral life cycles (Richardson and Rejmanek, 2004; Broncano et al.,2005). We define ‘‘invasive’’ species as those which produces largenumbers of offspring at considerable distances (>100 m) from par-ent plants (Richardson and Rejmanek, 2004).

Assessing ecological implications from the widespread use of anintroduced tree species requires an anchor from which to base anycomparison. As 47% of production forests’ standing volume insouthern Sweden (Götaland, approximately defined as south of59�N) is composed of Norway spruce (Pinus abies) (SFA, 2011),we have chosen this land-use as our baseline for such a compari-son. The four introduced tree species we consider are generallyestablished in southern Sweden on land previously dedicated tothe production of Norway spruce. The only notable exception is hy-brid aspen, which is also established on previous agricultural land.No species considered presently contributes more than a smallfraction of total standing volume on productive forest land in Swe-den (SFA, 2011).

In order to compare the risks associated with the four tree spe-cies, we adopted a graphical method based on a pie chart. Each treespecies is represented by a pie divided into five slices, each slicerepresenting one of five ecological issues considered. The biodiver-sity slice is green, whereas the other issues (hybridization, inva-siveness, pests and pathogens) are red. This is to indicate that

Table 1The number of articles returned from the Web of Science search conducted during 2012 using search string combinations indicated (see Section 2). Some combinations of treespecies and ecological concerns (e.g. hybridization and hybrid tree species) were more prone to returning multiple articles of limited relevance. Because of this, caution iswarranted when using the table to compare the relative availability of articles for the tree species and issues considered. These results formed the basis for obtaining additionalarticles from reference lists, colleagues, books, government studies, authorities, and reports.

Sycamore maple Douglas fir Hybrid aspen Hybrid larch

Invasiveness 28 81 24 26Hybridization 11 87 317 61Biodiversity 58 177 213 309Pests and pathogens 716 1020 237 312

A. Felton et al. / Forest Ecology and Management 307 (2013) 165–177 167

biodiversity operates on a positive scale, with likely increases inbiodiversity desirable; in contrast to the four ecological risks, forwhich increasing risk is undesirable. We can then indicate the rel-ative likelihood of a desirable or undesirable outcome arising, byincreasing or decreasing the size of each slice. To acknowledgethe course nature of such assessments we limit ourselves to thebroad categories of low, medium or high. For example, a tree spe-cies with a demonstrated capacity to invade ecosystems in Scandi-navia, either by seed spread or hybridization, would be deemed tobe of high risk of invasiveness for Sweden. Whereas a tree speciespossessing ecological attributes associated with the capacity to in-vade, but with no observed tendencies for invasiveness in northernEurope, would be deemed of medium risk. Finally, a tree specieswith no ecological attributes associated with invasiveness, andno previous invasive events after an introduction, would bedeemed of low risk of invasiveness. Fig. 1 is provided to indicatehow a hypothetical best and worst case scenario would be repre-sented using this approach. To acknowledge the inherent uncer-tainties involved in making such assessments, and the frequentlack of knowledge, we use faded colors when necessary to indicatean uncertainty range that exceeds broad categorization (i.e. low tomedium risk).

3. Results

3.1. Acer pseudoplatanus

Sycamore maple was first recorded in southern Sweden in the1800s (Berg and Nilsson, 1997), with naturalization taking placerelatively quickly (Fremstad and Elven, 1996). It is officially consid-ered an introduced species (Skogsstyrelsen, 2009). Plantations of

Fig. 1. Graphical illustration of a hypothetical best and worst case scenario for an introduslices for each ecological issue considered. Biodiversity is represented in green, with the fscale, with likely increases in biodiversity desirable; in contrast to the four ecological risdesirable or undesirable outcome arising by increasing or decreasing the size of each slicin other cases may be the causal process for a species being considered invasive. (For inteweb version of this article.)

sycamore maple currently occupy approximately 140 ha of forestarea in southern Sweden (pers. comm. Emma Holmström), thoughthis estimate is potentially low. This is due to limitations in theavailable database, stemming from the practice of grouping syca-more maple with ‘‘other broadleaves’’ in official statistics (SFA,2011), and because stands less than 0.5 ha are not required to bereported (Skogsstyrelsen, 2009). Sycamore maple’s natural distri-bution was proximate to Sweden by the 1600s, with the tree spe-cies endemic range extending to south-west Denmark by this time(Tillisch, 2001; Weidema and Buchwald, 2010). As many tree spe-cies in Europe have been shifting towards higher latitudes at a rateapproximating 50–200 km per century as part of a post-glacial ad-vance (Birks, 1989; Huntley and Webb, 1989), several authors ar-gue that sycamore maple’s potential range naturally extends wellinto Sweden (Weidema and Buchwald, 2010), raising questionsas to its introduced status.

Sycamore maple grows best on high pH soils possessing ade-quate depth and moisture, and is capable of germinating and estab-lishing on a relative wide range of pH regimes and soil types (Jones,1945; Evans, 1984). Once established, maturity is reached at arelatively young age, (Boyd, 1992; Burschel and Huss, 1997; Bart-hélémy et al., 2009) with seed production becoming increasinglyprolific with increasing size (Moller, 1965; Rusanen and Myking,2003). Their winged seeds are wind dispersed and can spread sev-eral kilometers (Hegi, 1924 cited in Weidema and Buchwald, 2010;Townsend, 2008), though the principle area of seed rain is generallylimited to 200 m from the parent tree (Tillisch, 2001; Townsend,2008). Sycamore maple’s seed bank is relatively short lived, andestablishment relies upon the repetitive blanketing of neighboringareas in a seed rain (Boyd, 1992; Deiller et al., 2003; Hérault et al.,2004; Townsend, 2008). Although shade tolerance is high in youth,

ced tree species, with each tree species assigned its own pie chart composed of fiveour ecological risks in red. This is to indicate that biodiversity operates on a positiveks, for which increasing risk is undesirable. We indicate the relative likelihood of ae. Note that hybridization risk can be disregarded in certain circumstances, whereasrpretation of the references to color in this figure legend, the reader is referred to the

168 A. Felton et al. / Forest Ecology and Management 307 (2013) 165–177

and can exceed that of beech and ash (Weidema and Buchwald,2010), the seedling’s demand for light nevertheless increases withage (Moller, 1965; Evans, 1984; Delagrange et al., 2006).

Sycamore maple is thus characterized as a shade-tolerant, light-demanding disturbance specialist that colonize woodland edgeswhere disturbance permits (e.g. water erosion, construction, for-estry), and where soils conditions are favorable. In such cases, thistree species can dominate tree species composition after distur-bances such as fire, landslide, wind-throw or clear cutting (Moller,1965; Evans, 1984; Boyd, 1992; Collet et al., 2008). It is these char-acteristics, and observed patterns of long-distance establishmentand occasional dominance of even natural forests that has led tosycamore maple’s classification as invasive in Norway, Sweden,Latvia, Lithuania, and potentially invasive in Finland (Skogsstyrel-sen, 2009; Weidema and Buchwald, 2010). In western Norwayfor example, sycamore maple is capable of invading nature re-serves previously dominated by native deciduous forests (Holtenand Brevik, 1999; Weidema and Buchwald, 2010). It is notablehowever that in regions projected to experience reduced rainfallor increased evapotranspiration due to climate change, sycamore’sinvasive status may be curtailed by its drought intolerance (More-croft et al., 2008; Townsend, 2008), as well as by frost damage toseedlings (Mollerova, 2005), browsing pressure (Hein et al., 2009;Weidema and Buchwald, 2010) and competition with grasses andother dense understory vegetation (Hein et al., 2009). It is alsonotable that under more natural forest conditions, as when co-occurring with ash and beech in its native range, Sycamore maplerarely exceeds 5% of stand composition (Larsen, 1997), and in someregions (UK) is considered to have limited risk of invading undis-turbed forest ecosystems (Townsend, 2008).

Sycamore’s ability to invade some ecosystems may be con-trasted with its capacity to provide habitat and resources for for-est-dependent biodiversity. Specifically, the texture, porosity andpH of sycamore’s bark makes it highly supportive of lichens, bryo-phytes and fungi, to an extent comparable to elm Ulmus laevis, ashFraxinus excelsior, and oak Quercus robur (Boyd, 1992; Binggeli,1993; Alexander et al., 2006). This raises the potential for sycamoremaple establishment to provide some level of ecological compen-sation for elm and ash dieback in the region. In addition to its highseed set, which is of benefit to rodents (Binggeli, 1993), sycamorealso produces vast quantities of nectar, pollen and sap, and therebycan support a high biomass of pollinating and herbivorous insects(Peck, 1989; Binggeli, 1993; but see Peterken, 2001; Alexanderet al., 2006; Townsend, 2008; Weidema and Buchwald, 2010). Syc-amore’s leaf litter is renowned for breaking down quickly, whichmay limit the depth of leaf litter, but provide for a high nutrientrecycling rate and an increased biomass of worms (Moller, 1965;Stern, 1989; Boyd, 1992; Alexander et al., 2006). The tree species’association with both high insect biomass and worm fauna isknown to facilitate a correspondingly high association with severalbird species, including chaffinch, goldcrest, tits, and woodcock inthe UK (Peck, 1989; Boyd, 1992).

The primary area for concern, with respect to stand level biodi-versity, arises from over-shading of the understory after canopyclosure in dense stands. The level of shade is comparable to thatfound in beech stands (Townsend, 2008), which may have adverseimpacts on some taxonomic groups like bryophytes (Binggeli,1993). Notably though, in a study of understory plants in southernSweden, the species composition of Sycamore stands did not sig-nificantly differ from that found under oak plantations (Brunet,2007).

With respect to pathogens, Sycamore maple is frequently in-fected by various foliar pathogens causing leaf spot (e.g. Cristular-iella depraedans) and tar spot (e.g. Rhytisma acerinum), but theseare not considered to cause significant damage to the trees. Poten-tially more damaging diseases include those caused by Phytophtho-

ra spp. The susceptibility of Sycamore maple to Phytophthorainfection varies depending on the specific species of pathogen,and inoculation studies have shown that Sycamore maple is rela-tively susceptible to P. cactorum (Holub et al., 2010), but displayedlow susceptibility against P. ramorum (Denman et al., 2005). P.cactorum has a broad host range and has recently been found ondiseased beech trees (Fagus sylvatica) in southern Sweden (Samu-elsson et al., 2012). Sycamore maple is also susceptible to sootybark disease (Cryptostroma corticale), which can cause cankersand bark necrosis; but outbreaks are mainly reported from theUK, and only in association with drought episodes (Paviour-Smith,1976; Gregory, 1982).

The insects attacking Sycamore maple are generally of little eco-nomic importance since they seldom kill or cause substantial dam-age to the trees. The notable exception is the scale insect Pulvinariaregalis, which attacks Sycamore maple and many other deciduoustree species. Sycamore maple seem to be more sensitive to P. reg-alis than other host trees, as exhibited by dramatic reductions inshoot length when heavily attacked as young trees (Speight,1991). Pulvinaria regalis was recently detected for the first timein Sweden (Gertsson, 2011) and has been shown to be increasingin density and range in the Netherlands, presumably as a conse-quence of climate change (Moraal and Akkerhuis, 2011).

In summary, the primary ecological concern associated withSycamore maple is its demonstrated capacity for invasiveness(Fig. 2). For example, individuals have now established themselvesin both disturbed and undisturbed sections of the beech forests ofSöderåsen national park in southern Sweden (pers. comm. RolandLarsson and Hans Wieslander). Here, as in other locations in Eur-ope (Townsend, 2008; Weidema and Buchwald, 2010), the syca-more regenerates vigorously in response to removal attempts. Incontrast to its invasiveness, biodiversity is projected to benefit, atleast at the stand scale, when replacing spruce monocultures withSycamore. Hybridization risks can be disregarded, due to the over-lapping distributions and evolved speciation barriers between Syc-amore maple and the two Swedish endemic species of maple Acercampestre and Acer platanoides. Whereas pest and pathogen risksare also considered to be low, the uncertainty regarding future out-breaks makes us judge the potential ecological risks as medium.The limited known pathogens and insect pests in Sweden, Syca-more maple’s proximate distribution, and endemic con-generics,all contribute to its low relative risk score. However, the suscepti-bility of Sycamore maple to Phytophthora spp. in general and P.cactorum in particular, could increase the potential risk of in-creased spread of these pathogens into nearby forest stands pos-sessing susceptible tree species. Furthermore, the risk ofdramatic growth reductions in young trees as a consequence ofP. regalis attacks could become increasingly problematic with fur-ther climatic change.

3.2. Pseudotsuga menziesii

Douglas fir has an extensive natural distribution which extendsfrom British Columbia to the mountains of central Mexico (Her-mann and Lavender, 1990). Its capacity to grow under a range ofgrowing conditions (Orellana and Raffaele, 2010) has contributedto its introduction and use throughout Europe, and as an alterna-tive to Norway spruce in Sweden (Skogsstyrelsen, 2009). As a re-sult Douglas fir currently has the widest distribution of anyintroduced conifer tree species in Europe, and has naturalized in12 countries (Amparo Carrillo-Gavilan and Vila, 2010). Douglasfir was first introduced to Sweden in the 1920s, and plantationsare currently estimated to occupy approximately 500 ha, primarilyon large estates in southern Sweden (pers. comm. Kristina Wall-ertz). Douglas fir has no extant con-generics in Europe, but fossil

Fig. 2. Graphical representation used for comparing the relative ecological risks, biodiversity benefits and corresponding uncertainties involved in converting Norway sprucemonocultures to any of the four introduced tree species considered for increased use in southern Sweden.

A. Felton et al. / Forest Ecology and Management 307 (2013) 165–177 169

evidence indicates the presence of Pseudotsuga spp. in Europe untilthe mid-Pleistocene (Hermann, 1985).

Douglas fir possesses a suite of ecological characteristics repeat-edly associated with invasive tendencies in conifers; includingsmall seed mass (<50 mg, further aided by the possession ofwings), early maturity (<10 years), and short intervals (<3 years)between large seed crops (Ledgard, 2002; Richardson and Rej-manek, 2004). The vast majority of seeds fall within 50–75 m ofthe parent tree (Malcolm et al., 2001; Jonasova et al., 2006),although seedlings may be encountered at 200 m or further (Sim-berloff et al., 2002; Orellana and Raffaele, 2010). Douglas fir ap-pears to benefit from canopy cover during regeneration, butsubsequently requires higher light levels for continued growth(Jonasova et al., 2006). Grasslands and shrublands thus appear tobe more susceptible to invasion by Douglas fir than forestlands(Richardson and Higgins, 1998), with the susceptibility of foreststo invasion dependent on their openness and disturbance fre-quency (Richardson and Bond, 1991; Hermann and Lavender,1999; Broncano et al., 2005). For example, Douglas fir is currentlyinvading from neighboring plantations into the heathlands ofMontseny Natural Park, Spain (Broncano et al., 2005). Douglas firmay therefore be considered a semi-shade tolerant pioneer specieswith a demonstrated invasiveness capacity, to the extent that it isconsidered one of the most invasive conifers used in forestryworldwide (Richardson and Rejmanek, 2004). The species is listedas invasive in seven countries, including Germany, Great Britain,and Bulgaria in Europe (Nunez and Medley, 2011). Notably how-ever, the risk of conifer invasion in Europe may be considered lessthan in the southern hemisphere, which may result from (1) a rel-atively limited conifer introduction effort, (2) the phylogenetic

similarities of introduced conifers to native taxa with associatedsusceptibility to native pests and pathogens, or (3) a time-lag inproblem development, perception, or recognition (Amparo Carril-lo-Gavilan and Vila, 2010). In Sweden, the capacity of this speciesto invade may also be limited by its sensitivity to browsing pres-sure, spring frosts, and freezing (Skogsstyrelsen, 2009).

With respect to biodiversity, Douglas fir generally hosts fewerinvertebrate taxa than native tree species (with the possible excep-tion of some bark beetle groups), although the precise taxonomicoutcomes will depend on the guild examined, the comparator treespecies, and the context of introduction in the region assessed(Alexander et al., 2006; Gossner and Ammer, 2006; Finch and Szu-melda, 2007; Bertheau et al., 2009; Gossner et al., 2009). For exam-ple, the thick and heterogeneous nature of Douglas fir bark issimilar to Scots pine (Pinus sylvestris), and after intensive Douglasfir introduction in France, and perhaps as a result of positive selec-tion pressure (Degomez and Wagner, 2001; Bertheau et al., 2009),Douglas fir has been found to support a substantial number of barkbeetles (Bertheau et al., 2009). In contrast, studies of Douglas fir inGermany suggest that the associated beetle fauna is limited rela-tive to Norway spruce, especially for those taxa breeding in thebark cortex (Gossner and Ammer, 2006). The generic reduction inarthropods appear to be sufficient to have corresponding adverseimpacts on foraging bird activity in some regions (Gossner and Uts-chick, 2004). With respect to the understory, the acidic nature ofDouglas fir’s needles, and the high density of foliage, results in anenvironment limited in light and low in pH. This environmenttends to restrict the growth of understory herbs and favor mosses(Finch and Szumelda, 2007; Dehlin et al., 2008). Similarly, coniferbark is generally acidic, and combined with dense foliage, greatly

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reduces the suitability of these stands as substrates for epiphytes(Alexander et al., 2006). Over the longer term, the similar plant sec-ondary compounds that occur in Douglas fir and Norway spruce(Picea abies), and the relatively rapid rate of endemic taxa associa-tion with Douglas fir in some European countries, might be indic-ative of the potential for colonization by endemic invertebratefauna in Sweden, despite the lack of Pseudotsuga congenerics(Burzlaff, 1998 cited in Gossner and Ammer, 2006; Roques et al.,2006).

Douglas fir has a history of disease problems in Europe, due tointroductions of Swiss needle cast fungi (Phaeocryptopus gaeuman-nii), and Douglas-fir needle blight (Rhabdocline pseudotsugae),which were introduced with the tree species to Europe. Bothcaused damage and production losses (see Hansen et al., 2000),but differences in susceptibility among subspecies of Douglas fir(P. menzeiesii vars. galuca and caesia is more susceptible than var.viridis), facilitated the use of targeted breeding, appropriate seedand provenance selection to control the problem (see Zobel et al.,1987). Notably however, the damage caused by P. gaeumannii inits native range has increased since the mid-1990s (Hansen et al.,2000), and at present causes significant production losses in NewZealand (Kimberley et al., 2011). In both continents disease out-breaks correlate with certain climatic and weather conditions (Ros-so and Hansen, 2003; Watt et al., 2010). Douglas fir production isalso affected in its native range by several different rot fungi,including Phellinus weirii (laminated root rot), that do not as yet oc-cur in Europe, but currently has a quarantine status (OEPP/EPPO,1979).

Other forest pathogens of potential concern in Sweden includePhytophthora ramorum, for which Douglas fir is known to be sus-ceptible to infection in inoculation trials (Hansen et al., 2005; Den-man et al., 2006). However, infection of Douglas fir by thispathogen in the field has only been observed in proximity to otherplant species with high levels of infection (Davidson et al., 2002).Inoculation trials have also revealed Douglas fir to be susceptibleto the Norway spruce pathogen Ophiostoma polonicum (Christian-sen and Solheim, 1994), which could be a threat should the Doug-las fir bark beetle (Dendroctonus pseudotsugae) be introduced andact as a vector (Roques et al., 2006). In addition, Douglas fir can alsobe damaged by the native Heterobasidion annosum sl. root rot,which raises the potential for disease transfer from infected Nor-way Spruce stands (Rönnberg et al., 1999) prevalent in southernSweden (Stenlid and Wasterlund, 1986).

At present, there are no known insect pests that constitute animminent threat to Douglas fir, which may be explained by the lackof close relatives among native tree species. However, insects withmore generalist host-plant utilization are of concern (cf. Dalin andBjörkman, 2006). For example, the native pine weevil Hylobiusabietis has the capacity to be a pest on seedlings during the estab-lishment phase (Wainhouse et al., 2001). In addition, the bark bee-tle Pityophtorus pityographus was recently discovered in Sweden,and can cause minor damage to trees by feeding on its smallerbranches (Ericson, 2010; Lindelöw, 2012a). Potential pest insectsthat do not as yet occur in Sweden include the beetle species Xylos-andrus germanus (widespread in Europe) and the North Americanaphid Gilleteella cooleyi (introduced to UK and Ireland) (Cameron,1936). The risk that these species will enter and establish in Swe-den will depend on trade, quarantine measures, climate change,and the scale to which other host tree species are planted (Liebholdet al., 2012).

In summary, Douglas fir is characterized by high risks associ-ated with invasiveness, pest and pathogen outbreaks, and highuncertainty with respect to the true extent of these risks (Fig. 2).Douglas fir has a demonstrated capacity for invasiveness in Europe,but unlike Sycamore maple, this capacity has not manifested itselfin Sweden as of yet. Likewise, there are a number of pests and

pathogens that are present in Sweden and currently causing signif-icant damage in other countries, as well as known pests which areyet to establish in Sweden, and endemic pests which have the po-tential to cause future outbreaks. There are thus a substantialamount of potential concerns regarding pests and pathogens,which must be balanced by the current absence of any eventuali-ties. With respect to biodiversity, Douglas fir is less favorable dueto its conifer associated attributes (e.g. acidic bark, dense foliage),which limit the potential landscape scale benefits of shifting to thistaxa from Norway spruce. At the stand level, its North Americanstatus and associated lack of co-evolved European species limitsbiodiversity. The only ecological risk not of concern is hybridiza-tion risk, due to the lack of any endemic con-generics in Europe.

3.3. Populus tremula tremuloides

Hybrid aspen (Populus � wettsteinii Hämet-Ahti) is a cross be-tween European aspen Populus tremula and North American trem-bling aspen P. tremuloides, first described in Germany in the 1920s(Wettstein, 1933). Populus tremula and P. tremuloides are closely re-lated species (Cervera et al., 2005), and some suggest they repre-sent a single circumboreal ring-species (MacKenzie, 2010; Tulluset al., 2012). The hybrid variant is considered to be one of the fast-est growing deciduous trees in Europe (Tullus et al., 2007, 2012),and is often used to meet wood fiber and bioenergy needs, oftenvia short rotation forestry (SRF) (EU, 2009; FAO, 2012). For SRF, fi-nal harvest may take place between 20 and 30 years after estab-lishment (Tullus et al., 2012). Hybrid aspen is best suited tonutrient-rich, well-aerated, moderately drained soil with a highcapacity for holding water (e.g. Albeluvisols, Luvisols and Plano-sols) (Tullus et al., 2012), and is often established on abandonedagricultural lands or where spruce monocultures previously oc-curred. The first trials of hybrid aspen plantations in Sweden beganin the 1940s (Rytter and Stener, 2005), with the present plantedarea now estimated to be over 2000 ha (Rytter et al., 2011). Dueto the attraction of this tree species to browsing herbivores, andthe high abundance of roe deer and moose in Sweden (Seiler,2004), fencing is often recommended (Rytter et al., 2011).

Aspens are light-demanding pioneers, often colonizing dis-turbed sites with their prolific production of wind dispersed smalldowny seeds, and may thereby establish in post-fire or clear-cutsites long distances from parent populations (Myking et al.,2011; Tullus et al., 2012). Hybrid aspen shares these characteris-tics, but possesses faster growth rates than either P. tremula or P.tremuloides (Rytter and Stener, 2005; Rytter, 2006; Koivurantaet al., 2012). Depending on site and climatic conditions aspensmay begin flowering and seeding anywhere from 10 to 20 yearsof age (Borset, 1985). They are dioecious, wind pollinated specieswhich are capable of reproducing sexually and asexually via theuse of root suckers (Koivuranta et al., 2012). Although someauthors emphasize the use of suckering for reproduction (Latva-Karjanmaa et al., 2003; Tullus et al., 2012), Myking et al. (2011)highlights the evidence for the importance of sexual reproductionin P. tremula, including the presence of genetic markers indicatingseed flow between aspen refugia during the Quaternary; rapidpost-glacial advance by this species in Europe (Huntley and Birks1983); high levels of within population genetic variation(Lopez-de-Heredia et al., 2004; Suvanto and Latva-Karjanmaa,2005), high levels of seed production (Worrell et al., 1999); andthe presence of aspens on islands and remote isolated areas(Myking et al., 2011). The importance of sexual reproduction forEuropean aspen has direct relevance to concerns regarding the po-tential for introgression between hybrid varieties and wild popula-tions of European aspen. These concerns include, temporal overlapin flowering (but see Tullus et al., 2012) in some regions of Fenno-scandia (R. Jaantinen, 2010 cited in Koivuranta et al., 2012); and

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their demonstrated capacity to cross and produce viable seed (Suv-anto and Pulkkinen, 2004; Koivuranta et al., 2012). Second, as pol-len disperses more effectively than seed, this greatly increases thepotential for invasive success, via the introgression with Europeanaspen (Koivuranta et al., 2012). Third, hybrid aspen clones are pro-duced using a small number of siring Populus tremuloides, thus rais-ing concerns regarding the implications of introgression on theEuropean aspen gene pool (Koivuranta et al., 2012). Fourth, hybridaspen’s faster growth capacity, increased production of viablehighly germinal seeds, and likely variation among clones in levelsof chemical defense against browsing damage (Bailey et al.,2007; Myking et al., 2011), all raise concerns regarding the poten-tial for introgression to result in the competitive exclusion of Euro-pean aspen and associated ecological impacts (Myking, 2002;Koivuranta et al., 2012). Furthermore, as European aspen occursthroughout Sweden, hybrid aspen will regionally co-occur withthe endemic European aspen wherever hybrid aspen is plantedwithin Sweden. It is for such reasons that gene flow from hybridaspen to native European aspen populations is often considered(but see Tullus et al., 2012) a ‘‘serious risk’’ in Sweden (Strengbom,2009 in Norén and Ringagård, 2009), and hybrid aspen is thereforelisted as an invasive species in Sweden (Skogsstyrelsen, 2009).

In contrast to these risks, stand-level biodiversity benefits fromthe conversion of Norway spruce stands to hybrid aspen could oc-cur due to the phylogenetic proximity of hybrid aspen to Europeanaspen, and due to increased landscape heterogeneity from theplanting of a broadleaf tree species within conifer dominated land-scapes. At the stand level, the nutrient-rich and high pH bark of as-pen benefits many epiphytic bryophytes and lichens in Scandinavia(Gustafsson and Eriksson, 1995), including many red-listed species(Hallingbäck, 1996). However, such benefits are primarily associ-ated with the bark conditions of old large trees (Esseen et al.,1997; Hazell et al., 1998) and not those provided by short rotationforestry (Peterken, 1999; Weih et al., 2003). Such benefits could beattained if short-rotation forestry includes green-tree retention atthe time of final felling. Furthermore, whereas vascular plant spe-cies richness may be similar in hybrid aspen stands to that foundon agricultural land, or even old-growth deciduous stands insouthern and central Sweden (Weih et al., 2003), species composi-tion is often dominated by light-demanding ruderals or generalists,and lacks forest-associated, rare, or endangered species (Weihet al., 2003; Soo et al., 2009a,b; Tullus et al., 2012). This patterncan alter as canopy cover increases, and depending on tree spacingand rotation periods (Soo et al., 2009a). Likewise, the degree towhich the original understory vegetation persists between rota-tions will depend on the intensity of chemical and mechanicalweed control carried out during establishment (Soo et al., 2009a;Tullus et al., 2012). Birds, mammals, and some invertebrate groupsare often more species rich in short rotation forestry stands thanon agricultural landscapes, but less than that found in native forestland (Christian et al., 1998; Blick and Burger, 2002; Dhondt andWrege, 2003; Schulz and Brauner, 2009; Tullus et al., 2012). Nota-bly, a recent meta-analysis found that Populus based short rotationforestry harbored lower levels of avian and mammalian diversitythan intensely managed forests in the United States (Riffell et al.,2011). The common practice of fencing hybrid aspen stands to re-move large browsing ungulates (Tullus et al., 2012) needs also tobe considered with respect to implications for landscape scalewildlife dispersal and migration requirements.

Selective breeding has led to improved resistance against pestsand diseases in today’s plantations, with hybrid aspen’s resistanceto many diseases now exceeding that of European aspen (Stener,2010). Outbreaks of some disease do still occur however, especiallyfrom canker causing pathogens and leaf rusts (Tullus et al., 2012).For example, damage has been reported from the stem canker fun-gi Neofabraea populi (Kasanen et al., 2002) and Entoleuca mammata

(Ilstedt and Gullberg, 1993). Other pathogens which can causegrowth reductions in hybrid aspen growth are Leucostoma niveum,causing branch cankers, Venturia spp., causing shoot blight, and thebacterial canker causing Xanthomonas populi. There are also severaldifferent leaf infecting pathogens (Melampsora, Septoria and Mars-sonina spp.) which may also have a negative impact on growth(Tullus et al., 2012). There are also other pathogens, such as Neofa-braea populi, with a large potential for causing damage (Kasanenet al., 2002), but not yet encountered in Sweden. In addition, hy-brid aspen can act as the alternate host for the pine rust fungus(Melampsora pinitorqua) which causes serious damage to youngScots pine (Mattila, 2005). Hybrid aspen does appear to be moreresistant than European aspen to infection (Kassfeldt, 2009), andensuring a distance in excess of 200 m from Scots pine can greatlylimit transmission rates and impacts (Eidmann and Klingström,1976). The most important rot fungi in hybrid aspen is Phellinustremulae, which causes heart rot. But since rot development ismainly observed in older trees (Basham, 1993), this problem mightbe of limited significance in short rotation forestry.

Among pest species, two wood and stem boring coleopteranspecies expected to cause problems in hybrid aspen are the long-horn beetles Saperda populnea and S. carcharias. Although S. carcha-rias cause damage to the base of the stem, it is S. populnea, whichattacks particularly smaller trees and causes shoot breakage thatis of potential economic importance (Rytter and Stener, 2011).The leaf-feeding chrysomelid beetle Chrysomela populi may alsocause growth reducing damage, due to the leaf feeding of boththe adult and larval stages (Rytter and Stener, 2011).

To summarize, the primary area of concern regarding hybrid as-pen is introgression with European aspen, the associated dilutionof its genome, and the invasive establishment of genetically mixedindividuals in Sweden (Fig. 2). Correspondingly, both invasivenessand hybridization risks are considered high. Of concern but less so,are the many known but presently manageable pathogens, whichtogether generates a relative risk score of medium (Fig. 2). Like-wise, pests are also considered of medium risk due to the relativelyfew but widespread pest insects of known and potential concern inSweden. We provide a medium to high evaluation with respect tobiodiversity due to both the stand and landscape level benefits ofshifting conifer dominated landscapes to stands composed ofbroadleaf tree species. In addition there is the potential for greatlyenhancing these benefits by reducing impacts on ground vegeta-tion during establishment, and the provision of habitat for taxaassociated with mature and decaying aspen (e.g. green tree reten-tion, snag creation).

3.4. Larix eurolepis/L. marschlinsii

Hybrid larch is a cross between European larch Larix deciduaand Japanese larch Larix kaempferi; neither of which are native toSweden. European larch is naturally distributed from the moun-tains of central Europe to the lowlands of southern Poland (Matrasand Paques, 2008), with Japanese larch restricted to the mountainsof central Honshu, Japan. The Siberian larch Larix sibirica is deemedendemic to Sweden, based on fossil remains dated to 8000 BPfound in the Scandes Mountains (Kullman, 1998). Nevertheless,there are no extant naturally occurring Larix populations inSweden.

European larch was introduced to Sweden in the middle of the1700s, with Japanese larch introduced in the late 1800s (Eko et al.,2004). The European and Japanese larch are closely related species,both morphologically and genetically, that hybridize easily in thewild and under experimental conditions (Paques et al., 2006). Hy-brid larch was introduced to Sweden in the mid-20th century, andthe standing volume in southern Sweden has now reached600,000 m3 (Eko et al., 2004; SFA, 2011). This makes Hybrid larch

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the most prevalent introduced species in southern Sweden (Skogs-styrelsen, 2009). Hybrid larch prefers moist rich soils, but is capa-ble of handling poorer, dry soils, as well as relatively high winddisturbance and high altitude conditions.

With respect to invasiveness, both of hybrid larch’s parent spe-cies possess positive scores for invasive potential (Richardson andRejmanek, 2004). In Europe, European larch is listed as invasive inthe Czech Republic, and Great Britain, with naturalization occur-ring in Lithuania, Ireland, Canada, and USA (Nunez et al., 2011),and potentially invasive in Latvia (NOBANIS, 2013a). Naturalregeneration does occur in hybrid larch stands, though little moreinformation is available regarding potential ecological repercus-sions (Larsson-Stern, 2003). What is established is that wildingscan escape from European larch plantations, as has been recordedto occur in Norway (Gederaas, 2007). Hybrid larch is established inSweden, but unknown with respect to its invasiveness (NOBANIS,2013b).

In relation to stand-level biodiversity, there are some indica-tions that the relatively sparse and deciduous nature of larch can-opies, can favor a relatively rich understorey of flora compared toother conifers (Barbier et al., 2008; Skogsstyrelsen, 2009; Wanget al., 2009). For example, the leaf area index for Norway spruceis twice that of European larch (Bolstad et al., 1990), with a corre-spondingly reduced diversity of understory flora (Larsson-Stern,2003). In addition, the understorey flora may in turn help to im-prove soil conditions (Wang et al., 2009), which otherwise wouldlikely consist of an acidic and nutrient-poor leaf litter (Skogsstyrel-sen, 2009). With respect to plant feeding invertebrates, Europeanlarch is species poor compared to native tree species in Britain(Peterken, 2001), but these communities are considered on parwith the amount of larch planted (Kennedy and Southwood,1984). In addition, some limited evidence suggests that coal tits,goldcrests and chaffinch may prefer foraging in European larchstands, relative to the prevalence of larch in a British landscape(Peck, 1989), with other potential benefits including this tree’sseed provisioning capacity (Alexander et al., 2006).

The pests and pathogens of hybrid larch overlap with otherEuropean Larix spp., some of which are of potential concern. Forexample, whereas the canker fungus Lachnellula willkommii canbe destructive to Larix spp., the hybrid larch is planted in partbecause of its resistance to infection by this pathogen (Sylvestre-Guinot et al., 1999). Hybrid larch can however incur substantialrot damage by Heterobasidion annosum s.s. when planted on standspreviously comprising infected Norway spruce (Rönnberg and Vol-lbrecht, 1999; Vollbrecht and Stenlid, 1999), or via primary infec-tion through the stump surface. Notably, this risk can be reducedusing stump treatments (Wang et al., 2012). Young larch treescan also succumb to infection and rot caused by different Armillariaspp., as well as sustain defoliation due to needle cast fungus Merialaricis (Cech, 2004; Maresi et al., 2004). Larch is also the alternatehost of several rust fungi which, although benign in larch, can haveadverse impacts on the growth of proximate and readily infectedbirch, poplar and Salix species (e.g. Samils et al., 2001). An addi-tional pathogen of specific concern is Phytophthora ramorum,which causes extensive damage on Japanese larch in the UK (Bra-sier and Webber, 2010). P. ramorum is of concern because of itsvery broad host range, which includes beech (Fagus sylvatica) andblueberries (Vaccinium myrtillus) (Herrero et al., 2010). It is cur-rently unknown how susceptible hybrid larch is to P. ramorum.

One of the insect species responsible for most stand relateddamage in Sweden, the pine weevil Hylobius abietis can also dam-age hybrid larch. The pine weevil damages and kills seedlings inthe regeneration phase, with clearcutting of stands apparentlyexacerbating the problem (Örlander and Nilsson, 1999; Nordlanderet al., 2003). Other bark beetles of potential concern include Ipscembrae that was recently discovered in Sweden, and I. subelonga-

tus that presently occurs in western Russia (Lindelöw, 2012b). Inaddition, the larch longicorn beetle Tetropium gabrieli was recentlydiscovered in Sweden, and is suspected to have killed medium-sized larch trees (Ericson, 2010); nevertheless its pest status re-mains unclear.

In summary, there are several ecological risks associated withhybrid larch, of which pest and pathogen outbreaks have the mostpotential concern. Its susceptibility to Heterobasidion root and buttrot, capacity to act as an alternate host to several rust fungi, anduncertainty regarding its susceptibility to P. ramorum infection,leads to a medium to high risk score for pathogens (Fig. 2). Like-wise, hybrid larch is susceptible to insect pests of substantial con-cern in Sweden, with additional concerns and uncertainties raisedregarding the potential risks of newly discovered pest species inSweden. In contrast, hybrid larch is deemed to have a medium le-vel risk of invasiveness, driven primarily by the ecological charac-teristics and demonstrated invasiveness of the parental treespecies. Biodiversity benefits are also deemed to be of mediumlikelihood, due primarily to its capacity to improve understory con-ditions relative to Norway spruce monocultures. Hybridization riskcan be disregarded, due to the lack of any naturally occurring Larixpopulations in Sweden.

4. Discussion

Considered together, the four tree species introduced to south-ern Sweden raise concerns spanning the spectrum of ecologicalrisks assessed, with each species deemed to be of potential high riskin at least one category. In addition, none of the tree species wereequivalent in terms of overall ecological risk, as each displayed aunique combination of specific risks and associated uncertainties.Using the issue of invasiveness as an example, none of the tree spe-cies considered was assessed to be of ‘‘low risk’’. A capacity to in-vade other ecosystems can in turn exacerbate other risks, such aspest and pathogen transfer to endemic tree species, as well asupgrading the scale of other ecological risks, beyond the stand. Thustree species with a demonstrated capacity for invasiveness oftenneed to be contained, and there are a number of management ap-proaches that can be used to help do so (Engelmark et al., 2001;Richardson and Rejmanek, 2004). These include the developmentof hybrids or transgenic varieties which are sterile or possess a re-duced capacity for invasiveness (Richardson, 2011), restrictingplantings to areas where the tree species is already present (Boyd,1992; Binggeli, 1993; Finch and Szumelda, 2007); limiting the totalallowable area of planting (Richardson and Rejmanek, 2004; Finchand Szumelda, 2007); preventing plantings where long distancedispersal of seed or pollen is favorable (hill tops, ridges) (Ledgard,2002), or from areas upwind or otherwise proximate to sensitiveareas (Boyd, 1992; Binggeli, 1993; Ledgard, 2002; Finch and Szu-melda, 2007). Such restrictions may then be supported by monitor-ing for wildings and targeted removal programs (Jonasova et al.,2006; Finch and Szumelda, 2007). Whereas such regulations andactions may be applicable to most invasive tree species, manage-ment strategies also need to consider the range of tree speciesspecific, and ecosystem implications that can result from suchinterventions. For example, sycamore maple populations can bene-fit from eradication attempts which cause disturbance to the forestcanopy or reduce competing ground vegetation; due to associatedbenefits to seedling establishment and regeneration (Townsend,2008). Likewise, both Sycamore maple and hybrid aspen arecapable of regenerating repeatedly after clearance attempts (Wei-dema and Buchwald, 2010; Myking et al., 2011). These responsesare consistent with the broader scale tenet that disturbance oftenbenefits the establishment and spread of invasive species (Hobbsand Huenneke, 1992; Ledgard, 2002; Richardson and Rejmanek,

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2004). As such, risks associated with the management interventionsthemselves must also be considered.

Specific issues of ecological concern may also be addressed bytargeted management interventions which may however, inadver-tently contribute to environmental problems not considered. Forexample, multiple studies highlight the potential benefits to thewithin-stand biodiversity of exotic plantations by placing themproximate to source populations of endemic understory plants(Weih et al., 2003; Brunet, 2007; Tullus et al., 2012) or inverte-brates (Gossner and Ammer, 2006; Finch and Szumelda, 2007;Gossner et al., 2009; Tullus et al., 2012). However, doing so canreadily come into conflict with management goals aimed at reduc-ing the risk of pest and pathogen transfer among exotic plantationsand native forests (Samils et al., 2001; Roques et al., 2006; Gossneret al., 2009), and the risk of invasive tree species establishing insensitive forest ecosystems (Boyd, 1992; Binggeli, 1993; Finchand Szumelda, 2007). As such, targeted efforts to reduce ecologicalrisks need to be conducted in conjunction with the wider-scaleconsideration of other environmental concerns and managementactivities (Hulme, 2006). Management activities which do comeinto conflict may in some circumstances be dealt with by prioriti-zation, the acceptance of likely trade-offs, and by suitably adjustingdifferent management strategies depending on the landscapecontext.

Notably, some of the ecological risks we evaluated lacked clearmanagement solutions. For example, there is no clear means bywhich hybridization may be prevented between hybrid aspenand European aspen. Furthermore, we do not presently know theextent to which this has already occurred, as hybrid aspen has beenplanted in Sweden since the 1940s (Rytter and Stener, 2005). Toprevent hybridization with European aspen requires preventingcross-pollination. However, pollination takes place over long dis-tances, thereby raising serious questions as to the long term effec-tiveness and feasibility of, for example, buffer zone approaches tocontainment. An added complication is that hybridization eventsare not readily detected in the field. Furthermore, experimentalevidence suggests that the genetic impact of hybrid aspen on Euro-pean aspen could increase with climatic change (Koivuranta et al.,2012).

Gene flow between hybrid aspen and European aspen is a typeof ecological risk which, like pest or pathogen introductions, maybe extremely difficult to prevent or contain once introductionhas taken place (Hulme, 2006). We refer to such ecological risksas having low reversibility. In the case of hybrid aspen, and accept-ing that the ecological risk is of concern (but see Tullus et al., 2012),the only effective means of preventing gene transfer may be thecultivation of a variety which cannot interbreed with European as-pen. In the absence of these or equivalent measures, stakeholdersneed to be made aware of the likelihood of the ecological risk(e.g. gene flow) taking place, and how low the prospect is forreversing such outcomes once they take place. In the absence ofinformation regarding the reversibility of an ecological risk, in-formed decision making cannot take place.

The issue of gene flow between hybrid aspen and European as-pen can also be used to illustrate the need to acknowledge societalvalues when evaluating ecological risks. Whereas ecologists canclarify the ecological risks involved in planting hybrid aspen, wecannot clarify the extent to which a society should prioritize avert-ing such a risk. A given set of stakeholders may decide that the po-tential benefits of using hybrid aspen for climate change mitigationsimply outweigh the potential ecological costs. In this regard sev-eral studies have emphasized that decisions regarding the impor-tation and use of introduced species have two components;ecological prediction of the relative risks, and evaluation of thoserisks through the perspectives of multiple interest groups (Ruesinket al., 1995; Hulme, 2006). The central point being that risk assess-

ments need to be undertaken, and the outcomes presented to allowfor individual stakeholders differing in their preference or aversionto specified outcomes (Richardson et al., 2009; Schlaepfer et al.,2009). For this process to take place requires that tradeoffs amongdifferent ecological risks and mitigation strategies are made expli-cit, as are levels of uncertainty and reversibility.

In some cases levels of uncertainty will be sufficiently high tohinder the capacity of researchers to clarify the ecological risksassociated with different tree species (Levine et al., 2003). If so, ex-plicit reference to these uncertainties can help motivate publicinterest and create support for further research to clarify the like-lihood of such risks. With respect to the tree species considered inthis study, the most substantial area of uncertainty is related to theecological risks of Douglas fir’s pests and pathogens. The uncer-tainty largely stemmed from inconsistency between Swedish andinternational observations of pest and pathogen damage in Doug-las fir stands. This uncertainty is then accentuated in Sweden dueto the inter-continental nature of the introduction, which therebyincludes the potential for further pest and pathogen introductionsfrom North America (Hansen et al., 2000; Bertheau et al., 2009).Likewise, Douglas fir may also provide an opportunity for indige-nous pests and pathogens to extend their host range and therebybecome problematic (Bertheau et al., 2009; Gossner et al., 2009).The difficulty in assessing such risks is not surprising consideringthe extensive variation in the susceptibility of different introducedtree species to pests and pathogens. For example, the geographicaldislocation of a tree species may decrease damage levels, due totheir escape from co-evolved pests and pathogens; a process oftenreferred to as the ‘enemy release theory’ (Colautti et al., 2004). Incontrast, introduced tree species may experience significantlyhigher levels of pest and pathogen related damage in their newenvironment (Lombardero et al., 2012). This may be due to the lackof co-evolution contributing to a higher susceptibility of the intro-duced tree species to native pests and pathogens; an interactionreferred to as the ‘biotic resistance hypothesis’ (Maron and Vila,2001). Introduced tree species may also suffer from pests andpathogens introduced simultaneously, but not considered signifi-cant problems in their native range (Boyce, 1940; Stone et al.,2008). The ability of pests and pathogens to spread and cause dam-age will likewise vary depending on the spatial distribution ofstands and may vary dramatically depending on local climate con-ditions. Climatic change is, therefore, a confounding factor thatmay enforce or mitigate pest and pathogen problems (Björkmanet al., 2011; Sturrock et al., 2011). Furthermore, pests and patho-gens associated with an introduced tree species may also threatenbiodiversity in other forest ecosystems, as some pathogens (e.g.Phytophthora spp.) and pests (e.g. Anoplophora glabripennis) areable to cause damage to many different host genera (Hansen,2008; Herrero et al., 2010). We can only suggest that the resolutionof such uncertainties preferably involves targeted research oradaptive management approaches (Rist et al., 2013), rather thanallowing the introduction to act itself as a large-scale and uncon-trolled experiment.

Our results are potentially relevant to several areas of currentregulation in Sweden relating to the use of introduced tree speciesin production forestry. For example, hybrid tree species containinggenes from Swedish tree species are not classified as introduced treespecies under the Swedish Forest Management Ordinance (Sko-gsvårdsförordningen 1993:1096), despite their apparent capacityto pose comparable ecological risks. Furthermore, whereas exclu-sion zones for introduced tree species are regulated for mountain-ous areas, they do not exist for areas proximate to national parksor nature reserves (Skogsvårdsförordningen 1993:1096). As such,whereas a production stand larger than 0.5 ha of Sycamore maplemay have its application rejected by the forest management author-ity (Skogsstyrelsen) if considered too proximate to a protected area,

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present guidelines are unclear as to what that distance is. It is alsounclear how this distance is adjusted for different introduced spe-cies varying extensively in the nature of their ecological risks. Like-wise, to be effective in the long term requires that the buffer zone iseither inhospitable to seedling establishment and regeneration, or ismonitored and cleared of wildlings. No such monitoring or clearingis currently required. Effective monitoring and containmentprograms are also often recommended with respect to pests andpathogens (see Stenlid et al., 2011; Pautasso et al., 2013), as suchprograms enable early detection and control, and ideally serve to in-crease understanding of the mechanisms involved in pest and path-ogen establishment and spread.

Our results also emphasize the influence of the comparativeland-use in dictating conclusions. We used Norway spruce mono-cultures as the relevant land-use from which to contrast the asso-ciated risks of planting introduced tree species. It is notablehowever that in many regions of southern Sweden, Norway sprucewas either absent or at extremely low densities throughout the pa-leo-ecological record (Lindbladh and Foster, 2010). Whereas thisdoes not negate its relevance as a suitable comparator (Stephensand Wagner, 2007; Paquette and Messier, 2010), it does highlightthe importance of the context in dictating outcomes. Norwayspruce monocultures are managed via a process of rotationallyclear-cutting even-aged stands. Although general environmentconsiderations are taken, including for example the retention ofsome living trees, dead wood, and high stumps (Gustafsson andPerhans, 2010), these are nevertheless intensively managed pro-duction forests (Brockerhoff et al., 2008). As such, from a standor landscape scale perspective, Norway spruce monocultures seta relatively low threshold for achieving biodiversity benefits. Atthe stand scale, Norway spruce monocultures are depauperate interms of tree species composition and structural diversity, withresultant limitations on the diversity of taxa supported by thesestands (Gärdenfors, 2005; Felton et al., 2010). At the landscapescale, any forest type which is sufficiently distinct from the domi-nant production forest to provide alternative habitat and resources,has the potential to enrich biodiversity at a landscape scale(Fischer et al., 2006; Lindenmayer et al., 2006). For both of thesereasons, the potential for biodiversity benefits arising from threeof the four tree species considered cannot be assumed to includebenefits for specialist or threatened taxa.

The intensely managed landscapes of southern Sweden, with itsdominance of production forests and agriculture, raise broaderscale concerns regarding the susceptibility of this region to theecological risks associated with introduced tree species and othernon-native species. Multiple studies have emphasized the impor-tance of the environment in determining its susceptibility to inva-sive taxa (Ledgard, 2002; Simberloff et al., 2002; Richardson andRejmanek, 2004; Broncano et al., 2005; Kulmatiski et al., 2006;Orellana and Raffaele, 2010; Nunez et al., 2011). Furthermore,the establishment of introduced species appears to be conduciveto the establishment of additional introduced species (Kulmatiski,2006). Exacerbating these concerns is global change itself, withanthropogenic climate change and associated changes to distur-bance regimes projected to further enhance establishment oppor-tunities for introduced species (Higgins and Richardson, 1998;Richardson and Rejmanek, 2004). As forestry is the second mostcommon source of introduced woody plant species worldwide(Richardson and Rejmanek, 2011), and it is those introduced treespecies already naturalized through forestry operations that are aregular source of future invasives (Richardson and Rejmanek,2004), the effective management of introduced tree species nowrequires assessments of how to make regions more resilient to ad-verse ecological outcomes. This may involve the identification ofareas of highest risk of invasion, as well as those in need of invasivemanagement (Rouget et al., 2002), and the targeted restoration of

biodiverse native ecosystems within the landscape matrix, to actas competitive barriers to the recruitment and spread of intro-duced species (Bakker and Wilson, 2004; Donald and Evans,2006; Hulme, 2006).

To conclude, the ecological risks associated with the four treespecies considered are diverse, often hampered by uncertainty,with outcomes that range from compensatory (the potential biodi-versity benefits of the invasive sycamore maple) to synergistic (e.g.invasiveness/hybridization risks in Hybrid aspen). Whereas weemphasize that there is an inherent subjectivity to the evaluationof these risks; by evaluating the available evidence we can never-theless help to identify mitigation priorities and managementstrategies, determine which ecological risks may be irreversible,and identify knowledge gaps in need of further research. By sodoing stakeholders can begin the process of informed decisionmaking regarding the effective management of introduced treespecies and their associated risks and benefits.

Acknowledgements

The research was funded through Future Forests, a multi-disci-plinary research program supported by the Foundation for Strate-gic Environmental Research (MISTRA). The funders had no role instudy design, data collection and analysis, decision to publish, orpreparation of the manuscript. We thank two anonymous review-ers for their constructive comments which improved the finalversion of this paper, and Karin Eklund and Åke Lindelöw forassistance with data collection.

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