This is the published version of the abstract: Available from

Deakin Research Online This is the published version of the abstract: Donaldson, Angie and Bennett, Andrew 2004, Ecological effects of roads : implications for the internal fragmentation of Australian parks and reserves, Parks Victoria, Melbourne, Vic Available from Deakin Research Online: http://hdl.handle.net/10536/DRO/DU:30010280 Reproduced with the kind permission of the copyright owner. Copyright : 2004, Parks Victoria

P A R K S V I C T O R I A T E C H N I C A L S E R I E S

NUMBER 12

Ecological Effects of RoadsImplications for the internal fragmentation

of Australian parks and reserves

Authors: A. Donaldson & A. Bennett

June 2004

© Parks Victoria

All rights reserved. This document is subject to the Copyright Act 1968, no part of this publication may be reproduced, stored in a retrieval system, or transmitted in any form, or by any means, electronic, mechanical, photocopying or otherwise without the prior permission of the publisher.

First published 2004

Published by Parks VictoriaLevel 10, 535 Bourke Street, Melbourne Victoria 3000

Opinions expressed by the Authors of this publication are not necessarily those of Parks Victoria, unless expressly stated. Parks Victoria and all persons involved in the preparation and distribution of this publication do not accept any responsibility for the accuracy of any of the opinions or information contained in the publication.

Authors:

Angie Donaldson – School of Ecology and Environment, Deakin UniversityAndrew Bennett – School of Ecology and Environment, Deakin University

National Library of AustraliaCataloguing-in-publication data

Includes bibliography.ISSN 1448-4935

Citation:

Donaldson A. & Bennett A. (2004) Ecological effects of roads: Implications for the internal fragmentation of Australian parks and reserves. Parks Victoria Technical Series No. 12. Parks Victoria, Melbourne.

Printed on environmentally friendly paper

Parks Victoria Technical Paper Series No. 12

Ecological effects of roads:

Implications for the internal fragmentation of Australian parks and

reserves

Angie Donaldson

Andrew Bennett

Deakin University

June 2004

Parks Victoria Technical Series No. 12 Ecological Effects of Roads

II

Executive Summary Roads are a widespread and common feature of landscapes throughout the world. The

primary function of roads is to facilitate the transport of people between locations, and to

provide access to otherwise remote areas. The total area of land directly covered by roads is

substantial, but their environmental impacts on surrounding landscapes extend their potential

effects even further.

Five primary impacts of roads on ecosystem function and biodiversity have been identified by

a review of published literature on the ecological effects of roads and road traffic. These

ecological impacts also result, often to a lesser degree, from the use of recreational tracks by

hikers and horseriders. The review focused primarily on literature relating to roads in parks,

reserves, and large continuous areas of vegetation, rather than roads in agricultural and

urban areas.

1 Roads are a source of habitat alteration

Roads alter surrounding habitats in numerous ways, consequently influencing the quality and

suitability of roadside areas for plants and animals. Increased edge effects, increased levels

of disturbance, and greater input of matter and energy are the primary ways that roads alter

conditions in adjacent habitats. Potential effects of these impacts on animal populations

include local reductions in population density, altered reproduction and mortality rates, and

altered movement and dispersal patterns. For vegetation, the potential effects of these

impacts include altered productivity, structure, and floristic composition of vegetation

communities. Relatively little is known about the distance to which these effects extend from

Australian roads; however, research has indicated that they extend further for animals than

for plants.

2 Roads provide conduits for the movement of plants and animals

Roads serve as a conduit for the movement of plants and animals. Human use of roads has

many impacts on surrounding environments, including the spread of human disturbance and

hunting pressure, often in otherwise remote areas. Many other organisms also use roads and

roadside habitats as movement corridors, including animals inhabiting road verges and

associated edge habitats, and predators or scavengers moving along the road itself. The

dispersal of weed propagules, and other alien flora, along roads is another way in which

roads can act as conduits. Other organisms such as the pathogen Phytophthora cinnamomi

are also spread via roads and road making activities, often with serious impacts on

surrounding vegetation. The movement of these organisms along roads and their

subsequent invasion of roadside habitats alters the habitat characteristics and composition of


III

roadside communities. The movement of pest species (including weeds, introduced animals

such as foxes and cane toads, and Phytophthora cinnamomi) along roads in Australia has

been identified as a problem in some areas. While little-used roads and tracks in parks and

reserves have reduced likelihood of facilitating the movement of some of these organisms,

such undisturbed areas are particularly susceptible to the potential impacts arising from the

movement of pest species.

3 Roads act as barriers to the movement of animals, potentially fragmenting and isolating populations and communities

Roads may form barriers to the movement of animals. This barrier effect results in one of the

more significant ecological impacts of roads: the fragmentation and isolation of wildlife

populations. Roads may limit the access of animals to vital resources, therefore decreasing

the area of available habitat, and may potentially limit the movement and dispersal of

individuals, fragmenting populations and consequently reducing gene flow. The barrier effect

of roads on animal movement depends primarily on road width and the intensity of its use.

Wide roads with heavy traffic loads have the greatest impact on animal movement. Tracks in

parks and reserves are likely to have less impact on animal movements.

4 Roads cause wildlife mortality

The mortality of wildlife on roads may act as a significant demographic ìsinkî for some

populations and species. Many animals are vulnerable to being killed on roads, with the

associated impact on populations differing between species. Most species are not adversely

affected by the impacts of road mortality, but for some, road mortality can be a significant

threat to the survival of populations.

To reduce the barrier effect of roads on the movement of animals due to both movement

inhibition and increased mortality, structures facilitating animal movement across roads such

as culverts and wildlife underpasses have been used. The success of such structures in

facilitating animal road crossings is primarily related to their dimensions and location.

However, little is known of the benefits of such movements for the overall status of

populations divided by roads; movement of few individuals across such structures, for

example, may not greatly benefit a population divided by a long stretch of road.

5 Roads are a source of biotic and abiotic effects

The impacts of roads on plants and animals, as well as other components of the environment

such as soil and water, are derived both from their structural characteristics and their use

and maintenance. Roads are a source of numerous particulate and chemical pollutants and

traffic-related disturbance. Water run-off from road surfaces alters the moisture balance of

roadside areas, and potentially affects the local hydrology by altering natural flow regimes.


IV

Increased sedimentation of road run-off also alters water flow regimes, and reduces the

quality of aquatic habitats. Roads and their traffic alter soil properties and structure, the

activity of micro and macroinvertebrate soil fauna, and also leaf litter layers. This influences

the structure and floristic composition of roadside vegetation communities. The degree to

which these effects of roads impact on surrounding environments depends on a number of

factors, including road traffic volume and surfacing material, and frequently varies between

locations and through time.

Some progress has been made in theoretically integrating the relative impacts of these

diverse and extensive road effects in a way that allows the ecological impacts of roads to be

understood and potentially quantified. One outcome, which incorporates the more significant

impacts of roads on the ecological functioning of ecosystems, is the concept of a îroad

effectî zone (Forman et al. 1997). It is hypothesised that the ecological ìroad effectî zone

extends for up to 300 m on either side of roads in the USA, and therefore affects a significant

area of land, an area substantially larger than that covered by roads themselves.

The bulk of the research on the environmental impacts of roads has been undertaken in

North US and Europe. Very little data are available on the specific effects of roads on

Australian environments. Most studies also relate to roads that transect urban or agricultural

areas. Less research has reported on the ecological effects of roads in forested or otherwise

undisturbed natural habitats such as national parks and nature reserves. Increased research

into the environmental impacts of roads on Australian systems is required to provide a basis

for managing the effects of roads in parks and reserves.


V

Contents

Executive Summary .............................................................................................................II 1 Introduction ......................................................................................................................1 2 Roads as a source of habitat alteration ..........................................................................2

2.1 HABITAT ALTERATION BY ROADS ñ IMPACTS ON WILDLIFE ............................................... 2 2.2 HABITAT ALTERATION BY ROADS ñ IMPACTS ON VEGETATION ........................................ 9

3 Roads as conduits for the movement of plants and animals ......................................12 3.1 DISPERSAL OF ANIMALS ALONG ROADS ............................................................................. 12 3.2 INCREASED PREDATOR MOVEMENT AND ACTIVITY ALONG ROADS .............................. 13 3.3 DISPERSAL OF PLANTS ALONG ROADS ............................................................................... 15 3.4 MOVEMENT OF A PATHOGEN, PHYTOPHTHORA CINNAMOMI, ALONG ROADS ............. 17 3.5 MOVEMENT OF HUMANS ALONG ROADS: IMPLICATIONS OF INCREASED ACCESS TO

NATURAL ECOSYSTEMS......................................................................................................... 18 4 Roads as barriers to the movement animals, potentially fragmenting and

isolating populations and communities.....................................................................19 5 Roads as sinks ñ wildlife mortality................................................................................26 6 Mechanisms that potentially reduce the barrier effect of roads..................................32 7 Roads as a source of biotic and abiotic effects ...........................................................36

7.1 POLLUTION FROM ROADS AND MOTOR VEHICLES............................................................ 36 7.2 IMPACTS OF ROADS ON HYDROLOGICAL SYSTEMS.......................................................... 39 7.3 IMPACTS OF ROADS ON SOIL CHARACTERISTICS ............................................................. 43

8 Road systems and their ecological impacts on landscapes ñ the ì road effectî zone ..............................................................................................................................46

9 Ecological impacts of recreational tracks ....................................................................49 References..........................................................................................................................53


1

1 Introduction Roads are a widespread and common landscape feature, present in even the most remote

areas, and are entirely anthropogenic in origin. The primary function of roads is to facilitate

the transport of people between locations, and to provide access to otherwise remote areas.

In doing so, roads create artificial corridors in landscapes, and as such perform a number of

functions that are characteristic of both natural and human-induced corridors. These

functions are habitat, conduit, filter, source, and sink (Forman 1995). The ecological impacts

of roads are related to the way roads perform these functions, and thereby modify the natural

processes, flows, and interactions in landscapes. These ecological impacts also result, often

to a lesser degree, from recreational tracks (Liddle 1997). Due to the influence of roads on a

range of biotic and abiotic processes, as well as on landscape structure and function, it has

been suggested that they should be considered an inherent component of landscape

structure (Miller et al. 1996; Saunders et al. 2002).

Roads in national parks and other such reserves serve four primary functions, including

access for fire suppression, resource use (such as timber harvesting), recreation, and land

management (Department of Conservation and Environment 1992). The type, size, and

average traffic volume of these roads varies widely: ranging from major, sealed access roads

to disused fire suppression tracks (Department of Conservation and Environment 1992). All

of them, however, affect the surrounding environment in many ways, from initial road

construction to subsequent vehicle use.

The management of areas reserved for the conservation of native biodiversity usually aims to

minimise the impact of anthropogenic disturbance. To this end, the management plans of

many Victorian national parks include provisions for the closure of redundant roads

(Department of Conservation and Environment 1992). An understanding of the potential

ecological effects of roads will aid in the process of identifying roads and tracks to be closed.

The primary aim of this review of literature on the ecological effects of roads is to identify

ways that roads potentially increase the internal fragmentation of parks and reserves. Most

research on this topic, however, has been undertaken in agricultural and urban landscapes.

Relatively few studies have focused on the impact of roads in forested or other undisturbed

habitats. Furthermore, most published research has been undertaken in Europe and North

America. Very little data on the specific effects of roads on Australian environments and

biota, especially in largely undisturbed areas, are available. This review focuses primarily on

Australian research where possible, however in many cases the findings relate to overseas

systems. The research cited in this review is comprehensive but does not necessarily

constitute the entire body of work on this topic.


2

2 Roads as a source of habitat alteration

2.1 HABITAT ALTERATION BY ROADS ñ IMPACTS ON WILDLIFE

Extent of current knowledge

Roads alter surrounding habitats in numerous ways, consequently affecting the quality and

suitability of roadside areas for wildlife habitat. Increased edge effects, disturbance, and the

input of matter and energy are the primary ways roads alter conditions in adjacent habitats.

The result of these impacts on fauna populations differs between species; however, a local

reduction in density is the most common outcome (Ferris 1979; Dhindsa et al. 1988; Reijnen

et al. 1997; Kuitunen et al. 1998; Baker et al. 1998; Huijser & Bergers 2000). Reproduction

and mortality (discussed in Section 5) rates are also commonly altered in roadside animal

populations (Reijnen & Foppen 1994; Reijnen et al. 1995; Reijnen et al. 1997; Ortega &

Capen 1999). Animals may also exhibit altered movement and dispersal patterns (Clarke et

Research into the effects of roads on fauna generally focuses on a particular species, or type of animal. Collectively they address a wide variety of potential impacts derived from roads. The outline of the research reviewed summarises studies based on the type of animal investigated while the following review discusses the findings in relation to the resultant impacts on fauna. This summary of relevant research does not include studies that are more appropriate to other sections of the review, for example road mortality of fauna.

The effect of forest roads on fauna has been broadly researched in Australia (Goosem 1997), the United States (US) (Burke & Sherburne 1982; Sherburne 1983), and South Africa (Pienaar 1968). Studies have researched roadside invertebrate populations in Australia (van Schagen et al. 1992), the US (Stiles & Jones 1998), the United Kingdom (UK) (Port & Thompson 1980; Spencer & Port 1988), and Germany (Maurer 1974). Amphibians have been studied in Australia (Seabrook & Dettmann 1996) and Canada (Fahrig et al. 1995), and reptiles in the US (Rudolph et al. 1999). The impact of roads on birds has been studied in Australia (Anonymous 1980; Brown et al. 1986; Lepschi 1992), India (Dhindsa et al. 1988), Scotland (Keller 1989), Czechoslovakia (Havlin 1987), Finland (Kuitunen et al. 1998), US (Stahlecker 1978; Ferris 1979; Whitford 1985; Rich et al. 1994; Knight et al. 1995; Ortega & Capen 1999), and The Netherlands (Reijnen & Foppen 1994; Foppen & Reijnen 1994; Reijnen et al. 1995; Reijnen et al. 1997). Research on the impact of roads on small mammals has been undertaken in Australia (Preuss 1989; Goosem 2000), the US (Huey 1941; Adams & Geis 1983), and The Netherlands (Huijser & Bergers 2000). Roadside populations of large mammals have been documented in the UK (Clarke et al. 1998), Norway (Nellemann et al. 2001), Alaska (Thurber et al. 1994), Canada (McLellan & Shackleton 1988), and the US (Carbaugh et al. 1975; Rost & Bailey 1979; Thiel 1985; Mech et al. 1988; Brody & Pelton 1989). In Australia, the impact of powerline easements on birds (Baker et al. 1998) and small mammals (Goldingay & Whelan 1997) has been studied.


3

al. 1998). Species distributions may change due to the density and spatial arrangement of

roads within a landscape, consequently influencing community composition and species

interactions (Thiel 1985; Thurber et al. 1994).

Edge effects of roads

Edge effects are generally associated with the interactions that occur between two adjacent,

but different, ecosystems or habitats (Murcia 1995; Lidicker 1999). A habitat edge exists

when habitat features as perceived by an individual or species are disrupted, therefore

affecting the performance of the organism(s) in some way (Lidicker 1999). The creation of

roads in areas of otherwise continuous habitat causes many of the disruptions characteristic

of edges.

The distance to which different biotic and abiotic factors penetrate across a habitat edge is

highly variable, and is influenced by the permeability of the habitat edge (Lidicker 1999).

Some of the abiotic factors that vary across edges include light, moisture and temperature of

air and soil, wind velocity, and chemical substances (Murcia 1995). Biotic factors influenced

by habitat edges include the abundance and distribution of species, with consequent impacts

on interactions such as predation, competition, pollination, and seed dispersal (Murcia 1995).

Habitat edges are frequently characterised by one or many of the following features:

increased productivity and diversity, the presence of edge-adapted species, the

establishment and spread of exotic species, altered community composition, and increased

vulnerability of interior species (Murcia 1995; Lidicker 1999). Roadside habitats often exhibit

these features.

Impact of roads on habitat quality

Increased noise disturbance from road traffic can reduce the quality of roadside habitats for

fauna, particularly birds (Reijnen & Foppen 1994; Reijnen et al. 1995; Reijnen et al. 1997).

The interference of traffic noise with bird communication during incubation and fledging

stages of reproduction is the main impact of road noise on bird communities (Forman &

Deblinger 2000). Research has shown that noise is the primary factor causing reduced

densities of breeding birds in roadside habitats in The Netherlands (Reijnen & Foppen 1994;

Reijnen et al. 1995; Reijnen et al. 1997).

The impact of roads on quality of aquatic habitats is discussed in Section 7.


4

Resources provided by roads

Resources provided by roads include food, additional habitat, and heat energy (Huey 1941;

Carbaugh et al. 1975; Stahlecker 1978; Whitford 1985; Dhindsa et al. 1988; Mader et al.

1990; Rudolph et al. 1999). Increased predator access to prey along roads also increases

roadside resources for some predators (see Section 3).

Insects, grit, grain, and carrion are resources often available in increased quantities along

roads. Spilled grain from the transport of agricultural produce provides short-term, localised,

and often highly abundant food for birds (Anonymous 1980; Dhindsa et al. 1988). However,

birds feeding on grain spillages are often at high risk of vehicle mortality. Birds use road grit

in their gizzards, including in Australia western rosellas (Platycercus icterotis) and common

bronzewings (Phaps chalcoptera) (Brown et al. 1986). Road-killed animals provide food for

scavengers, such as ravens (Corvus coronoides) in Australia (Lepschi 1992). Birds also feed

on insects along roads, both ground-dwelling invertebrates and those feeding, often in high

abundances, on roadside vegetation (Brown et al. 1986).

Road maintenance in many areas involves mowing, slashing, or herbicide application to

ensure that road verges contain low, grassy vegetation (Scheidt 1971). Road verges

consequently provide food for many herbivorous animals. For example, eastern grey

kangaroos (Macropus giganteus) and wombats (Vombatus ursinus) feed in roadside habitats

in Australia, as do white-tailed deer (Odocoileus virginianus) in the US (Carbaugh et al. 1975;

Bennett 1991). Likewise, small grassland mammals, including grassland melomys (Melomys

burtoni) and canefield rats (Rattus sordidus), are favoured by grassy road verges and

powerline easements in Australia (Goosem & Marsh 1997; Goosem 2000).

Roads provide habitat or favourable habitat features for many animals. In Australia, frogs

often spawn in rainforest road drains and also in the film of water that may occur on paved

roads (Goosem 1997). In Germany, lizards favour the habitat provided by gravel road

foundations (Mader et al. 1990). The creation of favourable habitat along American desert

roads is considered to be responsible for the significant range expansion of pocket gophers

(Thomomys) (Huey 1941). The distribution of raptors in the US is influenced by the presence

of powerlines due to their preferential use as perching sites (Stahlecker 1978).

Both birds and snakes may utilise the heat released from road surfaces (Whitford 1985;

Rudolph et al. 1999). Snakes are known to thermoregulate on road surfaces (Rudolph et al.

1999). Whitford (1985) determined that the use of heat from roads by small birds in the US

was a behavioural adaptation allowing the conservation of metabolic energy, potentially

increasing reproductive success and breeding range.


5

Impact of roads on the density of animal populations and species richness of communities

Due to habitat alteration, the density of animal populations and species richness of

communities is often altered and primarily reduced in roadside habitats (Dhindsa et al. 1988;

Reijnen et al. 1997; Baker et al. 1998). Species that favour edges are the common exception

to this trend, and usually occur at higher density and diversity in roadside habitats (Ferris

1979; Sherburne 1983). The presence of edge species in roadside habitats often maintains a

high species richness and abundance in roaded areas (Ferris 1979; Sherburne 1983).

The abundance of forest bird species and richness of communities in a powerline easement

in New South Wales, Australia was less than 20% of the values recorded in adjacent

forested areas (Baker et al. 1998). Significant reductions in the abundance and species

richness of forest birds were recorded for up to 125 m from the edge of the powerline

easement (Baker et al. 1998).

Densities of land, forest, and open-habitat bird species in roadside habitats comparative to

adjacent habitats were found to be reduced in Finland, the US, and The Netherlands,

respectively (Rich et al. 1994; Reijnen et al. 1997; Kuitunen et al. 1998). Species richness,

diversity, and evenness of bird communities are also decreased along Indian roads, with

negative correlations existing between these values and road width (Dhindsa et al. 1988).

Contrary to the majority of research, Havlin (1987) recorded increased bird densities along

Czechoslovakian roads.

The distance to which road disturbance may influence bird density varies with species (with

grassland species being more highly affected than woodland birds), traffic volume (higher

volumes having greater impacts), and adjacent habitat type (Reijnen et al. 1997). In The

Netherlands, the population densities of all bird species within 1.5 km of roads were reduced

by at least 30% (Reijnen et al. 1997). Ferris (1979) recorded a 50% reduction in bird

populations inhabiting roadside areas along a four-lane road in the US.

Hedgehog (Erinaceus europaeus) populations in The Netherlands were reduced by 30% by

roads and traffic, with an associated decrease in the survival potential of local populations

(Huijser & Bergers 2000). Likewise, Sherburne (1983) recorded reduced densities of red-

backed voles (Clethrionomys gapperi) inhabiting roadside habitats in the US. However, the

density and richness of small mammal communities inhabiting American roadside areas were

greater than those in adjacent habitats, primarily due to an increased diversity and

abundance of grassland species (Adams & Geis 1983).

The density of reptiles and amphibians may also decrease in areas close to roads (Fahrig et

al. 1995; Rudolph et al. 1999). In the US, populations of large snakes were reduced by up to


6

50% within 450 m of moderately used roads, with lower population densities recorded up to

850 m (Rudolph et al. 1999). While traffic volume had little impact on snake density (Rudolph

et al. 1999), it was found to influence the density of frogs and toads in roadside habitats in

Canada (Fahrig et al. 1995). In Australia, both the abundance and activity of cane toads

(Bufo marinus) is increased along roads (Seabrook & Dettmann 1996).

The density and richness of beetle and spider populations was found to decrease close to

roads in Germany (Maurer 1974). However, populations of other invertebrates, particularly

herbivorous species, have been found to inhabit roadside areas in significantly increased

densities (Port & Thompson 1980; van Schagen et al. 1992; Stiles & Jones 1998). In

roadside vegetation in Australia and the UK, increased populations of defoliating insects

have been related to the high nitrogen content of roadside soil (Port & Thompson 1980; van

Schagen et al. 1992). Traffic density modifies the impact of roads on invertebrate population

densities (Maurer 1974).

Impact of roads on community composition

The composition of faunal communities in roadside habitats is often different to those

occurring further away. In the US, the composition of bird communities varied with distance

from the road, with edge species being more diverse closer to the road (Ferris 1979). Similar

findings of differing small mammal composition in American roadside habitats compared to

adjacent areas indicated that grassland species were more common along roads (Sherburne

1983). Sherburne (1983) found that with increasing time since construction, the number of

edge species along roads increased while the number of forest species decreased.

Conversely, Kuitunen et al. (1998) did not record any differences in the composition of bird

communities sampled close (25 m) to roads in Finland and those further away (200 m)

In Queensland, Australia, the composition of small mammal communities was altered by the

edge effects of an unsealed rainforest road (Goosem 2000). Fawn-footed melomys

(Melomys cervinipes) were more abundant in roadside habitats than edge-avoiding Rattus

species, which were more abundant further from the road. This was related to the width of

the clearing and the grassy road verge habitat (Goosem 2000). The intrusion of grassland

species such as grassland melomys (Melomys burtoni) and canefield rats (Rattus sordidus)

along roadside habitats was also observed in this study, though these species had not

penetrated the adjacent rainforest habitat (Goosem 2000).

Avoidance of roads by animals

Invertebrates, birds, and small mammals frequently have differing population densities and

community structure in roadside habitats. However, larger vertebrates more commonly


7

respond to roads by altering their movement and behaviour rather than undergoing changes

to population and community structure.

Even minimal traffic volumes were found to displace Grizzly bears (Ursus arctos Ord) within

100 m of roads in Canada (McLellan & Shackleton 1988). The avoidance of roadside

habitats was influenced by the sex and age of animals, and this behaviour was calculated to

reduce the area of habitat available to grizzly bears (Ursus arctos) by 8.7% (McLellan &

Shackleton 1988).

In the US, mule deer (Odocoileus spp.) and elk (Cervus canadensis) are known to avoid

roads, with populations of both species being significantly reduced within 200 m of roads

(Rost & Bailey 1979). However, road avoidance in these species was lower when

surrounding habitats were reduced, or resources were limited (Rost & Bailey 1979).

The avoidance of roads and powerlines by wild reindeer (Rangifer tarandus tarandus) in

Norway was evident for up to 2.5 km, causing a reduction in population density of 79% in

these areas (Nellemann et al. 2001). Nellemann et al. (2001) concluded that the avoidance

of roads and powerlines by wild reindeer affects the distribution of species and reduces the

availability of winter habitat, which therefore reduces the carrying capacity of the mountain

ranges.

The dispersal and colonisation of badgers (Meles meles) in England is restricted by roads

(Clarke et al. 1998). Consequently, an increased density of roads in England poses a threat

to the potential increase in badger populations in the future (Clarke et al. 1998).

Pink-footed (Anser brachyrhynchus) and greylag (A. anser) geese avoid roads in agricultural

land in Scotland, with avoidance behaviour recorded for distances of 100 m from roads

(Keller 1989).

Impact of roads on the distribution of territories and species

Road density in a given area impacts the persistence and distribution of some animals,

usually large vertebrates (Mech et al. 1988; Brody & Pelton 1989; Thurber et al. 1994). For

example, Mech et al. (1988) determined that road densities of 0.58 km per km≤ in Minnesota,

US formed the threshold for the occurrence of gray wolves (Canis lupus) in an area. The

spatial organisation of wolf packs is influenced by the pattern of road occurrence within

landscapes; with increased road density, the likelihood of wolf persistence is reduced (Thiel

1985; Thurber et al. 1994).

Black bears (Ursus americanus) inhabiting Pisgah National Forest, US responded to roads

by shifting the location of home ranges rather than altering their movement patterns within


8

pre-established home ranges (Brody & Pelton 1989). Implications exist for alterations to the

interspecific interactions between black bears.

The density of ovenbird (Seiurus aurocapillus) territories in American forests was decreased

by up to 40% with increasing proximity to roads (Ortega & Capen 1999).

In the Central Highlands of Victoria, Australia, road widening for forestry and fire suppression

is considered a threat to the survival of the Leadbeaters possum (Gymnobelideus

leadbeateri) (Preuss 1989). The removal of hollow bearing trees to increase road widths will

decrease the habitat available for the persistence of this species (Preuss 1989).

Impact of roads on reproductive success

Ovenbirds inhabiting American forests exhibited lower pairing success in habitats within

150 m of roads (Ortega & Capen 1999). Likewise, pairing success for yearling male willow

warblers (Phylloscopus trochilus) in roadside habitats in The Netherlands was half that

recorded for males in other areas (Reijnen & Foppen 1994). The proportion of yearling males

in roadside habitats was significantly increased compared to habitats further from roads,

possibly due to reduced habitat quality, leading to the conclusion that roadside habitats were

a demographic îsinkî for males emigrating from better quality habitats further from the roads

(Reijnen & Foppen 1994). However, yearlings dispersing from roadside habitats were found

to disperse greater distances away from the road, thereby stabilising source populations

(Foppen & Reijnen 1994).

Relevance to roads in national parks

The most important issue arising from the impacts of habitat alteration by roads on wildlife is

the effect on animal persistence in roaded areas (Fahrig 2001). This is of particular

importance for species that are likely to be negatively affected by the occurrence of roads,

and therefore have decreased potential for maintaining populations in roadside habitats.

Species already vulnerable to extinction will be of particular concern. However, it is also

important to consider those species that are able inhabit roadside areas in higher than

normal densities. For example, predators or successful competitors that are either introduced

or native will require the most consideration[a1].

The degree to which these impacts influence the potential persistence of fauna populations

in natural habitats is important. The relative effect of each impact will differ for each type,

size, and species of animal, as well as for different individuals.


9

2.2 HABITAT ALTERATION BY ROADS ñ IMPACTS ON VEGETATION


Changes in the structure and floristic composition of vegetation communities at forest edges

reflect the altered abiotic conditions caused by the edge (Williams-Linera 1990). Changed

abiotic conditions are primarily caused by the removal of canopy vegetation, thereby altering

microclimatic conditions by increasing solar radiation and decreasing relative humidity (due

to increased temperatures and wind velocity and decreased transpiration; Williams-Linera

1990). Most research studying the impacts of habitat edges on vegetation structure and

floristic composition relates to the distinct edge between forest habitats and cleared land.

Few studies have looked at the impact of road edges on adjacent vegetation. This is an

important distinction because the clearing, with related edge effects, associated with roads is

unlikely to be as severe as that resulting from large-scale vegetation clearing. However,

other factors are associated with roads, including the effect on soil properties of road run-off

and the pollutants and particles that run-off contains (see Section 7).

Habitat alteration may affect plant populations in several ways. Resources available for

flowering and fruiting may be altered by changed abiotic conditions, predation of plant

reproductive structures may be affected by altered animal communities, and pollination may

be altered due to changes to pollinators and altered wind patterns (Cunningham 2000).

Impact of roads on vegetation productivity

Vegetation productivity is often increased in roadside areas (Johnson et al. 1975; Lamont &

Southall 1982; Spencer & Port 1988; Lamont et al. 1994a, 1994b). In the UK, plants grown in

roadside soil had higher growth rates than those growing in soil taken further from roads

(Spencer & Port 1988). Johnson et al. (1975) reported that the productivity of vegetation

adjacent to paved roads was 17 times, and unpaved roads 6 times, greater than that of

surrounding desert habitats in the US. However, it is hypothesized that there is an upper limit

to the width of roads beyond which these impacts will not increase [a2](Johnson et al. 1975).

Research into the effects of roads on adjacent vegetation communities has primarily studied the productivity and species composition of roadside communities. Increased productivity of roadside vegetation has been reported in Australian studies (Lamont et al. 1994a, 1994b), and in the US (Johnson et al. 1975) and the UK (Spencer & Port 1988). Changes to the floristic composition of roadside vegetation has been addressed by three Australian studies (Lamont & Southall 1982; Reid 1997; Norton & Stafford Smith 1999), as well as work in Panama (Williams-Linera 1990), the UK (Angold 1997), and Puerto Rico (Olander et al. 1998). The impact of herbivory on roadside vegetation was investigated by an Australian (van Schagen et al. 1992), and an English study (Port & Thompson 1980).


10

The productivity of particular species may also increase in roadside habitats. For example,

Menzies' banksia (Banksia menziesii) and Hookers' banksia (B. hookeriana) exhibited

increased size and fecundity along roads in Western Australia (Lamont et al. 1994a, 1994b).

Crown size and the number of flower heads of roadside banksia trees were both found to be

two and a half times greater than trees growing 50 m from the road (Lamont et al. 1994a,

1994b). Seed store in roadside plants was almost five times greater due to increased cone

production, follicles per cone, and seed viability per follicle (Lamont et al. 1994a, 1994b).

Increased fecundity was related to the increased access of roadside plants to moisture and

nutrients from road run-off, and reduced competition from other plants. Lamont et al. (1994a,

1994b) suggested that the health of banksia populations in roadside areas may provide a

buffer against local decline and extinction of these species.

Impact of roads on floristic composition and structure of vegetation communities

In the UK, Angold (1997) recorded increased growth of vascular plants, particularly grass

species, and decreased lichen abundance in heathland vegetation within 200 m of major

roads. The extent of the edge effect was positively correlated with traffic density and road

width (Angold 1997). The enhanced growth of vascular plants was related to the increased

nitrogen available in roadside areas, and the eutrophication of these areas increased the

competitive advantage of grass species (Angold 1997).

In Puerto Rico, roads in tropical forests affected the composition of adjacent vegetation for a

distance of 10 m (Olander et al. 1998). Monocots were more abundant along roads whereas

trees were more abundant in areas of mature forest. Olander et al. (1998) also recorded an

increased abundance of introduced species within 2 m of the road, a finding related to the

altered soil properties adjacent to roads.

In Australia, research into the floristic composition of roadside vegetation has focused on the

increased abundance of mistletoes (Loranthaceae) in roadside areas (Lamont & Southall

1982; Reid 1997; Norton & Stafford Smith 1999). While these studies have been conducted

in linear strips of remnant vegetation along roads, as opposed to continuous vegetation

fragmented by roads, they indicate that mistletoes are denser in roadside habitats than in

vegetation further from roads (Lamont & Southall 1982; Norton & Stafford Smith 1999).

Lamont and Southall (1982) recorded 13 times more mistletoes on potential host trees in

road verges than in an adjacent nature reserve in Western Australia. Increased water and

nutrient availability, reduced roadside perching sites, and increased movement of birds along

roadside corridors increased mistletoe occurrence in roadside vegetation (Norton & Stafford

Smith 1999). The redistribution of resources by roads, rather than the impact of the road

itself, is considered to the most influential factor in the pattern of mistletoe occurrence


11

(Norton & Stafford Smith 1999). However, Reid (1997) proposed that a reduction in roadside

habitats of two factors known to regulate mistletoe populations under natural conditions, fire

frequency and herbivorous marsupial abundance, may increase the abundance of roadside

mistletoes.

Defoliation of roadside vegetation due to increased density of herbivorous insects has been

related to the impacts of nitrogen pollution arising from vehicle emissions (Port & Thompson

1980; van Schagen et al. 1992). In Western Australia, van Schagen et al. (1992) reported

that defoliation of roadside vegetation by caterpillars (Ochrogaster lunifer) resulted in leaf

loss and early shoot formation, leading to roadside trees with younger foliage. Port and

Thompson (1980) found that the increased nitrogen content of roadside vegetation in the UK,

in conjunction with the stressed condition of roadside plants, potentially caused greater

populations of herbivorous insects. Angold (1997) recorded increased plant damage from

herbivore attack in roadside heath species in the UK.


From the limited data available it is apparent that the impacts of roads on adjacent vegetation

communities are diverse and potentially significant. Alterations to the floristic and structural

composition of roadside vegetation communities will also affect the habitat available for

fauna, as well as ecosystem processes in roadside areas. More research needs to focus on

the edge effects associated with roads in otherwise undisturbed vegetation.


12

3 Roads as conduits for the movement of plants and animals One primary function of roads is the facilitation of the movement of people and goods

between locations. However, roads also serve as conduits for the movement of other biota,

thereby potentially affecting surrounding environments. In particular, the potential invasion of

habitats by non-indigenous species threatens native biodiversity by changing the

characteristics of habitats as well as by altering species composition (Greenberg et al. 1997).

3.1 DISPERSAL OF ANIMALS ALONG ROADS


Roads can facilitate the dispersal and range expansion of animals in two main ways. Habitat

alteration in road verges may allow the movement of species that favour altered roadside

habitats, or animals may move directly along the road surface. Animals may also be

transported into previously uninhabited areas by vehicles. However, conditions in roadside

habitats must suit the requirements of the moving, or translocated, animals for populations of

these species to establish in roadside areas.

In Australia, the creation of cleared, grassy habitats in powerline easements was found to

facilitate the intrusion of small mammals that did not occur in the natural vegetation on either

side of the easement (Goosem & Marsh 1997). However, while three species, grassland

melomys (Melomys burtoni), canefield rats (Rattus sordidus), and introduced house mice

(Mus domesticus), were present in the easement, none had invaded the adjacent forest

vegetation (Goosem & Marsh 1997).

Cane toads (Bufo marinus) use roads and tracks in Australia as activity and dispersal

corridors (Seabrook & Dettmann 1996). The potential for movement of cane toads along

roads is increased in areas where the adjacent roadside vegetation is dense (Seabrook &

Dettmann 1996).

Transportation via movable homes facilitated the establishment of a spotted marsh frog

(Limnodynastes tasmaniensis) population in Kununurra, Western Australia almost 2000 km

The potential of animals to extend their distribution along roads has been recorded for small mammals and anurans. Australian studies report range expansion along roads by frogs (Martin & Tyler 1978) and toads (Seabrook & Dettmann 1996). The range expansion of small mammals has been studied along powerline easements in Australia (Goosem & Marsh 1997) and roads in the US (Huey 1941; Getz et al. 1978).


13

outside the natural range of the species (Martin & Tyler 1978). The species was recorded to

have extended its new roadside range lengthwise by almost 7 km over a year.

In the US, two species of small mammal have used roadside habitats as routes for dispersal,

with resultant range expansion (Huey 1941; Getz et al. 1978). Huey (1941) recorded the

expansion of the natural range of pocket gophers along desert roads. Increased moisture

and vegetation growth along roads provided appropriate habitat for these animals, therefore

facilitating their movement into otherwise unsuitable areas (Huey 1941). Similarly, habitat

alteration of roadside areas facilitated the movement of the meadow vole (Microtus

pennysylvanicus) into previously uninhabited areas (Getz et al. 1978).


The potential impact on natural communities of faunal range expansion along roads will be

determined by any resulting interactions in the invaded habitats. For example, cane toads

have significant negative impacts on native Australian wildlife, and consequently the range

expansion of this species poses a serious threat to the conservation of Australian fauna

(Seabrook & Dettmann 1996). However, the example of the isolated, localised population of

a non-indigenous frog species in Western Australia is less likely to threaten the survival of

other species.

The range expansion of species known to competitively exclude indigenous species is also

cause for concern. Some of the generalist small mammal species adapted to roadside

vegetation may outcompete native species, with potential repercussions for their long-term

survival (Goosem & Marsh 1997).

3.2 INCREASED PREDATOR MOVEMENT AND ACTIVITY ALONG ROADS


Roads are commonly thought to facilitate the movement of feral predators into natural

habitats, potentially increasing the predation of native fauna (May & Norton 1996, and

references within). In Australia, foxes (Vulpes vulpes) use roads as movement paths (Catling

& Burt 1995). Claridge (1998) found that the localised disturbance of dense vegetation along

Two aspects of the relationship between roads and predation have been studied. The use of roads as movement corridors by vertebrate predators has been reported in three Australian studies (Catling & Burt 1995; May & Norton 1996; Claridge 1998), and one from New Zealand (Alterio 2000). No studies, however, have specifically investigated the use of roads by feral predators (May & Norton 1996). The occurrence of nest predation along roads has been investigated in Australia (Lindenmayer et al. 1999), the US (Small & Hunter 1988; Rich et al. 1994; Keyser et al. 1998), and Mexico (Burkey 1993).


14

trap-lines in Australia facilitated the access of foxes and feral cats (Felis catus), thereby

increasing the foraging range of these predators.

In New Zealand, the use of sodium monoflouroacetate (1080) baits for predator control was

most successful when baits were placed along roads (Alterio 2000). Consequently, the most

cost-effective management of feral predators in New Zealand forests involved bait placement

along roads.

Findings of an association between nest predation and roads are inconsistent. Three studies

did not record increased nest predation along road edges (Small & Hunter 1988; Keyser et

al. 1998; Lindenmayer et al. 1999) while two others did (Burkey 1993; Rich et al. 1994).

In Australia, Lindenmayer et al. (1999) determined that the rate of egg predation from

artificial nests in different types of forest was not related to distance from roads. Two

American studies returned similar results; however, both found that predation rate, as well as

predator occurrence, increased with forest fragmentation and reduced fragment size (Small

& Hunter 1988; Keyser et al. 1998). Keyser et al. (1998) recorded increased activity of large

predators close to the edge of forest fragments, though this activity was not correlated with

increased nest predation.

In the US, forest roads as narrow as 8 m wide attract nest predators to both the corridor

itself, and the adjacent forest interiors (Rich et al. 1994). Birds inhabiting roadside habitats

were at increased risk of predation and brood parasitism (Rich et al. 1994). In Mexico,

Burkey (1993) similarly recorded that egg predation by gray foxes (Urocyon

cinereoargenteus) and coatis (Nasua nasua) was higher within 100 m of rainforest roads.

Such increased nest predation near forest edges may lead to altered bird community

composition in fragmented habitats (Burkey 1993).

In Mexico, the incidence of seed predation was lower with increased proximity to forest roads

(Burkey 1993). This finding was related to edge avoidance by seed predators, and to low

population densities of these species along edges due to the impacts of predation, and may

impact roadside plant recruitment and community structure (Burkey 1993).


Increased access of feral predators, particularly foxes, to undisturbed habitats via roads is

likely to have significant consequences for the conservation of native Australian fauna. Foxes

have been implicated in the extinction, or reduced population densities, of many native

Australian ground-dwelling mammals, especially those within the critical weight range of 35ñ

5,500 g (May & Norton 1996; Claridge 1998). The increased access of this introduced

predator to native prey species in national parks is therefore of great concern.


15

Increased rates of nest predation along roads are considered to be of less concern due to

the inconclusive nature of research findings on this topic.

3.3 DISPERSAL OF PLANTS ALONG ROADS


Roads facilitate the transport of source propagules into otherwise remote areas, by acting as

dispersal corridors and via vehicular dispersal of plants or seeds (Wace 1977; Schmidt 1989;

Wilson et al. 1992; Greenberg et al. 1997). Roadside habitats are also often suitable for plant

establishment, and may contain a reservoir of propagules for future invasion (Parendes &

Jones 2000). Many factors influence the potential establishment of exotic plant species in

roadside habitats, including the level of habitat disturbance, light availability, climatic

conditions, and fire history (Amor & Stevens 1976; Brothers & Spingarn 1992; Cowie &

Werner 1993; Greenberg et al. 1997; Parendes & Jones 2000).

Potential for transport of propagules on vehicles

Cars in Australia and Germany may transport enormous numbers of seeds belonging to a

large number of species (Wace 1977; Schmidt 1989). The potential dispersal of plant

propagules by motor vehicles extends over large distances, and incorporates a wide number

of species exhibiting diverse life cycles, morphologies and dispersal techniques (Wace

1977). In Canberra, Australia, 18,526 seedlings from 259 species germinated from sludge

obtained from an automatic carwash over two and a half years (Wace 1977). In Kakadu

National Park, Australia, 1,960 seedlings belonging to 88 species were collected in a year

(Lonsdale & Lane 1994). The propagules obtained from one car in Germany over almost six

months yielded 3,926 seedlings from 124 species (Schmidt 1989).

Nevertheless, the potential of cars to transport plant propagules does not ensure the

successful germination and survival of these species in roadside habitats (Wace 1977).

Germination success of transported plant propagules has been related to three main factors:

the time of year, the transport position of the propagule on the vehicle, and the provision of

germination cues (Schmidt 1989).

The majority of the literature reviewed in this section relates to research undertaken in Australia. Most reports concern the dispersal of flora via vehicles, and the features of roadside habitats that influence successful establishment (Amor & Stevens 1976; Wace 1977; Cowie & Werner 1993; Lonsdale & Lane 1994). Other such studies have been undertaken in New Zealand (Wilson et al. 1992), Germany (Schmidt 1989), and the US (Tyser & Worley 1992; Greenberg et al. 1997; Parendes & Jones 2000).


16

Schmidt (1989) found that many of the species transported by cars were not common in

roadside vegetation of the study area, indicating the potential for transport of alien species

via motor vehicles. Almost two-thirds of the species obtained from an Australian carwash

were introduced (though not necessarily classed as weeds) and accounted for nearly three

quarters of the total germinating seedlings (Wace 1977). Acknowledged weed species were

one of the most commonly recorded plant groups (Wace 1977).

Lonsdale and Lane (1994) found that most cars entering Kakadu National Park carried either

one or no seeds. However, as weed species carried by cars occurred at three times as many

sites in Kakadu National Park as those not transported by cars, it was considered likely that

weed transport on tourist vehicles inside the park was responsible for weed infestations.

Lonsdale and Lane (1994) highlight the potential for vehicle seed movement to cause subtle

alterations to natural biogeographic boundaries of native species.

Features of roadside habitats favouring establishment of alien flora

Many characteristics of roadside habitats make them particularly susceptible to invasion by

exotic species, including increased substrate disturbance, soil modification, moisture, nutrient

availability, and light (Amor & Stevens 1976; Wace 1977; Tyser & Worley 1992; Cowie &

Werner 1993; Greenberg et al. 1997; Parendes & Jones 2000). The reduced cover of native

vegetation common in roadside habitats due to mowing and herbicide applications also

reduces competition for establishing seedlings (Wace 1977; Greenberg et al. 1997;

Parendes & Jones 2000).

Fire frequency may increase the potential spread of alien species (Amor & Stevens 1976).

However, Cowie and Werner (1993) report that fire occurrence in Kakadu National Park

decreased the invasion of weeds in natural habitats.

Occurrence of plant invasion along roads

The invasion of alien flora has been correlated with distance from road edge. Amor and

Stevens (1976) found that the frequency of alien plants in Australian forests decreased with

distance from roads. Similar findings have been reported for along primary and secondary

roads in grassland communities in Glacier National Park, US (Tyser & Worley 1992). The

frequency of exotic species occurrence along high and low-use roads in the US was likewise

greater than that along abandoned roads and streams (Parendes & Jones 2000)

In Kakadu National Park, Australia, most of the invasive flora of the park was associated with

human activities, roadways and other disturbed areas; however, these areas comprised only

a small proportion of the total park (Cowie & Werner 1993).


17


While cars potentially disperse significant numbers of plant propagules, it is likely that the

majority of these seeds will fail to establish permanent populations in roadside habitats.

Furthermore, the invasion of undisturbed native vegetation by alien species is low (Willson &

Chrome 1989; Brothers & Spingarn 1992). However, the potential for roadside areas to be

reservoirs of exotic seeds and support communities composed predominately of introduced

species is cause for concern. This is particularly true in areas managed for the protection and

conservation of native biodiversity. Two studies reporting on weed invasion in Kakadu

National Park highlighted this; while the occurrence of exotic species was insignificant in

undisturbed areas, these species were found in increasing abundance in human disturbed

areas, therefore reducing the conservation value of these regions.

The potential for alien species to interrupt ecosystem processes such as nutrient cycles,

hydrology, soil erosion and fertility and disturbance regimes is also an important

consideration (Greenberg et al. 1997).

3.4 MOVEMENT OF A PATHOGEN, PHYTOPHTHORA CINNAMOMI, ALONG ROADS


The spread of P. cinnamomi is associated with road-making activities and materials and

water drainage from road surfaces, with the degree of road usage being influential (Weste &

Taylor 1971; Dawson & Weste 1985; Wilson et al. 2000). The use of P. cinnamomi-infected

gravel in road construction is listed in the Flora and Fauna Guarantee Act (Vic) as a

Potentially Threatening Process (Parks Victoria 1997).

In the Brisbane Ranges National Park, Victoria, all sites infected with P. cinnamomi were

associated with road-making activities (Weste & Taylor 1971; Dawson & Weste 1985). Once

established on road verges, P. cinnamomi commonly extended down-slope via water-borne

zoospores into areas of undisturbed vegetation (Weste & Taylor 1971; Dawson & Weste

1985).

Wilson et al. (2000) recorded that 60% of sites infected with P. cinnamomi in the Eastern

Otway Ranges, Victoria, occurred within 150 m of roads or tracks. This proportion decreased

until only 6% of infected sites occurred between 451ñ600 m from roads and tracks (Wilson et

Three Australian studies investigating the relationship between roads and the distribution of the pathogen Phytophthora cinnamomi have been reviewed. All three were undertaken in Victoria (Weste & Taylor 1971; Dawson & Weste 1985; Wilson et al. 2000).


18

al. 2000). The frequency of occurrence of P. cinnamomi was inversely related to traffic load

of roads and tracks: the lowest infection was recorded close to paved roads, then gravel

tracks, while the greatest incidence of infection was recorded near 4WD tracks (Wilson et al.

2000).


These findings have important implications for the conservation of native vegetation

communities in national parks in Victoria. The îdie-backî of native vegetation caused by P.

cinnamomi is extensive, and impacts of the disease also affect fauna communities (Wilson et

al. 2000). Research shows that roads and tracks are the primary way in which P. cinnamomi

spreads to uninfected areas. It is vital that the spread of this pathogen be limited. Effective

management of infected areas is therefore important, as is the prevention of future infection.

Wilson et al. (2000) recommended management involving controlled access to infected

areas and vehicle wash-down points.

3.5 MOVEMENT OF HUMANS ALONG ROADS: IMPLICATIONS OF INCREASED ACCESS TO NATURAL ECOSYSTEMS

Increased human access to otherwise remote areas is one of the main functions of roads,

particularly those in national parks. Many studies suggest that greater human access to

wilderness areas has significant impacts on the biota in these areas. The most commonly

cited impact is the increased potential for hunting pressure to impact wildlife populations

(Trombulak & Frissell 2000). Many overseas studies on the impact of roads on populations of

large mammals suggest that one of the primary effects of roads on animals such as wolves

(Canis lupus; Thiel 1985; Mech et al. 1988; Thurber et al. 1994), and grizzly and black bears

(Ursus arctos and Ursus americanus, respectively; McLellan & Shackleton 1988; Brody &

Pelton 1989) is the increased access and efficiency of hunters.


19

4 Roads as barriers to the movement animals, potentially fragmenting and isolating populations and communities


Roads function as barriers to the movement of biota in many ways. Discontinuities in

otherwise continuous habitat, in conjunction with the creation of a strip of foreign habitat, may

limit the movement of animals (Mader 1984; Baur & Baur 1990). Behavioural avoidance may

restrict animal movement across roads, as might the actual physical barrier that the roads

form (Barnett et al. 1978; Bakowski & Kozakiewicz 1988; Burnett 1992). Another causal

factor is the tendency of some animals to develop home ranges or territories along physical

boundaries (Barnett et al. 1978; Burnett 1992; Develey & Stouffer 2001). In such situations,

the road itself does not act as a barrier to movement but influences the social interactions of

animals in a way that reduces movement across the road (Barnett et al. 1978; Burnett 1992).

Other more indirect impacts of roads also cause a barrier to animal movement. For example,

traffic noise leads to road avoidance in many animals, as does the physical presence of

vehicles (Singer 1978; Mader 1984). Wildlife mortality associated with road crossing forms

another type of barrier to animal movement (Mader 1984; see Section 5). Forman and

Alexander (1998) hypothesized that the barrier effect of roads on fauna populations is likely

Most research on the effect of roads as a barrier to animal movement has been undertaken in the US and Europe. The movement of mammals across roads in forested and agricultural land has been studied in the US, Canada, Alaska, and Norway (Oxley et al. 1974; Schrieber & Graves 1977; Singer 1978; Rost & Bailey 1979; Kozel & Fleharty 1979; Wilkins 1982; Swihart & Slade 1984; Curatolo & Murphy 1986; Nellemann et al. 2001). Small mammal movement across forest roads has been studied in Poland and Germany (Bakowski & Kozakiewicz 1988; Mader 1984) and across agricultural roads in the UK (Richardson et al. 1997). Australian research on the effects of roads and powerlines has been undertaken in native forest and pine plantations in New South Wales (Barnett et al. 1978), in alpine Victoria (Mansergh & Scotts 1989), and in Queensland rainforest (Burnett 1992; Goosem & Marsh 1997; Goosem 1997). The movement of invertebrate species across roads, mainly through agricultural land, has been studied in Germany and Sweden (Mader 1984; Mader 1988; Baur & Baur 1990; Mader et al. 1990). Studies in Brazil and the US reported on tropical rainforest birds and fish, respectively (Travis & Tilsworth 1986; Develey & Stouffer 2001). Amphibians have been studied in the US (Gibbs 1998), while Reh and Seitz (1990) reported on the genetic isolation of frog populations by German roads. The genetic isolation of bank vole populations by roads has been studied in Germany and Switzerland (Gerlach & Musolf 2000).


20

to affect more species, over a greater area, than either road mortality or road avoidance (see

Section 2), and is potentially the greatest ecological impact of roads with vehicles.

The effect of roads as a barrier to animal movement differs between species, and depends

on road width, road clearing width, and traffic density (Barnett et al. 1978; Baur & Baur 1990;

Richardson et al. 1997). Small terrestrial species with reduced mobility are affected to a

greater degree than larger, more mobile species. The movement of birds and bats is less

restricted by roads due to their mobility. The movement of aquatic organisms may be

inhibited where roads cross streams, potentially affecting aquatic habitats significant

distances from the intersection of road networks and stream systems.

Small mammals

Roads are widely considered to inhibit, but rarely prevent, the movement of small mammals

(Oxley et al. 1974; Barnett et al. 1978; Wilkins 1982; Swihart & Slade 1984; Mader 1984).

Many studies have shown that small mammals will move across roads, though these

movements are usually random and possibly related to dispersal (Burnett 1992; Richardson

et al. 1997). The movement of translocated animals across roads when returning to their

original habitat highlights the permeability of the road barrier to small mammal movement

(Schreiber & Graves 1977; Kozel & Fleharty 1979; Bakowski & Kozakiewicz 1988; Burnett

1992; Richardson et al. 1997).

The degree to which roads inhibit small mammal movement is strongly influenced by road

width; wide roads are thought to act as physical barriers to small mammal movement,

whereas narrower roads are more likely to act as sociological or behavioural barriers (Burnett

1992). Due to the number of factors influencing the potential barrier effect of roads on small

mammals it is difficult to draw definite conclusions as to the minimum road width restricting

animal movement. It is also unknown whether the barrier effect increases gradually with road

width, or whether a critical road width exists above which animals are unable to cross

(Richardson et al. 1997). Roads of widths ranging from 3 m (Barnett et al. 1978; Swihart &

Slade 1984) to 20 m (Oxley et al. 1974; Richardson et al. 1997) inhibited small mammal

movements, while divided roads 90 m wide were suggested to be as effective a movement

barrier as a water body 180 m wide (Oxley et al. 1974). Burnett (1992) concluded that

bitumen roads narrower than 12 m are unlikely to cause genetic isolation in small mammal

populations. If genetic isolation occurred, it would most likely be due to the behavioural,

rather than physical, nature of the barrier (Burnett 1992).

Other factors also influence the potential for small mammals to cross roads. For example, in

a Queensland, Australia rainforest, the presence or absence of culverts was more influential


21

than clearing width on the ability of animals to cross roads between 12 m and 20 m wide

(Goosem 1997).

Studies have recorded that the width of powerline easements that inhibit small mammal

movement is comparatively greater than roads. In Australia, Goosem and Marsh (1997)

found that powerline easements 60 m wide inhibited the movement of small mammals. In

America, translocated mammals crossed easements over 100 m wide (Schreiber & Graves

1977). The barrier that roads form to the movement of small mammals, however, is

essentially different to that created by powerline easements. The barrier effect of roads is

related to many additional factors beyond the physical presence of the road itself, whereas

the barrier caused by powerline easements has been linked primarily to clearing width, and

the structural and microclimatic differences between the natural habitat and the easement

(Goosem & Marsh 1997).

The differential mobility of individuals and species, together with population density and

resource availability, influence the degree to which a road restricts animal movement (Wilkins

1982; Burnett 1992). Australian research has shown that white-tailed rats (Uromys

caudimaculatus) and bush rats (Rattus fuscipes) cross roads more frequently than smaller

species such as brown antechinus (Antechinus stuartii), yellow-footed antechinus (A.

flavipes), and fawn-footed melomys (Melomys cervinipes; Barnett et al. 1978; Burnett 1992).

For white-tailed rats, the findings were attributed to the greater size and mobility of the

species (Burnett 1992). In Poland, roads did not limit the movements of yellow-necked mice

(Apodemus flavicollis) but did restrict bank vole (Clethrionomys glareolus) movements

(Bakowski & Kozakiewicz 1988). However, not all studies have found that roads act as

selective filters for species or gender in small mammals (see Richardson et al. 1997).

Larger mammals

Many factors influence the potential movement of larger mammals across roads, including

resource availability, habitat type, species characteristics, and traffic volume (Singer 1978;

Rost & Bailey 1979; Curatolo & Murphy 1986). In America, mountain goats (Oremnos

americanus) established favoured road-crossing points to reach vital food resources;

crossing success increased with the size of the crossing group and was discouraged by

traffic and humans (Singer 1978). Conversely, the movement patterns of caribou (Rangifer

tarandus) in Alaska were not limited by the presence of roads (Curatolo & Murphy 1986).

In Norway, roads and powerlines have fragmented the last remaining population of wild

reindeer into 26 separate subpopulations that have little or no exchange of individuals

(Nellemann et al. 2001). This finding highlights the importance of determining the impact of


22

roads on different species, as the inhibitory effect of roads on reindeer movements is more

severe than other research on large vertebrates might indicate.


23

Invertebrate fauna

Research shows roads can inhibit the movement of invertebrates such as beetles, snails,

and spiders (Mader 1984, 1988; Baur & Baur 1990; Mader et al. 1990). Road type also

influenced crossing by beetles, which demonstrated greater crossing frequency over grassy

field tracks compared to gravel or paved field tracks (Mader et al. 1990). In Germany and

Sweden, the movement patterns of snails and beetles were altered by roads, resulting in the

increased longitudinal movements of individuals parallel to roads (Baur & Baur 1990; Mader

1984).

Amphibians

In the US and Germany, roads have been found to inhibit the movement of frogs, newts, and

salamanders (Reh & Seitz 1990; Gibbs 1998). However, the cause of the barrier effect of

roads to amphibians is unknown; behavioural avoidance and increased mortality are both

considered possible factors (Gibbs 1998). German research has recorded that roads can

cause the genetic isolation of frog populations within distances of 3 km (Reh & Seitz 1990).

Due to the importance of dispersal in their life-history strategies, it has been hypothesized

that the impact of movement restriction by roads is particularly detrimental to amphibians

(Gibbs 1998).

Avifauna

The effect of roads as a barrier to the movement of birds is less than for small terrestrial

species due to birdsí increased mobility and ability to cross roads with decreased contact

with road surfaces and traffic. However, roads may form significant barriers to the movement

of some birds, particularly understorey species (Develey & Stouffer 2001). In Brazil,

understorey birds crossed roads with closed canopies more readily than roads with open

canopies (Develey & Stouffer 2001). In fact, the territories of understorey birds incorporated

both sides of closed canopy roads, whereas open canopy roads commonly formed territory

boundaries (Develey & Stouffer 2001).

Aquatic organisms

Roads may also form barriers to the movement of aquatic species. Culverts running under

roads allow the passage of water but may hinder the movement of fish and other aquatic

organisms (Travis & Tilsworth 1986). An important factor influencing the barrier effect of

culverts to fish movement is the associated water velocity (Travis & Tilsworth 1986). Water

temperature and fish life stage, primarily the degree of spawning motivation, also affect the

ability of individual fish to utilise culverts for movement (Travis & Tilsworth 1986). In the US,

Travis and Tilsworth (1986) found that excessive pipe velocities restricted the upstream


24

migration of arctic grayling (Thymallus arcticus) for eight days, during which time fisherman

removed almost half the population.

In Australia, Tyson (1980) similarly hypothesized that the increased occurrence of road-killed

platypus (Ornithorhynchus anatinus) was related to the inability of these animals to move

through culverts due to increased water velocity. This conclusion was based on observation

rather than scientific study.

Implications of road barrier effect for individuals and species

Fragmentation and isolation of species populations due to movement inhibition, with

consequent impacts on the genetic structure and composition of populations, is one of the

most commonly discussed implications of the barrier effect of roads (Reh & Seitz 1990).

Reduced gene flow between populations increases the risk of inbreeding and the incidence

of homozygosity, and decreases fecundity, therefore increasing the probability of extinction

of the population (Hartl et al. 1992, cited in Gerlach & Musolf 2000). The impact of genetic

isolation on populations contributes to the fact that habitat fragmentation is one of the

primary causes of faunal extinction (Reh & Seitz 1990; Gerlach & Musolf 2000).

Although the implications of genetic isolation for species survival in fragmented landscapes

are relatively well researched, the specific impacts of roads on genetic isolation are poorly

understood. Reh and Seitz (1990) found that roads in Germany have a significant barrier

effect on the transfer of genes between frog populations within 3 km. The increased isolation

of frog populations by roads led to common frog (Rana temporaria) populations being highly

inbred and homozygous, with reduced gene flow creating local subpopulations each

undergoing different microevolutionary changes (Reh & Seitz 1990). As the roads in this

study were no more than 30 years old, corresponding to ten to 12 frog generations, it was

concluded that significant evolutionary processes at the population level might be rapid for

frog populations in road-fragmented habitats (Reh & Seitz 1990).

A similar study of the genetic isolation of bank vole populations by roads in Germany and

Switzerland found that highways resulted in significant population subdivision, but this impact

was not caused by narrower roads (Gerlach & Musolf 2000). Gerlach and Musolf (2000)

concluded that the barrier effect, caused by reduced migration across the highway, led to

genetic substructuring of bank vole populations. It was hypothesized that the impacts of

highways on genetic isolation of animal populations would be greater for larger species with

smaller population sizes and for endangered species with low dispersal capabilities (Gerlach

& Musolf 2000).

As the mobility of animals and their home range size increases, the primary impact of roads

will shift from the potential physical barrier to movement imposed by roads, to the


25

fragmentation of habitats by roads. Habitat fragmentation reduces the area in which

individuals can obtain necessary resources, therefore reducing the size of populations that

can survive in remaining areas of habitat and increasing competition in these areas.

Consequently, as the density of roads in an area increases, the potential for habitat

fragmentation to render remaining areas unsuitable for animal use likewise increases. For

example, in the US the habitat of grizzly bears (Ursus arctos) was reduced by 8.7% by road

construction (McLellan & Shackleton 1988).


Studies on the barrier effects of roads on Australian fauna have been few. However, the

available data together with international research highlight the potential implications for

native fauna. It is important to consider the contribution of roads to the increasing

fragmentation of habitats, particularly for species that may react negatively to roads as

physical, behavioural or sociological barriers. The associated possibility of genetic isolation

of animal populations is also important. The extent of these impacts generally appears to be

related to road width and traffic volume, and therefore is likely to be much less severe for

tourist or management roads in parks and reserves where traffic volume is low, than for

major highways with high volume traffic. However, the barrier effect of roads leading to

increased habitat fragmentation is one of the primary ways in which roads and tracks through

otherwise undisturbed native vegetation might affect animal species in such areas.


26

5 Roads as sinks ñ wildlife mortality


Animals frequently come into contact with roads during everyday activities and breeding and

dispersal movements, and consequently risk being killed by vehicles. Animals with life

strategies requiring dispersal between different habitat types, such as amphibians, are at

increased risk of mortality (Fahrig et al. 1995). Animals may be attracted to some aspect of

roadside habitats and therefore be at high risk of mortality (e.g., frogs, which can spawn in

roadside drains; Goosem 1997). Animals are also attracted to food resources that are often

increased along roadways (e.g., grain spills from transport vehicles or carrion from previous

roadkills). Some animals are attracted to an attribute of the road itself (e.g., snakes, which

place themselves at increased risk of road mortality by thermoregulating on road surfaces;

Rudolph et al. 1999). Factors influencing road mortality of wildlife include seasonal

differences in animal behaviour, patterns of animal movement, traffic speed and volume,

width of road clearing, roadside vegetation type, and the density and age structure of

roadside wildlife populations (Case 1978; Osawa 1989; Goosem 1997).

Wildlife mortality resulting from roads has been relatively well documented. There is less focus on international research in this section due to the number of Australian studies on wildlife road mortality. Eleven Australian studies, ranging from observational data to scientific research, have addressed road-related wildlife mortality (Vestjens 1973; Anonymous 1980; Tyson 1980; Coulson 1982, 1985, 1989; Ehmann & Cogger 1985; Brown et al. 1986; Osawa 1989; Lepschi 1992; Goosem 1997). These studies include data on birds, macropods, small mammals, frogs, reptiles, and the platypus. Research undertaken in the UK (Hodson 1966, Clarke et al. 1998), the US (Carbaugh et al. 1975; Case 1978), and The Netherlands (Huijser & Bergers 2000) has also reported on road mortality of mammals. Snake mortality has been investigated in the US (Rudolph et al. 1999), amphibian road kills have been researched in Canada (Fahrig et al. 1995), and an Indian study has reported on bird mortality (Dhindsa et al. 1988). Only a small proportion of the extensive international research on road related wildlife mortality has been included in this review. It is important to recognize that while there are numerous reports of the types of species and numbers of animals killed on roads, there is only limited understanding of the relative impact of this mortality on the population dynamics and conservation status of the species concerned.


27

Impact of road mortality on mammals

In New South Wales, Australia, mammals comprised 29% of the total road mortality recorded

over a two-year period (Vestjens 1973). Twelve species, half of which were introduced, were

recorded in total, with European rabbits (Oryctolagus cuniculcus) being most frequently

recorded (Vestjens 1973).

In Victoria, Australia, road mortality of swamp wallabies (Wallabia bicolor) and eastern grey

kangaroos (Macropus giganteus) varied spatially, with increased fatalities along boundaries

between farms and woodlots (Coulson 1982). Temporal variation in the road-kills of these

species was also identified, with increased mortality correlated with periods of drought,

season, and lunar cycle (Coulson 1982, 1989). On North Stradbroke Island, the nocturnal

activity of swamp wallabies increased the number of night-time road fatalities (Osawa 1989).

The occurrence of food resources utilised by eastern grey kangaroos and swamp wallabies

on roadsides increases the potential road mortality of these species (Coulson 1989; Osawa

1989).

Due to the increased proportion of adult males in macropod road fatalities, Coulson (1982)

hypothesized that vehicles may act as selective mortality filters, with unknown impacts on the

social organisation and genetics of local populations. Coulson (1985) attributed this trend to

the greater size of male home ranges, therefore increasing the potential for road contact.

Road mortality of macropods in Australia is unlikely to have a negative impact on population

densities (Coulson 1982, 1989; Osawa 1989).

The road mortality of swamp wallabies on North Stradbroke Island was correlated with traffic

density (Osawa 1989). Traffic volume, however, had little impact on mammal road mortality

in Queensland rainforests or vertebrate road mortality in the US (Case 1978; Goosem 1997).

In the US, a positive correlation was recorded between road mortality and vehicle speed

(Case 1978).

Tyson (1980) observed four occurrences of road mortality of platypus in Tasmania. The

fatalities were related to the difficulty in travelling through culverts carrying water under the

road. These records were not collected as part of a scientific study.

In the US, road mortality of white-tailed deer was found to decrease with animal habituation

to roads (Carbaugh et al. 1975). In The Netherlands, road mortality of hedgehogs potentially

[a3]reduced population densities by 30%, thereby affecting the survival probability of local

populations (Huijser & Bergers 2000). In England, road mortality of badgers is the largest

single cause of death of these animals (Clarke et al. 1998). The differing degree of effect of

road mortality on these animals highlights the potential variability between species.


28

Impact of road mortality on amphibians and reptiles

Road mortality has been cited as a potentially important factor in the worldwide decline of

amphibians (Fahrig et al. 1995). Amphibians move between different habitats for foraging,

breeding, and over-wintering, consequently increasing the potential for their movements to

intersect with roads and result in mortality (Fahrig et al. 1995). Ehmann and Cogger (1985)

estimated that 4,450,000 frogs are killed annually on sealed roads in Australia, and in the UK

frogs had the highest recorded road mortality of any animal group (Hodson 1966). A study by

Goosem (1997) in the Wet Tropics World Heritage Area in Queensland, however, found that

amphibian road mortality was not significant enough to depress local population sizes. In

contrast, local reductions in anuran densities have been recorded in roadside areas in

Canada (Fahrig et al. 1995). In Canada, traffic volume was an influential factor [a4]in frog and

toad fatality and reduced population density (Fahrig et al. 1995).

Roads and associated traffic have also been implicated in the decline of snake populations

(Rudolph et al. 1999). In the US, populations of large snakes can be reduced by up to 50%

within 450 m of moderately used roads (Rudolph et al. 1999). Rudolph et al. (1999)

extrapolated that road mortality may depress populations of large snakes by 50% across the

entire area of eastern Texas. Road mortality therefore has the potential to act as a îsinkî for

local populations experiencing such significant levels of road mortality (Rosen & Lowe 1994,

cited in Rudolph et al. 1999).

While specific data on road-related snake mortality is unavailable for Australia, Ehmann and

Cogger (1985) estimated that 1,030,000 reptiles are killed annually on paved roads in

Australia. In New South Wales, Vestjens (1973) reported that reptiles constituted only 5% of

the total road-related wildlife mortality. Six species of lizard, two of snake and one of tortoise

contributed to these figures, with the common bearded dragon (Amphibolurus barbatus

barbatus) being the most frequently killed (Vestjens 1973).

Impact of road mortality on birds

In a study in New South Wales, Australia, birds were killed more than any other animal,

accounting for 66% of total road mortalities (Vestjens 1973). Contradictory results were

recorded in Queensland, where only 2% of all animals killed were birds (Goosem 1997). This

discrepancy may be related to the differing roadside habitats sampled, with more birds killed

on open, agricultural roads than in rainforests, where birds avoid wide roads (Goosem 1997).

Bird mortality in Australia affects a large number of species. Vestjens (1973) and Lepschi

(1992) recorded 64 and 36 species, respectively, in New South Wales studies, and Brown et

al. (1986) recorded 32 species in Western Australia. However, only a few of these species

were killed in large numbers (Vestjens 1973; Brown et al. 1986, Lepschi 1992).


29

Nevertheless, an account of the road mortality of regent parrots (Polytelis anthopeplus) in

Victoria provides an extreme example of the degree to which animals frequenting roadsides

are at risk of mortality. An anonymous observer (1980) provided an account of around 175

birds killed within a period of just over a week due to a large grain spill attracting large flocks

to a roadside area. Even so, road mortality in Australia is unlikely to endanger any bird

species, with vehicles being likened to any other predator (Brown et al. 1986, Lepschi 1992).

In New South Wales and Western Australia, the species most affected by road mortality were

magpies (Gymnorhina tibicen), galahs (Cacatua roseicapilla), magpie-larks (Grallina

cyanoleuca), and western rosellas (Platycercus icterotis; Vestjens 1973; Brown et al. 1986,

Lepschi 1992). The high mortality of these species was related to their increased usage of

roadside areas for foraging and nesting (Vestjens 1973; Brown et al. 1986, Lepschi 1992).

Mortality of juvenile magpies was significantly higher than adult mortality (Vestjens 1973;

Lepschi 1992).

The occurrence of road-killed birds varies spatially, with mortality being highest at the

intersection of roads and creeks (Brown et al. 1986). Increased road mortality of birds in the

localised area of grain spills also influences the spatial occurrence of road-kills. Temporal

variation in road mortality of birds in Australia has been related to breeding season;

increased deaths occur between November and March (Brown et al. 1986).

In India, research has positively correlated bird mortality with traffic speed (Dhindsa et al.

1988). Dhindsa et al. (1988) suggested that, due to bird mortality, roads might act as

ecological traps.

Impact of road mortality on wildlife populations

The number of animals killed on roads is widely considered to be underestimated.

Contributing factors include the potential movement of injured animals away from the road,

removal by scavengers, or decomposition prior to inclusion in road-kill tally (Coulson 1985).

However, estimates of the annual road mortality for some animals do exist, including:

5,480,000 frogs and reptiles in Australia (Ehmann & Cogger 1985); 365,000,000 vertebrates

in the US (Lalo 1987, cited in Rudolph et al. 1999); 113,000ñ340,000 hedgehogs in The

Netherlands (Huijser & Bergers 1998, cited in Huijser & Bergers 2000); 50,000 badgers in

the UK (Clarke et al. 1998); and 7,000,000 birds in Bulgaria (Nankinov & Todorov 1983, cited

in Dhindsa et al. 1988). While these calculations may be outdated, they highlight the

magnitude of road related fauna mortality.

The impact of road mortality on the density of wildlife populations is species specific, with

influential factors including mobility, reproductive, and adult mortality rate (Coulson 1985;

Rudolph et al. 1999). Road deaths in some species may also be influenced by the sex and


30

age of individual animals, with juveniles and males usually considered at higher risk (Coulson

1985). Habitat preference also impacts the potential for a species to be affected by road

mortality (Coulson 1985).

Some animals are seen to be at greater risk of road mortality than others. For example, the

impact of road mortality on reptile and amphibian populations is thought to be more

significant than on small mammals (Bennett 1991; Fahrig et al. 1995; Rudolph et al. 1999).

The impact of road mortality on endangered species is also often high. In Queensland, a

population of the southern cassowary (Casuarius casuarius johnsonii) decreased by around

14% over three years due to road mortality (Bentrupperbaumer 1988, cited in Goosem

1997). Generally, however, common species are killed most frequently, and the proportion of

mortality attributable to roads is usually only a small percentage of the total mortality of a

population (Coulson 1985). Nevertheless, it has been suggested that road mortality has a

greater, and more direct, impact on the persistence of wildlife populations than reduced

movement caused by roads (Jaeger & Fahrig 2001, cited in Fahrig 2001).

Wildlife road mortality as a source of data on animal ecology and biology

Road-killed animals can provide valuable information on a wide range of aspects of animal

species ecology and biology. In Tasmania, surveys of road-killed red-bellied pademelons

(Thylogale billardierii) were used to monitor the distribution of the species (Coulson 1985). In

Victoria, road-kills helped identify the existence of the long-footed potoroo (Potorous

longipes), with the second two recorded specimens of the species being road-killed animals

(Coulson 1985). The number of road-killed animals has also been used to determine

population indexes for different species (Case 1978 and references within).

Analysis of road-killed animals may also provide important dietary or reproductive information

on a species (Coulson 1985). On North Stradbroke Island, the stomach contents of swamp

wallabies (Wallabia bicolor) were used to determine the food preferences of the species

(Coulson 1985).


The impact of road mortality on wildlife populations in national parks will differ between

species, and may also be affected by traffic volume and speed. Road mortality is generally

considered not to be significant as a threatening process to most species, though threatened

and endangered species likely to be at risk of road mortality should be given particular

attention. The greatest impacts in parks and reserves are likely to be where busy highways

or major roads pass through the reserve, rather than from internal tracks used for access or

management purposes.


31

The potential for humans to be harmed in collisions with wildlife is another issue for

consideration. Coulson (1985) reported the death of three people in Victoria over a six-year

period as a result of wildlife collisions.


32

6 Mechanisms that potentially reduce the barrier effect of roads


The barrier effects of roads, caused by the inhibition of animal movement and animal

mortality, potentially reduce the viability of habitats for sustaining wildlife populations and the

viability of the populations themselves. The barrier to the movement of biota formed by roads

can be reduced in many ways and to differing degrees for a wide range of animals. Pre-

existing culverts allowing water passage under roads are often used by wildlife for

movement, underpasses allowing animal movement can be created, and landscape

connectors allow the movement of many natural flows across roads (Forman & Hersperger

1996). Structures allowing the movement of animals across roads usually incorporate road

fencing to reduce road mortality of wildlife and to funnel animal movement through the

crossing (Reed 1981; Foster & Humphrey 1995). Such road culverts and underpasses serve

two primary functions: to increase the permeability of road barriers to wildlife movement and

to reduce the number of animals killed on roads (Yanes et al. 1995; Clevenger & Waltho

2000).

Factors influencing fauna use of wildlife road crossings

Research shows that attributes of the underpass are the dominant factor influencing the use

of these structures by animals (Reed 1981; Hunt et al. 1987; Foster & Humphrey 1995;

Yanes et al. 1995; Clevenger & Waltho 2000). Openness, and the potential for animals to

see the habitat on the other side of the underpass, are commonly cited as the most important

factors (Reed 1981; Foster & Humphrey 1995; Clevenger & Waltho 2000). The location of

wildlife underpasses at wildlife crossing points also increases their success (Foster &

Humphrey 1995). Associated levels of human activity, and the complexity of roadside

vegetation, also influence the degree to which underpasses are used by all species (Hunt et

al. 1987; Yanes et al. 1995; Clevenger & Waltho 2000).

Most studies into the success of wildlife underpasses for increasing animal movement across roads have been undertaken in the US and Canada, and primarily relate to large mammals (Reed 1981; Foster & Humphrey 1995; Clevenger & Waltho 2000). The potential for road culverts and underground tunnels to facilitate animal movement has also been studied in Australia (Hunt et al. 1987; Mansergh & Scotts 1989). Culvert use by wildlife has been reviewed in Spain (Yanes et al. 1995). Forman and Hersperger (1996) and Forman et al. (1997) provide comprehensive overviews of different wildlife-crossing structures used by animals around the world.


33

The crossing of small and medium-sized mammals is affected in different ways (Yanes et al.

1995). A study in Spain found that small mammals were predominately influenced by culvert

dimensions, whereas medium-sized mammals responded primarily to total road width (Yanes

et al. 1995). Carnivores using underpasses in the US were primarily influenced by landscape

variables such as proximity to drainage structures, while the structural features of

underpasses affect their use by ungulates (Reed 1981; Clevenger & Waltho 2000).

Australian research has determined that age of the crossing structure influenced its

differential use by animal species (Hunt et al. 1987). For example, long-established drainage

culverts were predominately used by native small mammals, whereas newly constructed

tunnels were mainly used by feral predators (Hunt et al. 1987). These findings were related

to the degree of native vegetation cover at the tunnel entrance, and led to the hypothesis that

native mammals will use tunnels specifically constructed for wildlife movement following

revegetation. Larger mammals such as wombats and macropods did not use culverts despite

the adequate size of these structures (Hunt et al. 1987). This may be related to the reduced

inhibition of movement of these species by roads, therefore decreasing the necessity for

wildlife crossing structures, or possibly to the increased predation risk associated with tunnel

use (Hunt et al. 1987).

Species use of wildlife road crossings

In Victoria, Australia, habitat continuity for mountain pygmy-possum (Burramys parvus)

populations was restored by the construction of tunnels connecting road-bisected habitat

(Mansergh & Scotts 1989). The natural rocky scree required for movement by this species

was replicated in the tunnel interior, therefore increasing successful tunnel use and reducing

the need for drift fences along the road (Mansergh & Scotts 1989). The tunnels were found to

facilitate the dispersal of male pygmy-possums within two weeks of completion, subsequently

restoring the social organisation of the population, and increasing the survival of females

(Mansergh & Scotts 1989).

In the US, wildlife underpasses facilitated the movement of panthers (Felis concolor coryi),

bobcats (Lynx rufus), and black bears (Ursus americanus) between sections of their home

ranges otherwise isolated by roads, thereby reducing habitat fragmentation (Foster &

Humphrey 1995). Foster and Humphrey (1995) found that underpasses reduced the road

mortality of some species, therefore preventing roads from becoming demographic sinks for

these animals. Some species also utilised underpasses for foraging, including wading birds,

raccoons (Procyon lotor), and deer (Foster & Humphrey 1995).

In Spain, many animal groups, including amphibians, reptiles, and mammals, were found to

use culverts for road crossing, with greatest use by small mammals (Yanes et al. 1995).


34

Crossing by most groups, however, was negatively correlated with the presence of detritus

pits at the entrance of culverts (Yanes et al. 1995).

In Europe and America, amphibian tunnels allowed animal passage across roads, and

reduced mortality (Forman & Hersperger 1996, Forman et al. 1997). Tunnels specifically

targeting movement of amphibians have not been used in Australia.

Species adaptation to wildlife road crossings

Research findings differ with regard to the potential for animals to adapt to underpass use. In

Australia, mountain pygmy-possums took only two weeks to use a tunnel connecting

otherwise isolated habitat (Mansergh & Scotts 1989). Similarly, Yanes et al. (1995) found

that none of the animal groups studied in Spain showed any sign of culvert avoidance.

Animals are known to become used to the structural features of underpass crossings with

time (Clevenger & Waltho 2000). However, Reed (1981) states that 75% of mule deer in

studies in the US exhibited wary and frightened behaviour on exiting underpasses. This

decade-long study indicated minimal familiarisation with underpasses for this species (Reed,

1981).

Wider implications of wildlife road crossings

One important impact of wildlife underpasses is their potential to alter the balance of

predator-prey interactions by providing carnivores with greater access to prey species in the

area of underpass crossings (Hunt et al. 1987; Clevenger & Waltho 2000). Increased prey

mortality or the altered use of road underpasses by prey species are potential outcomes of

this impact. For example, an English study found that badgers disrupted underpass use by

hedgehogs due to the predator avoidance behaviour of hedgehogs (Clevenger & Waltho

2000). In the US, Foster and Humphrey (1995) likewise recorded deer avoidance of

underpasses used by bobcats (Felis rufus) and panthers (Felis concolor coryi). Waters

(1988, cited in Foster & Humphrey 1995) found that wolves (Canis lupus) and coyotes (Canis

latrans) learned to herd their prey against the road fencing to aid capture.

Other types of wildlife road crossing structures

In Europe and America, overpasses have been constructed to allow the movement of

animals over roads (Forman et al. 1997). Overpasses in the US have been successful in

facilitating the movement of deer and other animal species across roads (Forman et al.

1997). No such overpasses have been constructed in Australia.

Landscape connectors in Switzerland 140 m and 200 m wide have been created to increase

the potential movement of all natural flows across roads, including ground water as well as

animals of all sizes (Forman et al. 1997). There are plans for similar connectors with widths

between 1.5 km and 2 km to be constructed in both Switzerland and Holland (Forman et al.


35

1997). The degree to which natural processes move across such landscape connectors is

unknown, however the maintenance of all the natural flows occurring in landscapes is

important, and the ability to achieve this would decrease the barrier effect of roads

significantly.


The use of underpasses by animal species has the potential to maintain habitat continuity in

roaded areas, and also to reduce road-related wildlife mortality. Research into the barrier

effects of roads and road mortality shows that both these impacts can affect Australian

species and populations. It is therefore important to reduce the effects of these impacts on

native fauna species. A study by Mansergh and Scotts (1989) documenting the creation, and

rapid utilisation, of under-road tunnels by the mountain pygmy-possum illustrates the degree

to which such structures can successfully reduce habitat fragmentation. However, the

findings of Hunt et al. (1987) of increased use of wildlife tunnels by feral predators have

important implications for the increased predation of native species and the greater facility of

predator movement in native habitats. Further research into the degree of habitat

fragmentation caused by roads for Australian fauna species will help to identify species

requiring the aid of road-crossing structures, and therefore direct the construction of suitable

underpasses.


36

7 Roads as a source of biotic and abiotic effects The biotic and abiotic effects of roads on surrounding environments are numerous, diverse,

and far-reaching. These effects may be derived from specific structural features of the road

itself or from road use and maintenance (Forman 1999). The degree to which the biotic and

abiotic effects of roads have an impact on the environment depends on a number of factors

and frequently varies both spatially and temporally. The combination of these road effects

and increased human access to an area may alter local disturbance regimes such as the

pattern of fire occurrence.

7.1 POLLUTION FROM ROADS AND MOTOR VEHICLES


Numerous chemical and physical pollutants result from all phases of road construction,

maintenance, and use. All affect the surrounding environment in various ways and to

differing degrees (Scheidt 1971). Vehicle emissions contribute some of the primary pollutants

arising from road usage (Lagerwerff & Specht 1970; Scheidt 1971; US Department of Health,

Education, and Welfare 1971). Motor vehicles are also the source of pollutants from engine

oils and greases, depositions from tyres, and litter and debris from vehicles and their

passengers (Lagerwerff & Specht 1970; Scheidt 1971). Roadway maintenance contributes

pollutants in the form of de-icing salts and herbicides (Scheidt 1971).

Atmospheric pollution from vehicle emissions contains carbon monoxide, atmospheric lead,

hydrocarbons, nitrogen oxides, sulphur oxides, particulates, and sometimes nickel

(Lagerwerff and Specht 1970; Scheidt 1971; US Department of Health, Education, and

Welfare 1971). Most research relates to the environmental impacts of lead from vehicle

emissions. Studies have investigated the roadside lead content of soil (Chow 1970;

Lagerwerff & Specht 1970; Motto et al. 1970; Wylie & Bell 1973; Quarles et al. 1974;

Research into the impact of pollution from roads and vehicles has been undertaken in Australia, Scotland, New Zealand, and the US. The bulk of this research, encompassing the effects of road pollution on soil, vegetation and fauna, has been conducted in the US (Chow 1970; Lagerwerff & Specht 1970; Motto et al. 1970; Quarles et al. 1974; Goldsmith et al. 1976, OíNeill et al. 1983; Goldsmith & Scanlon 1977; Clark 1979; Grue et al. 1984; Harrison & Dyer 1984). The impacts of road pollution have been studied in Australia (Wylie & Bell 1973; Noller & Smythe 1974; Clift et al. 1983), New Zealand (Collins 1984), and Scotland (Iredale et al. 1993). All studies, except that by Harrison and Dyer (1984), were undertaken on heavily used urban arterials. Many studies have reported on the environmental impacts of de-icing salts, but as this practice is not undertaken in Australia this research has not been reviewed.


37

Goldsmith et al. 1976, Clift et al. 1983; OíNeill et al. 1983; Collins 1984; Iredale et al. 1993),

vegetation (Chow 1970; Lagerwerff & Specht 1970; Motto et al. 1970; Wylie & Bell 1973;

Noller & Smythe 1974; Quarles et al. 1974; Goldsmith et al. 1976, OíNeill et al. 1983; Collins

1984), and invertebrate and vertebrate fauna (Goldsmith & Scanlon 1977; Clark 1979; OíNeill

et al. 1983; Grue et al. 1984; Harrison & Dyer 1984). The following results were primarily

obtained from research undertaken on roads with high traffic volumes (15,000ñ50,000 cars

per day).

Lead contamination of soil and vegetation

Traffic volume, distance from road, soil depth, prevailing wind direction, and the occurrence

of roadside vegetation all influence the deposition and accumulation of lead in the soil. Most

studies have recorded a strong positive correlation between soil lead concentration and road

proximity, with significantly increased lead concentrations occurring up to 80 m from roads

(Chow 1970; Lagerwerff & Specht 1970; Motto et al. 1970; Wylie & Bell 1973; Quarles et al.

1974; Goldsmith et al. 1976, Clift et al. 1983; Collins 1984; Harrison & Dyer 1984). Lead

compounds contained in the soil are relatively insoluble and accumulate in higher

concentrations near the surface (Chow 1970; Wylie & Bell 1973; Clift et al. 1983; Collins

1984).

Lead concentrations in vegetation are likewise higher near roads, but do not reach levels

recorded in soil (Chow 1970; Lagerwerff & Specht 1970; Motto et al. 1970; Wylie & Bell

1973; Noller & Smythe 1974; Quarles et al. 1974; Goldsmith et al. 1976, Collins 1984;

Harrison & Dyer 1984; Iredale et al. 1993). The concentration of lead in roadside vegetation

varies seasonally and differs between species (Quarles et al. 1974; Goldsmith et al. 1976,

Iredale et al. 1993). However, the contamination of plants by insoluble lead in surface soil is

unlikely to reduce plant growth (Quarles et al. 1974).

Lead contamination of wildlife

Lead levels in invertebrate and vertebrate animals have been positively correlated with

proximity to roads (Quarles et al. 1974; Goldsmith & Scanlon 1977; Grue et al. 1984) and

traffic volume (Clark 1979; OíNeill et al. 1983; Grue et al. 1984). Invertebrates accumulate

lead relative to their proximity to the soil and the amount of soil and litter they ingest (Grue et

al. 1984). Lead concentrations are consequently highest in detritivores and lowest in

herbivorous insects (Goldsmith & Scanlon 1977).

The lead content of invertebrates and plants affects the level of lead in small mammals, bats,

birds and larger herbivorous species, as animals primarily accumulate lead through dietary

intake (Quarles et al. 1974; Goldsmith & Scanlon 1977; Clark 1979; OíNeill et al. 1983; Grue

et al. 1984; Harrison & Dyer 1984; Collins 1984). Metabolic differences, sex, quantity of food


38

consumed, home range, burrowing habit, and life span all influence lead uptake by animals

(Quarles et al. 1974; OíNeill et al. 1983; Harrison & Dyer 1984). Lead concentrations in

aerially feeding birds are commonly lower than those in birds and small mammals that feed

on terrestrial invertebrates (Grue et al. 1984). The levels of lead in bats have been found to

exceed those in both terrestrial mammals and birds (Clark 1979; Grue et al. 1984).

Excess lead in mammals has been correlated with reproductive problems, decreased body

weight, renal abnormalities, and mortality (Goldsmith & Scanlon 1977; OíNeill et al. 1983). In

birds, lead contamination has been implicated in weight and vision loss, wing and leg

paralysis, altered nerve function, behavioural alterations, different immune responses, and

altered levels of brain enzymes (Grue et al. 1984). However, studies of wild animal

populations have not recorded such impacts, nor found populations to be limited by lead.

Other chemical pollutants derived from vehicle emissions

Nitrogen concentrations in plants and soil have been recorded at higher than normal levels in

roadside areas, and have been related to the nitrogen oxides emitted in vehicle exhausts

(Port & Thompson 1980; Spencer & Port 1988; van Schagen et al. 1992). Increased nitrogen

in plants has been linked to outbreaks of herbivorous insects, and increased nitrogen in soil

has been correlated with reduced plant germination (Port & Thompson 1980; Spencer & Port

1988; van Schagen et al. 1992).

Concentrations of cadmium and zinc from engine oils and tyres, and nickel from gasoline, in

roadside soil and vegetation are likewise higher closer to roads and the soil surface

(Lagerwerff & Specht 1970). These are just some of the pollutants other than lead that arise

from roads and vehicle usage. It has been suggested that the more diffuse effects of such

less obvious pollutants are likely to have greater impacts on surrounding environments

(Bennett 1991).


Pollutants derived from road use affect surrounding environments. However, these impacts

are strongly related to road usage; the impact of road pollutants on lesser-used roads is

unlikely to be significant. The possible exceptions include the increased availability of

nitrogen along roadsides, a commonly limiting resource in Australian soils, and the impact of

herbicide use. The effects of herbicide use on surrounding environments, however, are

poorly understood.


39

7.2 IMPACTS OF ROADS ON HYDROLOGICAL SYSTEMS


Roads affect the natural horizontal flows in a landscape by altering the pattern of water

movement above and below the soil surface (Forman 1999). Increased surface run-off

generated by impervious road surfaces is one of the primary impacts on natural water flows

(King & Tennyson 1984). Roads may also divert subsurface flows to surface flows, further

increasing flow volumes and altering soil moisture regimes (King & Tennyson 1984). The

concentration of these increased volumes of water in roadside ditches and culverts changes

natural water flow paths (King & Tennyson 1984). These impacts affect streamflow quantities

and regimes, commonly resulting in increased magnitude and frequency of peak flows (King

& Tennyson 1984; Jones et al. 2000). Water generated by road systems also carries

sediment and chemical pollution from the road surface into aquatic ecosystems, affecting the

quality of aquatic habitats (Gjessing et al. 1984; Costantini et al. 1999). The potential for

roads to negatively impact aquatic ecosystems is significant (Eaglin & Hubert 1993; Ruediger

& Ruediger 1999).

Roads pose a significant risk to wetland biodiversity on local and regional scales, as any

impact of roads affecting a single biotic or abiotic feature of a wetland is likely to affect the

dynamics of the entire wetland (Thrasher 1983; Department of Conservation and

Environment 1985; Findlay & Houlahan 1997; Findlay & Bourdages 2000). In Canada, the

species richness of wetlands was negatively correlated with the density of paved roads within

2 km (Findlay & Houlahan 1997). The full extent of the impact roads have on wetlands,

however, is unlikely to be immediately apparent due to the lag time existing between road

construction and the resultant species loss (Findlay & Bourdages 2000).

The impact of roads on hydrological systems has been studied in Australia (Costantini et al. 1999), the US (King & Tennyson 1984; Reid & Dunne 1984; Bilby et al. 1989; Eaglin & Hubert 1993; Wemple et al. 1996, Rummer et al. 1997; Ruediger & Ruediger 1999; Jones et al. 2000), New Zealand (Fahey & Coker 1992), and Norway (Gjessing et al. 1984). All studies except Gjessing et al. (1984) reported on the hydrological impacts of roads in forested catchments. Wetlands have been identified by many studies as being highly susceptible to the impacts of roads and road construction. However, few studies have actually investigated the specific impacts of roads on wetland ecosystems. Such studies have been undertaken in Canada (Findlay & Houlahan 1997; Findlay & Bourdages 2000), and the US (Thrasher 1983). An Australian government report raised ecological issues associated with road impacts on wetlands (Department of Conservation and Environment 1985).


40

Run-off

Road run-off commonly increases the moisture content in roadside soil. Roadside soil

receives at least twice as much rainfall as other areas, and the interception of subsurface

flows by roads further increases the water availability in roadside areas (Huey 1941; King &

Tennyson 1984). Run-off distance is influenced by topography, surface morphology, and

vegetation (Costantini et al. 1999).

Run-off from roads may increase the rate and extent of erosion, reduce percolation and

aquifer recharge rates, alter stream channel morphology, and increase stream discharge

rates (Forman & Alexander 1998). The increased access of roadside habitats to water has

been cited as an influential factor in the increased productivity and diversity common to

roadside vegetation communities (Huey 1941; Johnson et al. 1975; Lamont et al. 1994a,

1994b, Williams et al. 2001). Large volumes of run-off from forest roads potentially threaten

aquatic ecosystems and marine environments (Costantini et al. 1999).

Alterations to stream flow

Road run-off affects the flow regimes of streams and other hydrological systems in roaded

catchments (King & Tennyson 1984; Wemple et al. 1996, Jones et al. 2000). Hydrological

systems and road networks are fundamentally different in that stream networks flow down-

slope while roads cut across slopes (Jones et al. 2000). The impact of roads on stream

hydrology is greatest at points where roads intersect streams, but may extend large

distances from the point of direct disturbance (Forman 1999; Ruediger & Ruediger 1999).

Two hydrological flow paths link road systems and stream channels: roadside ditches

draining directly to streams and roadside ditches draining to culverts (Wemple et al. 1996). In

the US, Wemple et al. (1996) found that over half the total road length in two basins was

hydrologically connected to stream channels, leading to an increased drainage density of

between 21% and 50%.

The impact of roads on forested catchments depends on the density, width, location, spatial

arrangement, and design of roads and on the catchment topography (King & Tennyson 1984;

Costantini et al. 1999). Consequently, the impact of roads on catchment hydrology is highly

variable (King & Tennyson 1984). Nevertheless, estimates have been made as to the

minimum road coverage that alters streamflow in forested catchments in America. Only small

increases in water yield occur if less than 8% of the total catchment area is roaded, average

peak flows may be significantly increased when between 8% and 12% of the total catchment

is roaded, and large peak flows are significantly increased when more than 12% of the

catchment area is covered by roads (Harr et al. 1975, cited in King & Tennyson 1984).


41

Roads also alter the natural balance that exists between peak flows and the resistance to

change of stream and riparian networks (Jones et al. 2000). Therefore, increased flow

regimes associated with roads may permanently alter stream and riparian habitats, in the

process reducing their ability to respond to disturbance and impacting ecosystem resilience

(Jones et al. 2000).

Chemical transport

Road run-off often carries organic and inorganic pollutants and other suspended matter

(Gjessing et al. 1984). The distance to which these pollutants are transported varies. In

Norway, pollutants, including organic carbon and metal ions, were found in high quantities in

snow within 5 m of roads, and were still recorded at distances of up to 300 m (Gjessing et al.

1984). Implications exist for the fate of these pollutants following snowmelt.

Pollutants contained in road run-off also have the potential to contaminate water bodies and

stream systems. In Norway, road run-off caused moderate pollution of water systems by

inorganic pollutants including lead, zinc, and chromium, and high pollution of water systems

by organic pollutants and suspended matter (Gjessing et al. 1984). Water bodies such as

lakes have been known to act as sinks for most of these pollutants, reducing the

contamination of any outflowing water (Gjessing et al. 1984).

Sedimentation

Roads in forested catchments utilised for timber production in the US significantly increase

the sedimentation of natural water flows (King & Tennyson 1984; Eaglin & Hubert 1993;

Rummer et al. 1997; Ruediger & Ruediger 1999). The impact of such road sediments on

streams depends on the characteristics and quantity of sediment entering the stream system

(Bilby et al. 1989). However, sediment movement beyond the limits of road clearings is often

insignificant, with ditches accumulating the majority of road sediment (Rummer et al. 1997).

Road/stream intersections, road length, road area, position in a catchment, traffic volume,

and precipitation all increase the rate of sediment delivery to streams (Bilby et al. 1989;

Eaglin & Hubert 1993; Costantini et al. 1999). In the US, for example, heavily used roads

contribute 130 times more sediment to hydrological systems than abandoned roads (Reid &

Dunne 1984). Natural catchment features including site hydrology, climate, soils, topography,

and quantity of vegetation along roadsides also affect the rate of erosion and sedimentation

from forest roads (Rummer et al. 1997).

Road surface material and ballast depth may also influence the rate and quantity of erosion

from roads and sediment deposition to stream systems (Bilby et al. 1989). In the US, paved

roads yield less than one percent of the sediment quantity produced by heavily used gravel


42

roads (Reid & Dunne 1984). Gravel roads are recognised as an important source of fine-

grained sediment (Reid & Dunne 1984).

Channel disturbance and alteration and the removal and deposition of material are the

primary effects of sedimentation on hydrological systems (Bilby et al. 1989). The deposition

or suspended movement of sediment affects habitat quality for aquatic biota directly by filling

pools and other interstitial spaces in streambeds (Bilby et al. 1989). The effects of deposited

sediment are better understood than those of suspended material, the impacts of which are

harder to quantify but potentially affect a greater area (Bilby et al. 1989). The ecosystem

resilience of hydrological systems may also be affected by alterations to the spatial

distribution of starting and stopping points of debris flow in landscapes (Jones et al. 2000).

Fine-grained sediments have particularly detrimental impacts on aquatic ecosystems, fish

habitat, and water quality (Reid & Dunne 1984; Eaglin & Hubert 1993; Ruediger & Ruediger

1999). In New Zealand, up to 200 t of fine sediment may be transported annually to local

embayments from a harvested forest, increasing the suspended sediment concentration of

seawater (Fahey & Coker 1992). In the US, direct positive correlations have been drawn

between the quantity of fine sediment in salmon-spawning gravels and the length of roads in

the basin (Reid & Dunne 1984). Road related sedimentation potentially reduces salmonid

fish density, survival of salmonid eggs and alevins (due to reduced water flow through

streambed gravel), and the survival of anadromous fish (Reid & Dunne 1984; Bilby et al.

1989; Eaglin & Hubert 1993).


The impact of roads on hydrological systems is particularly important in forests managed for

timber production, though impacts will also occur in unlogged areas. Roads in mountainous

areas are likely to have an even greater impact on the hydrology of the surrounding

catchment. Of great significance is the impact that roads have on stream habitats and

aquatic ecosystems; an area requiring greater research in Australia. The impact of road run-

off on the water and nutrient content of roadside soils is also important in the management of

terrestrial habitats. The effects of this impact will be greater in areas where vegetation is

adapted to conditions of low soil nutrients (e.g., heathlands) and water availability.


43

7.3 IMPACTS OF ROADS ON SOIL CHARACTERISTICS


Roads alter the properties and structure of soil, the activity of micro and macroinvertebrate

soil fauna, and also leaf litter layers, consequently affecting the structure and floristics of

roadside vegetation communities (Muskett & Jones 1981; Olander et al. 1998; Bolling &

Walker 2000; Haskell 2000). These impacts are modified by the type of road construction

and the duration and intensity of road use (Bolling & Walker 2000). These impacts affect the

immediate and surrounding environment from the time of road construction until often

significant periods after road abandonment (Olander et al. 1998; Bolling & Walker 2000).

Soil compaction

Roads cause soil compaction on the road itself and, due to road building activities, in

adjacent areas (Guariguata & Dupuy 1997; Bolling & Walker 2000). Soil compaction

influences the succession of vegetation communities developing on roadsides or abandoned

roads, as increased compaction commonly suppresses plant growth (Guariguata & Dupuy

1997; Olander et al. 1998; Bolling & Walker 2000). In the US, there was no difference in the

soil compaction of roads abandoned for 5 and 88 years, leading to the recommendation to

decompact abandoned roads to enhance plant growth potential (Bolling & Walker 2000).

Soil substrate properties, invertebrate fauna, and leaf litter

Roads affect the abundance and richness of macro and microinvertebrate soil fauna and leaf

litter layers (Muskett & Jones 1981; Haskell 2000). In the US, the abundance and diversity of

macroinvertebrate fauna and depth of leaf litter were significantly reduced with increasing

proximity to roads (Haskell 2000). These impacts on litter layer depth and macroinvertebrate

abundance extended for 100 m while impacts on faunal diversity persisted for 15 m (Haskell

2000). Similarly in the UK, soil respiration related to decreased microbial activity was reduced

up to 100 m from roads (Muskett & Jones 1981).

Research on the impact of roads on soil characteristics has reported on soil substrate

properties, soil compaction, litter layers, and invertebrate soil fauna. Forest habitats

have been studied in the US (Bolling & Walker 2000; Haskell 2000; Holl et al. 2000),

Alaska (Auerbach et al. 1997), Costa Rica (Guariguata & Dupuy 1997), and Puerto Rico

(Olander et al. 1998). Agricultural landscapes have been studied in the UK (Muskett &

Jones 1981; Spencer & Port 1988). No studies specifically on this topic have been

undertaken in Australia.


44

Reductions in abundance and diversity of macroinvertebrate soil fauna may alter the ability of

soil to process energy and nutrients, with associated repercussions for the distribution and

abundance of other organisms, particularly plants and animals that forage on litter-dwelling

macroinvertebrates (Haskell 2000). Haskell (2000) suggested that road related alterations to

leaf litter depth might change the growth rate, abundance, and diversity of vegetation in

linear strips along roads.

Organic matter accumulation (Muskett & Jones 1981; Guariguata & Dupuy 1997; Auerbach

et al. 1997; Olander et al. 1998) and moisture content (Muskett & Jones 1981; Auerbach et

al. 1997; Olander et al. 1998) are both commonly decreased in roadside soils. Similarly,

Haskell (2000) suggested that increased wind velocity and solar radiation along forest roads

might cause drying litter, thereby altering decomposition rates. The physical displacement of

litter by wind also influenced litter depths (Haskell 2000). The findings of decreased moisture

in roadside soil conflict with the findings of other studies discussed previously (Huey 1941;

Johnson et al. 1975, Williams et al. 2001).

Temperature and oxygen concentration (Olander et al. 1998), bulk density (Auerbach et al.

1997; Olander et al. 1998), and pH (Muskett and Jones 1981; Auerbach et al. 1997; Olander

et al. 1998) are all commonly increased in roadside soils.

The effect of roads on soil nutrient availability is debated. Some studies report increased soil

nutrients in roadside areas (Lamont et al. 1994a, 1994b, Olander et al. 1998) while others

report contradictory findings (Guariguata & Dupuy 1997; Auerbach et al. 1997; Bolling &

Walker 2000). Each will be discussed separately.

In Australia, nutrients that potentially increase in roadside soil include nitrogen, potassium,

and calcium (Lamont et al. 1994b), as well as phosphorus and ammonium (Morgan 1998,

cited in Williams et al. 2001). In the UK, increased levels of nitrogen have likewise been

recorded near roads (Port & Thompson 1980). In Puerto Rico, Olander et al. (1998) found

that pools of exchangeable nutrients, except total nitrogen, were higher in recent roadfills

than in mature forest soils. Research indicated that even after 35 years, regenerating

roadfills had different soil nutrient dynamics and pool sizes in comparison to those in soils of

mature, undisturbed forests (Olander et al. 1998).

The availability of soil nutrients, including nitrogen and phosphorus, is commonly reduced in

roadside soils, with three possible explanations (Guariguata & Dupuy 1997; Auerbach et al.

1997; Bolling & Walker 2000). First, increased compaction of roadside soils may reduce

decomposition, with associated repercussions for nutrient cycling (Guariguata & Dupuy 1997;

Bolling & Walker 2000). Second, the cation-exchange capacity of soils might be altered by

road deposited dust, therefore lowering total nutrient availability (Auerbach et al. 1997). Last,


45

the disturbance common to roadside habitats may favour the growth of plant species that are

efficient at nutrient uptake, thereby depleting the total nutrient pool of the soil (Auerbach et al.

1997; Bolling & Walker 2000).

Associated impacts on vegetation

In Puerto Rico, the reduced growth of vegetation in roadfill habitats was related to increased

soil compaction and decreased nitrogen availability and organic matter (Olander et al. 1998).

These factors were also implicated in altered floristic composition of roadside communities, in

which the abundance of monocot species was increased and the recruitment of natives

significantly reduced (Olander et al. 1998). In America, the relative abundance of native and

exotic seeds germinating in abandoned roads compared to undisturbed vegetation was

likewise altered; natives germinated more frequently in undisturbed areas while exotic

species germination was greater along roads (Holl et al. 2000). In Costa Rica, reduced

densities of seed propagules were recorded in roadside soils, causing reduced regeneration

of roadside habitats (Guariguata & Dupuy 1997). In Alaska, the richness and biomass of

arctic tundra was reduced in areas within 100 m of roads (Auerbach et al. 1997).


Soil properties and soil structure directly influence the growth and survival of vegetation

communities. Consequently, any alterations to natural soil characteristics due to roads will

affect vegetation communities in roadside areas. However, the degree of impact on soil will

vary according to the road features. The impact of roads on soil nutrient status is likely to be

of particular concern for Australian ecosystems, especially the potential addition or loss of

nitrogen and phosphorus.


46

8 Road systems and their ecological impacts on landscapes ñ the ì road effectî zone


The impact of roads on landscape structure and function

Most of the preceding discussion has focused on the localised effects of roads on

surrounding environments, for example, the increased lead content of soil up to 80 m from

roads (Quarles et al. 1974) or reduced bird densities within 1.5 km of roads (Reijnen et al.

1997). However, understanding the way in which these road effects have a wider impact on

the functioning and diversity of landscapes is essential.

When addressing the effect of road networks on a landscape scale, one of the most

frequently discussed impacts is increased habitat fragmentation (Reed et al. 1996, Tinker et

al. 1998; Saunders et al. 2002). An inevitable consequence of this impact is the reduced size

of remaining habitat patches (Miller et al. 1996, Reed et al. 1996, Tinker et al. 1998). In the

US, habitat fragmentation by roads and the associated increase in edge habitat increased

the area affected by roads in Medicine Bow-Routt National Forest by 2.5ñ3.5 times the actual

area occupied by roads (Reed et al. 1996). The consequent reduction in core habitat away

from roads will have significant impacts on îinterior speciesî requiring undisturbed habitat for

survival (Tinker et al. 1998; Saunders et al. 2002).

Due to the linearity of roads, habitat patches created by roads are often simplified in shape

(Reed et al. 1996, Tinker et al. 1998; Saunders et al. 2002). This may minimise the impact of

edge effects due to the decreased edge to interior ratio (Saunders et al. 2002). However, the

impact of topography, land cover, and road juxtaposition are all likely to influence the degree

to which roads simplify patch shape (Miller et al. 1996, Saunders et al. 2002).

The spatial distribution of roads may have a greater impact on landscape structure than road

density (Tinker et al. 1998). Roads that are evenly distributed across a landscape will have a

greater impact on patch size, core area, and the quantity of edge habitat than those

occurring in a small area (Reed et al. 1996, Tinker et al. 1998). However, in the US, no

Research on the impact of roads on landscape biodiversity and structure has been undertaken in the US and Canada (Schonewald-Cox & Buechner 1992; Miller et al. 1996, Reed et al. 1996, Tinker et al. 1998; Brosofske et al. 1999; Rivard et al. 2000; Saunders et al. 2002). The concept of the ì road effectî zone has been developed in the US (Forman 1999, 2000; Forman & Hersperger 1996, Forman et al. 1997; Forman & Deblinger 1998, 2000), and applied to a native ecosystem in Australia (Williams et al. 2001). Road disturbance of birds in the wider landscape has been researched in The Netherlands (Reijnen et al. 1997).


47

correlation has been identified between road density and altered landscape structure (Tinker

et al. 1998). Nevertheless, road density may serve as an index for the level of disturbance in

a landscape, or the degree of divergence from natural conditions (Miller et al. 1996, Forman

& Hersperger 1996).

It has been suggested that the effect of roads on landscape structure may be localized, being

strongest in areas adjacent to roads (Miller et al. 1996). For example, early-seral forest

stands are more abundant than late-seral stands along roads in Roosevelt National Forest,

US (Miller et al. 1996). Road occurrence was also found to influence the distribution of some

plant species in Chequamegon National Forest, US (Brosofske et al. 1999).

In the US, paved roads substantially fragment the majority of large parks (Schonewald-Cox &

Buechner 1992). Most of the resulting fragments were small (less than 100 km≤), and the

smallest parks were most highly fragmented (Schonewald-Cox & Buechner 1992).

The ecological ì road effectî zone

The environmental effects of roads have each been addressed as separate, and essentially

isolated, impacts. However, to assess the full impact of roads on surrounding environments,

these effects must be considered together, as their interactions contribute to the total

ecological effect of roads. The concept of the îroad effectî zone incorporates all the different

impacts of roads and provides a basis for estimating the area ecologically affected by roads.

The îroad effectî zone is defined by the distance to which each different ecological road

impact extends outward (Forman 1999). These distances differ for each impact, ranging from

a few metres to over a kilometre (Forman et al. 1997). The distance to which ecological road

impacts affect adjacent environments also differs on each side of the road, with asymmetry

of slope, habitat suitability, and wind direction all modifying the effect distance (Forman et al.

1997; Forman 1999). Furthermore, the environmental effects of roads are often greater in

some locations, for example, the point of intersection between streams and roads (Forman et

al. 1997). These factors all contribute to the formation of a îroad effectî zone that varies in

width along roads, is asymmetrical, and has highly convoluted boundaries (Forman et al.

1997; Forman 1999). Forman and Deblinger (1998, 2000) calculated that the average

distance that ecological impacts of roads extend outwards is 300 m. Therefore, the îroad

effectî zone averages 600 m in width. In short, the area of land ecologically affected by roads

is significantly greater than the area covered by roads themselves (Forman et al. 1997).

In the US, the concept of a îroad effectî zone was applied to a section of highway, leading to

the estimation that 22% of the US is ecologically affected by roads (Forman & Deblinger

1998, 2000; Forman 2000). This calculation does not include a quantification of the impacts

of habitat loss and population fragmentation due to roads (Forman & Deblinger 2000).


48

Reijnen et al. (1997) estimated that, based on the disturbance distance for bird communities,

8% of The Netherlands was disturbed by roads in 1986; 11ñ16% of The Netherlands was

predicted to be potentially disturbed by roads in 2010.

In Victoria, Australia, the concept of the îroad effectî zone (with parameters relating to

American environments) was used to determine the potential impact of a highway

development on native grasslands (Williams et al. 2001). Williams et al. (2001) estimated that

the proposed development options for the highway would result in between 13% and 55% of

existing grasslands being ecologically affected by the upgraded highway. Fragmentation of

the existing grasslands and reduction of available habitat for grassland birds were identified

as impacts resulting from the highway development (Williams et al. 2001).


The concept of a îroad effectî zone enhances the understanding of the way roads affect

natural systems on a broad scale, and may affect landscape structure and function. Such

understanding is essential for effective conservation management of natural habitats. The

synthesis of the numerous impacts of roads into the concept of an ecological îroad effectî

zone allows identification of the potential impact of roads on the ecosystems they transect.

While the formula for the îroad effectî zone has been developed for American systems, it

does provide a basis for the development of similar figures applying specifically to Australian

systems. In the absence of such figures, these American parameters allow a rough

estimation of the area ecologically affected by roads in Australia.


49

9 Ecological impacts of recreational tracks


The impact of human trampling and horseriding on surrounding environments is of particular

concern in areas protected for wilderness and biodiversity values. While the effects of these

recreational activities do not affect an extensive area, they often have significant impacts in

the local track area. Vegetation communities, soil properties, and microhabitat conditions in

trackside areas are all impacted by trampling (Adkison & Jackson 1996). The disturbance

associated with recreational tracks usually only extends between 2 m and 3 m from the track

(Dale & Weaver 1974; Cole 1987, cited in Adkison & Jackson 1996).

Both the width and creation of additional tracks increase with track use (Dale & Weaver

1974; Lance et al. 1989). The quality and condition of existing tracks also influence the

potential development of new tracks (Whinam & Comfort 1996). In Tasmania, Whinam and

Comfort (1996) found that once original tracks become impassable, horseriders created

multiple new tracks to allow continued movement along trails. In Israel, track development

was significantly greater in unprotected areas than in protected areas, consequently leading

to increased habitat fragmentation (Kutiel 1999). In Scotland, the width of recreational tracks

increased by a maximum of 1.3 m over five years (Lance et al. 1989). Increased width and

density of recreational tracks will consequently increase the area affected by the following

environmental impacts.

Environmental impacts arising from recreational track and trail use by hikers and horseriders have been relatively well researched. Australian studies have reported on the impact of human trampling (Liddle & Thyer 1986, Calais & Kirkpatrick 1986, Whinam & Chilcott 1999) and horseriding (Whinam et al. 1994; Whinam & Comfort 1996, Weaver & Adams 1996, Landsberg et al. 2001) on natural ecosystems. Australian research has studied the impact of track disturbance on an endangered snake population (Goldingay 1998), and also the issue of track management (Cubit & McArthur 1995). Studies on the environmental impact of human trampling have also been undertaken in England (Burden & Randerson 1972), the US (Hall & Kuss 1989; Adkison & Jackson 1996), and Scotland (Lance et al. 1989). Research into the effect of horseriding on natural environments has been undertaken in the US (Dale and Weaver 1974) and Canada (McQuaid-Cook 1978). The pattern of track development has been studied in Israel (Kutiel 1999). Many aspects of the environmental impacts of recreational activities are addressed by Liddle (1997).


50

Impacts of human trampling

Numerous factors influence the distance to which the impacts associated with human

trampling extend from recreational tracks. These include habitat type, vegetation type, soil

properties, traffic volume, climatic conditions, season, and characteristics of track location

such as slope, drainage, and altitude (Burden & Randerson 1972; Adkison & Jackson 1996,

Whinam & Chilcott 1999). For example, alpine environments are generally considered most

susceptible to the impacts of human recreation (Whinam & Chilcott 1999). The trampling

threshold that an environment can withstand varies greatly between sites and habitats

(Calais & Kirkpatrick 1986). The impacts of trampling on surrounding environments, both to

vegetation and the soil profile, are often long lasting (Whinam & Chilcott 1999).

The vegetation directly adjacent to walking tracks is most greatly affected by hiking activity,

with reduced height and cover of vegetation being a common outcome (McQuaid-Cook 1978;

Adkison & Jackson 1996, Whinam & Chilcott 1999). Resultant loss of vegetation cover often

continues for some time after the impact (Whinam & Chilcott 1999). Some vegetation types

are more resistant to the effects of trampling than others, for example Whinam and Chilcott

(1999) determined that grasses and graminoids were more resilient than shrubs along

walking tracks in Tasmania. In America, Hall and Kuss (1989) likewise found that the

graminoid species, as well as low growing, early flowering species, were more abundant

along tracks. Weedy species often increase in trackside areas in the US (Dale & Weaver

1974). Consequently, the composition of trackside vegetation communities may be altered

due to the variable trampling resistance of different species (Adkison & Jackson 1996).

Nevertheless, a study of Tasmanian tracks found that no plant species were lost as a result

of track impacts; in fact the diversity of trackside vegetation was enhanced in some areas

(Whinam & Chilcott 1999). Likewise, in the US, increased species diversity of trackside

vegetation, caused by increased light availability, was recorded in Shenandoah National Park

(Hall & Kuss 1989). Calais and Kirkpatrick (1986) stated, however, that no vegetation

community in the Cradle Mountain-Lake St Clair National Park, Tasmania, would survive in

habitats adjacent to tracks that were used by more than 2,000 people each year.

Erosion and compaction are two of the primary impacts of trampling on trackside soil (Burden

& Randerson 1972; McQuaid-Cook 1978; Adkison & Jackson 1996). Soil properties and

trampling levels both influence the compaction of trackside soils (Burden & Randerson 1972;

McQuaid-Cook 1978). Litter layers and soil organic matter are also commonly reduced in

areas adjacent to recreational tracks (Burden & Randerson 1972; Liddle & Thyer 1986,

Adkison & Jackson 1996).


51

Impacts of horseriding

The impact of horseriding on trackside environments is influenced by a number of factors,

including degree of use, adjacent vegetation community, and track characteristics such as

slope and substrate (Whinam et al. 1994; Landsberg et al. 2001). The impacts of horseriding

on trackside environments are usually considered to be greater- and longer-lasting than

those associated with human trampling (Whinam & Chilcott 1999; Landsberg et al. 2001).

Horseriding tracks are more deeply incised than those created by human trampling (Dale and

Weaver 1974; McQuaid-Cook 1978; Whinam & Chilcott 1999). However, in Canada, soil

compaction resulting from horseriding was less than that associated with human trampling

(McQuaid-Cook 1978). Soil density and erosion are also increased along horseriding tracks

(Whinam et al. 1994; Whinam & Comfort 1996). Horse manure increases the nutrient content

and pH of trackside soil (Whinam et al. 1994).

Horses are also known to facilitate the invasion of weeds in trackside areas (Weaver &

Adams 1996, Landsberg et al. 2001). The potential for weed invasion along tracks used by

horseriders is increased due to the nutrient enrichment of soil by manure (Whinam et al.

1994).

Other effects of recreational tracks on ecosystem values

Other impacts of recreational track use on surrounding environments include lake and

stream pollution, the spread of the pathogen P. cinnamomi, and increased littering (Cubit &

McArthur 1995). In Queensland, wide walking tracks have been known to act as fire breaks

(Liddle & Thyer 1986).

The distribution of the endangered broad-headed snake (Hoplocephalus bungaroides) in

Royal National Park, Sydney is affected by the occurrence and associated disturbance of

recreational tracks (Goldingay 1998). Goldingay (1998) concluded, due to the reduced area

of undisturbed habitat in Royal National Park, that the closure of some walking tracks would

be required to ensure the survival of this species in the park.

Relevance to tracks in national parks

The impact of recreational tracks, used for either hiking or horseriding, on surrounding

environments varies widely. The habitat through which tracks pass, together with the degree

of use, are the most influential factors modifying the environmental impacts of trampling. For

example, heavily used tracks through alpine areas will almost definitely require active

management, and it is likely that many other recreational tracks will require monitoring to

ensure their impact on surrounding environments is known, and appropriately managed. The

findings of Whinam and Comfort (1996), showing the level of horseriding in sub-alpine


52

environments in Tasmania to be unsustainable, highlight the significance of track impacts in

some areas.


53

References Adams L. W. & Geis A. D. (1983) Effects of roads on small mammals. Journal of Applied

Ecology. 20, 403ñ415.

Adkison G. P. & Jackson M. T. (1996) Changes in ground-layer vegetation near trails in

midwestern US forests. Natural Areas Journal. 16(1), 14ñ23.

Alterio N. (2000) Controlling small mammal predators using sodium monofluoroacetate

(1080) in bait stations along forestry roads in a New Zealand beech forest. New Zealand

Journal of Ecology 24(1), 3ñ9.

Amor R. L. & Stevens P. L. (1976) Spread of weeds from a roadside into sclerophyll forests

at Dartmouth, Australia. Weed Research. 16, 111ñ118.

Angold P. G. (1997) The impact of a road upon adjacent heathland vegetation: effects on

plants species composition. Journal of Applied Ecology. 34(2), 409ñ417.

Anonymous (1980) Road kills of Regent Parrots at Kiama, Victoria. Bird Observer. 588, 99.

Auerbach N. A., Walker M. D. & Walker D. A. (1997) Effects of roadside disturbance on

substrate and vegetation properties in arctic tundra. Ecological Applications. 7, 218ñ235.

Baker J., Goldingay R. L. & Whelan R. J. (1998) Powerline easements through forests: a

case study of impacts on avifauna. Pacific Conservation Biology. 4, 79ñ89.

Bakowski C. & Kozakiewicz M. (1988) The effect of a forest road on Bank Vole and Yellow-

necked Mouse populations. Acta Theriologica. 33, 345ñ53.

Barnett J. L., How R. A. & Humphreys W. F. (1978) The use of habitat components by small

mammals in eastern Australia. Australian Journal of Ecology. 3, 277ñ285.

Baur A. & Baur B. (1990) Are roads barriers to dispersal in the land snail Arianta

arbustorum? Canadian Journal of Zoology. 68, 613ñ617.

Bennett A. F. (1991) Roads, roadsides and wildlife conservation: a review. In: Nature

Conservation 2: the role of corridors (Eds D.A. Saunders & R. J. Hobbs) pp 99ñ118.

Chipping Norton: Surrey Beatty and Sons.

Bilby R. E., Sullivan K. & Duncan S. H. (1989) The generation and fate of road surface

sediment in forested watersheds in southwestern Washington. Forest Science 35(2), 453ñ

468.

Bolling J. D. & Walker L. R. (2000) Plant and soil recovery along a series of abandoned

desert roads. Journal of Arid Environments. 46(1), 1ñ24.


54

Brody A.J. & Pelton M.R. (1989) Effects of roads on black bear movements in western North

Carolina. Wildlife Society Bulletin. 17, 5ñ10.

Brosofske K. D., Chen J., Crow T. R. & Saunders S. C. (1999) Vegetation responses to

landscape structure at multiple scales across a northern Wisconsin, USA, pine barrens

landscape. Plant Ecology. 143, 203ñ218.

Brothers T. S. & Spingarn A. (1992) Forest fragmentation and alien plant invasion of central

Indiana old-growth forests. Conservation Biology. 6, 91ñ100.

Brown R. J., Brown M. N. & Pesotto B. (1986) Birds killed on secondary roads in Western

Australia. Corella. 10, 118ñ122.

Burden R. F. & Randerson P. F. (1972) Quantitative studies on the effects of human

trampling on vegetation as an aid to the management of semi-natural areas. Journal of

Applied Ecology. 9, 439ñ458.

Burke R. C. & Sherburne J. A. (1982) Monitoring wildlife populations and activity along I-95 in

northern Maine before, during and after construction. Transportation Research Record. 859,

1ñ7.

Burkey T. V. (1993) Edge effects in seed and egg predation at two neotropical rainforest

sites. Biological Conservation. 66, 139ñ143.

Burnett S. E. (1992) Effects of a rainforest road on movements of small mammals:

mechanisms and implications. Wildlife Research. 19, 95ñ104.

Calais S. S. & Kirkpatrick J. B. (1986) Impact of trampling on natural ecosystems in the

Cradle Mountain-Lake St Clair National Park. Australian Geographer. 17, 6ñ14.

Carbaugh B. T., Vaughan J. P., Bellis E. D. & Graves H. B. (1975) Distribution and activity of

white-tailed deer along an interstate highway. Journal of Wildlife Management. 39, 570ñ581.

Case R. M. (1978) Interstate highway road-killed animals: a data source for biologists.

Wildlife Society Bulletin. 6, 8ñ13.

Catling P. C. & Burt R. J. (1995) Why are red foxes absent from some Eucalypt Forests in

eastern New South Wales? Wildlife Research. 22(5), 535ñ546.

Chow T. J. (1970) Lead accumulation in roadside soil and grass. Nature. 225, 295ñ296.

Claridge A. W. (1998) Use of tracks and trails by introduced predators: an important

consideration in the study of native ground-dwelling mammals. Victorian Naturalist. 115(3),

88ñ93.


55

Clark D. R. (1979) Lead concentrations: bats vs. terrestrial small mammals collected near a

major highway. Environmental Science and Technology. 13, 338ñ340.

Clarke G. P., White P. C. L. & Harris S. (1998) Effects of roads on badger Meles meles

populations in south-west England. Biological Conservation. 86(2), 117ñ124.

Clevenger A. P. & Waltho N. (2000) Factors influencing the effectiveness of wildlife

underpasses in Banff National Park, Alberta, Canada. Conservation Biology. 14(1), 47ñ56.

Clift D., Dickson I. E., Roos T., Collins P., Jolly M. & Klindworth A. (1983) Accumulation of

lead beside the Mulgrave Freeway, Victoria. Search. 14, 155ñ157.

Collins J. A. (1984) Roadside lead in New Zealand and its significance for human and animal

health. New Zealand Journal of Science. 27, 93ñ98.

Costantini A., Loch R. J., Connolly R. D. & Garthe R. (1999) Sediment generation from forest

roads: bed and eroded sediment size distributions, and runoff management strategies.

Australian Journal of Soil Research. 37(5), 947ñ964.

Coulson G. (1982) Road-kills of macropods on a section of highway in central Victoria.

Australian Wildlife Research. 9, 21ñ26.

Coulson G. (1985) Road kills: wheels vs. wildlife. Wildlife Australia. Summer 1985, 26ñ28.

Coulson G. (1989) The effect of drought on road mortality of macropods. Australian Wildlife

Research. 16, 79ñ83.

Cowie I. D. & Werner P. A. (1993) Alien plant species invasive in Kakadu National Park,

tropical northern Australia. Biological Conservation. 63, 127ñ135.

Cubit S. & McArthur S. (1995) Walking track management: only half the answer. Australian

Parks and Recreation. 31, 15ñ20.

Cunningham S. A. (2000) Effects of habitat fragmentation on the reproductive ecology of four

plant species in mallee woodland. Conservation Biology. 14, 758ñ768.

Curatolo J.A. & Murphy S.M. (1986) The effects of pipelines, roads, and traffic on the

movements of Caribou, Rangifer tarandus. Canadian Field Naturalist. 100, 218ñ224.

Dale D. & Weaver T. (1974) Trampling effects on vegetation of the trail corridors of north

Rocky Mountain Forests. Journal of Applied Ecology. 11, 769ñ772.

Dawson P. & Weste G. (1985) Changes in the distribution of Phytophthora cinnamomi in the

Brisbane Ranges National Park between 1970 and 1980ñ81. Australian Journal of Botany.

33, 309ñ315.


56

Department of Conservation and Environment. (1985) Roadworks and wetlands in the Perth

metropolitan region: recommendations for conflict resolution. Perth. Bulletin. 233.

Department of Conservation and Environment. (1992) Alpine National Park Cobberas-

Tingaringy Unit Management Plan: Department of Conservation and Environment.

Develey P. F. & Stouffer P. C. (2001) Effects of roads on movements by understorey birds in

mixed-species flocks in central Amazonian Brazil. Conservation Biology. 15(5), 1416ñ1422.

Dhindsa M. S., Sandhu J. S., Sandhu P. S. & Toor H. S. (1988) Roadside birds in Punjab

(India), relation to mortality from vehicles. Environmental Conservation. 15, 303ñ310.

Eaglin G.S. & Hubert W.A. (1993) Effects of logging roads on substrate and trout in streams

of the Medicine Bow National Forest, Wyoming. North American Journal of Fisheries

Management. 13, 844ñ846.

Ehmann H. & Cogger H. (1985) Australia's endangered herpetofauna: a review of criteria and

policies. In: Biology of Australasian frogs and reptiles (Eds G. Grigg, R. Shine & H. Ehmann)

pp 435ñ447. New South Wales: Surrey Beatty & Sons.

Fahey B. D. & Coker R. J. (1992) Sediment production from forest roads in Queen Charlotte

Forest and potential impacts on marine water quality, Marlborough Sounds, New Zealand.

New Zealand Journal of Marine and Freshwater Research. 26, 187ñ195.

Fahrig L. (2001) Animal populations and roads. Colorado. In: Proceedings of the

International Conference of Ecology and Transportation, Keystone, CO, Sept 24ñ28 2001.

Raleigh, NC: Centre for Transportation and Environment, North Carolina State University.

Fahrig L., Pedlar J. H., Pope S. E., Taylor P. D. & Wegner J. F. (1995) Effect of road traffic

on amphibian density. Biological Conservation. 73, 177ñ182.

Ferris C. R. (1979) Effects of Interstate 95 on breeding birds in northern Maine. Journal of

Wildlife Management. 43, 421ñ427.

Findlay C. S. & Bourdages J. (2000) Response time of wetland biodiversity to road

construction on adjacent lands. Conservation Biology. 14(1), 86ñ94.

Findlay C. S. & Houlahan J. (1997) Anthropogenic correlates of species richness in

southeastern Ontario wetlands. Conservation Biology. 11, 1000ñ1009.

Foppen R. & Reijnen R. (1994) The effects of car traffic on breeding bird populations in

woodland. II. Breeding dispersal of male willow warblers (Phylloscopus trochilus) in relation

to the proximity of a highway. Journal of Applied Ecology. 31, 95ñ101.


57

Forman R. T. T. (1995) Land mosaics: the ecology of landscapes. Cambridge: Cambridge

University Press.

Forman R. T. T. (1999) Horizontal processes, roads, suburbs, societal objectives, and

landscape ecology. In: Landscape Ecological Analysis: Issue and Applications (Eds: J. M.

Klopatek & R. H. Gardner) pp. 35ñ53. New York: Springer-Verlag.

Forman R. T. T. (2000) Estimate of the area affected ecologically by the road system in the

United States. Conservation Biology. 14, 31ñ35.

Forman R. T. T. & Alexander L. E. (1998) Roads and their major ecological effects. Annual

Review of Ecology and Systematics. 29, 207ñ231.

Forman R. T. T. & Deblinger R. D. (1998) The ecological road-effect zone for transportation

planning and a Massachusetts highway example. In: Proceedings of the International

Conference on Wildlife Ecology and Transportation (Eds: G. L. Evink, P. Garrett, D. Zeigler &

J. Berry) pp. 78ñ96. Tallahassee: Florida Department of Transportation.

Forman R. T. T. & Deblinger R. D. (2000) The ecological road-effect zone of a

Massachusetts (USA.) suburban highway. Conservation Biology. 14(1), 36ñ46.

Forman R. T. T., Friedman D. S., Fitzhenry J. D., Martin J. D., Chen A. S. & Alexander L. E.

(1997) Ecological effects of roads: toward three summary indices and an overview for North

America. In: Habitat fragmentation & infrastructure: proceedings (Ed: K. Canters). pp. 40ñ54.

Delft, The Netherlands: Ministry of Transport, Public Works and Water Management.

Forman R. T. T. & Hersperger A. M. (1996) Road density in different landscapes with

international planning and mitigation solutions. (Eds: G. L. Evink, P. Garrett, D. Zeigler & J.

Berry). Tallahassee, Florida: Florida Department of Transportation.

Foster M. L. & Humphrey S. R. (1995) Use of highway underpasses by Florida panthers and

other wildlife. Wildlife Society Bulletin. 23(1), 95ñ100.

Gerlach G. & Musolf K. (2000) Fragmentation of landscape as a cause for genetic

subdivision in bank voles. Conservation Biology. 14(4), 1066ñ1074.

Getz L. L., Cole F. R. & Gates D. L. (1978) Interstate roadsides as dispersal routes for

Microtus pennsylvannicus. Journal of Mammalogy. 59, 208ñ212.

Gibbs J. P. (1998) Amphibian movements in response to forest edges, roads and

streambeds in southern New England. Journal of Wildlife Management. 62, 584ñ589.

Gjessing E., Lygren E., Berlind I., Gulbrandsen T. & Skanne R. (1984) Effect of highway

runoff on lake water quality. Science of the Total Environment. 33, 247ñ257.


58

Goldingay R. (1998) Between a rock and a hard place: conserving the broad-headed snake

in Australia's oldest National Park. Proceedings of the Linnean Society of New South Wales.

120, 1ñ10.

Goldingay R. L. & Whelan R. J. (1997) Powerline easements: do they promote edge effects

in eucalypt forest for small mammals? Wildlife Research. 24, 737ñ744.

Goldsmith C. D. & Scanlon P. F. (1977) Lead levels in small mammals and selected

invertebrates associated with highways of different traffic densities. Bulletin of Environmental

Contamination and Toxicology. 17, 311ñ316.

Goldsmith C. D., Scanlon P. F. & Pirie W. R. (1976) Lead concentrations in soil and

vegetation associated with highways of different traffic density. Bulletin of Environmental

Contamination and Toxicology. 16(1), 66ñ70.

Goosem M. (2000) Effects of tropical rainforest roads on small mammals: Edge changes in

community composition. Wildlife Research. 27(2), 151ñ163.

Goosem M. & Marsh H. (1997) Fragmentation of a small-mammal community by a powerline

corridor through tropical rainforest. Wildlife Research. 24, 613ñ629.

Goosem M. W. (1997) Internal fragmentation: the effects of roads, highways and powerline

clearings on movements and mortality of rainforest vertebrates. In: Tropical rainforest

remnants: ecology, management and conservation of fragmented communities (Eds: W. F.

Laurance & R. O. Bierregaard Jr.) pp. 241ñ255. Chicago: University of Chicago Press.

Greenberg C. H., Crownover S. H. & Gordon D. R. (1997) Roadside soil: a corridor for

invasion of xeric scrub by nonindigenous plants. Natural Areas Journal. 17, 99ñ109.

Grue C. E., O'Shea T. J. & Hoffman D. J. (1984) Lead concentrations and reproduction in

highway-nesting barn swallows. The Condor. 86, 383ñ389.

Guariguata M. R. & Dupuy J. M. (1997) Forest regeneration in abandoned logging roads in

Lowland Costa Rica. Biotropica. 29(1), 15ñ28.

Hall C. N. & Kuss F. R. (1989) Vegetation alteration along trails in Shenandoah National

Park. Biological Conservation. 48, 211ñ227.

Harrison P. D. & Dyer M. I. (1984) Lead in mule deer forage in Rocky Mountain National

Park, Colorado. Journal of Wildlife Management. 48, 510ñ517.

Haskell D. G. (2000) Effects of forest roads on macroinvertebrate soil fauna of the Southern

Appalachian mountains. Conservation Biology. 14(1), 57ñ63.

Havlin J. (1987) Motorways and birds. Folia Zoologica. 36, 137ñ153.


59

Hodson N. L. (1966) A survey of road mortality in mammals (and including data for the Grass

snake and Common frog). Journal of Zoology. 148, 576ñ579.

Holl K. D., Steele H. N., Fusari M. H. & Fox L. R. (2000) Seed banks of maritime chaparral

and abandoned roads: potential for vegetation recovery. Journal of the Torrey Botanical

Society. 127(3), 207ñ220.

Huey L. M. (1941) Mammalian invasion via the highway. Journal of Mammalogy. 22, 383ñ

385.

Huijser M. P. & Bergers P. J. M. (2000) The effect of roads and traffic on hedgehog

(Erinaceus europaeus) populations. Biological Conservation. 95(1), 111ñ116.

Hunt A., Dickens H. J. & Whelan R. J. (1987) Movement of mammals through tunnels under

railway lines. Australian Zoologist. 24, 89ñ93.

Iredale S. E., Scott D. & Young F. W. (1993) The effect of vehicular traffic on the lead levels

of wayside flora in Glasgow. Journal of Biological Education. 27, 283ñ286.

Johnson H. J., Vasek F. C. & Yonkers T. (1975) Productivity, diversity and stability

relationships in Mojave Desert roadside vegetation. Bulletin of the Torrey Botanical Club.

102, 106ñ115.

Jones J. A., Swanson F. J., Wemple B. C. & Snyder K. U. (2000) Effects of roads on

hydrology, geomorphology, and disturbance patches in stream networks. Conservation

Biology. 14(1), 76ñ85.

Keller V.E. (1989) The effect of disturbance from roads on the distribution of feeding sites of

geese (Anser brachyrhynchus, A. anser), wintering in north-east Scotland. Ardea. 79, 229ñ

232.

Keyser A. J., Hill G. E. & Soehren E. C. (1998) Effects of forest fragment size, nest density,

and proximity to edge on the risk of predation to ground-nesting passerine birds.

Conservation Biology. 12(5), 986ñ994.

King J. G. & Tennyson L. C. (1984) Alteration of streamflow characteristics following road

construction in north central Idaho. Water Resources Research. 20(8), 1159ñ1163.

Knight R. L., Knight H. A. L. & Camp R. J. (1995) Common ravens and number and type of

linear rights-of-way. Biological Conservation. 74(1), 65ñ67.

Kozel R. M. & Fleharty E. D. (1979) Movements of rodents across roads. Southwestern

Naturalist. 24, 239ñ248.


60

Kuitunen M., Rossi E. & Stenroos A. (1998) Do highways influence density of land birds?

Environmental Management. 22(2), 297ñ302.

Kutiel P. (1999) Tendencies in the development of tracks in open areas. Environmental

Management. 23(3), 401ñ408.

Lagerwerff J. V. & Specht A. W. (1970) Contamination of roadside soil and vegetation with

cadmium, nickel, lead, and zinc. Environmental Science and Technology. 4, 583ñ586.

Lamont B. B., Rees R. G., Witkowski E. T. F. & Whitten V. A. (1994a) Comparative size,

fecundity and ecophysiology of roadside plants of Banksia hookeriana. Journal of Applied

Ecology. 31, 137ñ144.

Lamont B. B. & Southall K. J. (1982) Biology of mistletoe Amyema preissii on road verges

and undisturbed vegetation. Search. 13, 87ñ88.

Lamont B. B., Whitten V. A., Witkowski E. T. F., Rees R. G. & Enright N. J. (1994b) Regional

and local (road verge) effects on size and fecundity in Banksia menziesii. Australian Journal

of Ecology. 19, 197ñ205.

Lance A. N., Baugh I. D. & Love J. A. (1989) Continued footpath widening in the Cairngorn

Mountains, Scotland. Biological Conservation. 49, 201ñ214.

Landsberg J., Logan B. & Shorthouse D.. (2001) Horse riding in urban conservation areas:

reviewing scientific evidence to guide management. Ecological Management and

Restoration. 2(1), 36ñ46.

Lepschi B. J. (1992) Birds killed on a primary road in southern New South Wales. Corella

16(3), 75ñ77.

Liddle M. (1997) Recreation Ecology ñ the ecological impact of outdoor recreation and

ecotourism. London: Chapman & Hall.

Liddle M. J & Thyer N. (1986) Trampling and fire in a subtropical dry sclerophyll forest.

Environmental Conservation. 13, 33ñ39.

Lidicker W. Z. (1999) Responses of mammals to habitat edges: an overview. Landscape

Ecology. 14, 333ñ343.

Lindenmayer D. B., Pope M. L. & Cunningham R. B. (1999) Roads and nest predation: an

experimental study in a modified forest system. Emu. 99, 148ñ152.

Lonsdale W. M. & Lane A. M. (1994) Tourist vehicles as vectors of weed seeds in Kakadu

National Park, northern Australia. Biological Conservation. 69, 277ñ283.


61

Mader H. J. (1984) Animal habitat isolation by roads and agricultural fields. Biological

Conservation. 29, 81ñ96.

Mader H.J. (1988) The significance of paved agricultural roads as barriers to ground dwelling

arthropods (Ed: K. F. Schreiber) pp. 97ñ100. Munster: Munstersche Geographische

Arbeiten.

Mader H. J., Schell C. & Kornacker P. (1990) Linear barriers to arthropod movements in the

landscape. Biological Conservation. 54, 209ñ222.

Mansergh I. M. & Scotts D. J. (1989) Habitat continuity and social organization of the

mountain pygmy-possum restored by tunnel. Journal of Wildlife Management. 53, 701ñ707.

Martin A. A. & Tyler M. J. (1978) The introduction into Western Australia of the frog

Limnodynastes tasmaniensis Gunther. Australian Zoologist. 19(3), 321ñ325.

Maurer R. (1974) The beetle and spider fauna of meadows affected by traffic pollution.

Oecologia. 14, 327ñ351.

May S. A. & Norton T. W. (1996) Influence of fragmentation and disturbance on the potential

impact of feral predators on native fauna in Australian forest ecosystems. Wildlife Research.

23, 387ñ400.

McLellan B. N. & Shackleton D. M. (1988) Grizzly bears and resource-extraction industries:

effects of roads on behaviour, habitat use and demography. Journal of Applied Ecology. 25,

451ñ460.

McQuaid-Cook J. (1978) Effects of hikers and horses on mountain trails. Journal of

Environmental Management. 6, 209ñ212.

Mech L. D., Fritts S. H., Radde G. L. & Paul W. J. (1988) Wolf distribution and road density in

Minnesota. Wildlife Society Bulletin. 16, 85ñ87.

Miller J. R., Joyce L. A., Knight R. L. & King R. M. (1996) Forest roads and landscape

structure in the southern Rocky Mountains. Landscape Ecology. 11, 115ñ127.

Motto H. L., Daines R. H., Chilko D. M. & Motto C. K. (1970) Lead in soils and plants: its

relationship to traffic volume and proximity to highways. Environmental Science and

Technology. 4, 231ñ238.

Murcia C. (1995) Edge effects in fragmented forests: implications for conservation. Trends in

Ecology & Evolution. 10, 58ñ62.

Muskett C. J. & Jones M. P. (1981) Soil respiratory activity in relation to motor vehicle

pollution. Water, Air, & Soil Pollution. 23, 231ñ242.


62

Nellemann C., Vistnes I., Jordhoy P. & Strand O. (2001) Winter distribution of wild reindeer in

relation to power lines, roads and resorts. Biological Conservation. 101(3), 351ñ360.

Noller B. N. & Smythe L. E. (1974) The distribution of lead in vegetation bordering roads in

Sydney metropolitan area. Search. 5, 108ñ110.

Norton D. A. & Stafford Smith M. (1999) Why might roadside mulgas be better mistletoe

hosts? Australian Journal of Ecology. 24, 193ñ198.

Olander L. P., Scatena F. N. & Silver W. L. (1998) Impacts of disturbance initiated by road

construction in a subtropical cloud forest in the Luquillo experimental forest, Puerto Rico.

Forest Ecology & Management. 109, 33ñ49.

O'Neill D. H., Robel R. J. & Dayton A. D. (1983) Lead contamination near Kansas highways:

implications for wildlife enhancement programs. Wildlife Society Bulletin. 11, 152ñ160.

Ortega Y. K. & Capen D. E. (1999) Effects of forest roads on habitat quality for Ovenbirds in

a forested landscape. Auk. 116, 937ñ946.

Osawa R. (1989) Road-kills of swamp wallaby, Wallabia bicolor, on North Stradbroke Island,

south-east Queensland. Australian Wildlife Research. 16, 95ñ104.

Oxley D. J., Fenton M. B. & Carmody G. R. (1974) The effects of roads on populations of

small mammals. Journal of Applied Ecology. 11, 51ñ59.

Parendes L. A. & Jones J. A. (2000) Role of light availability and dispersal mechanisms in

invasion of exotic plants roads and streams in the H. J. Andrews Experimental Forest,

Oregon. Conservation Biology. 14, 64ñ75.

Parks Victoria. (1997) Brisbane Ranges National Park Management Plan. Parks Victoria.

Pienaar UdV. (1968) The ecological significance of roads in a national park. Koedoe. 11,

169ñ174.

Port G. R. & Thompson J. R. (1980) Outbreaks of insect herbivores on plants along

motorways in the United Kingdom. Journal of Applied Ecology. 17, 649ñ656.

Preuss P. (1989) Forest roadworks pose threat to rare possum. Conservation News. July, 9.

Quarles H. D., Hanawalt R. B. & Odum W. E. (1974) Lead in small mammals, plants, and soil

at varying distances from a highway. Journal of Applied Ecology. 11, 937ñ949.

Reed D. F. (1981) Mule behaviour at a highway underpass exit. Journal of Wildlife



63

Reed R. A, Johnson-Barnard J & Baker W. L. (1996) The contribution of roads to forest

fragmentation in the Rocky Mountains. Conservation Biology. 10, 1098ñ1106.

Reh W. & Seitz A. (1990) The influence of land use on the genetic structure of populations of

the common frog Rana temporaria. Biological Conservation. 54, 239ñ249.

Reid L. M. & Dunne T. (1984) Sediment production from forest road surfaces. Water

Resources Research. 20, 1753ñ1761.

Reid N. (1997) Control of the mistletoes by possums and fire: a review of the evidence.

Victorian Naturalist. 114(3), 149ñ158.

Reijnen R. & Foppen R. (1994) The effects of car traffic on breeding bird populations in

woodland. I. Evidence of reduced habitat quality for willow warblers (Phylloscopus trochilus)

breeding close to a highway. Journal of Applied Ecology. 31, 85ñ94.

Reijnen R., Foppen R., Ter Braak C. & Thissen J. (1995) The effects of car traffic on

breeding bird populations in woodland. III. Reduction of density in relation to the proximity of

main roads. Journal of Applied Ecology. 32, 187ñ202.

Reijnen R., Foppen R. & Veenbaas G. (1997) Disturbance by traffic of breeding birds:

Evaluation of the effect and considerations in planning and managing road corridors.

Biodiversity & Conservation. 6(4), 567ñ581.

Rich A. C., Dobkin D. S. & Niles L. J. (1994). Defining forest fragmentation by corridor width:

the influence of narrow forest-dividing corridors on forest-nesting birds in southern New

Jersey. Conservation Biology 1109ñ1121.

Richardson J. H., Shore R. F. & Treweek J. R. (1997) Are major roads a barrier to small

mammals? Journal of Zoology (London). 243, 840ñ846.

Rivard D. H., Poitevin J., Plasse D., Carleton M. & Currie D. J. (2000) Changing species

richness and composition in Canadian national parks. Conservation Biology. 14(4), 1099ñ

1109.

Rost G. R. & Bailey J. A. (1979) Distribution of Mule Deer and Elk in relation to roads.

Journal of Wildlife Management. 43, 634ñ641.

Rudolph D. C., Burgdorf S. J. & Conner R. N. (1999) Preliminary evaluation of the impact of

roads and associated vehicular traffic on snake populations in eastern Texas. The

International Conference on Wildlife Ecology and Transportation III


64

Ruediger B. & Ruediger B. (1999) The effects of highways on trout and salmon rivers and

streams in the western U. S. The International Conference on Wildlife Ecology and

Transportation III

Rummer B., Stokes B. & Lockaby G. (1997) Sedimentation associated with forest road

surfacing in a bottomland hardwood ecosystem. Forest Ecology & Management. 90(2ñ3),

195ñ200.

Saunders S. C., Mislivets M. R., Chen J. & Cleland D. T. (2002) Effects of roads on

landscape structure within nested ecological units of the Northern Great Lakes Region, USA.

Biological Conservation. 103(2), 209ñ225.

Scheidt M. E. (1971) Environmental effects of highways. In: Man's Impact on Environment

(Ed: T. R. Detwyler) pp. 419ñ427. New York: McGraw-Hill Book Company.

Schmidt W. (1989) Plant dispersal by motor cars. Vegetatio. 80, 147ñ152.

Schonewald-Cox C. & Buechner M. (1992) Park protection and public roads. In:

Conservation biology. The theory and practice of nature conservation, preservation and

management (Eds: P. L. Fiedler & S. K. Jain) pp. 373ñ395. New York: Chapman and Hall.

Schreiber R. K. & Graves J. H. (1977) Powerline corridors as possible barriers to the

movements of small mammals. American Midland Naturalist. 97, 504ñ508.

Seabrook W. A. & Dettman E. B. (1996) Roads as activity corridors for cane toads in

Australia. Journal of Wildlife Management. 60, 363ñ368.

Sherburne J. (1983) Wildlife populations utilizing right-of-way habitat along Interstate 95 in

Northern Maine. Transportation Research Record. 1016, 16ñ20.

Singer F. (1978) Behavior of mountain goats in relation to US Highway 2, Glacier National

Park, Montana. Journal of Wildlife Management. 42, 591ñ597.

Small M. F. & Hunter M .L. (1988) Forest fragmentation and avian predation in forested

landscapes. Oecologia. 76, 62ñ64.

Spencer H. J. & Port G. R. (1988) Effects of roadside conditions on plants and insects II. Soil

conditions. Journal of Applied Ecology. 25, 709ñ714.

Stahlecker D. W. (1978) Effects of new transmission line on wintering prairie raptors. The

Condor. 80, 444ñ446.

Stiles J. H. & Jones R. H. (1998) Distribution of the red imported fire ant, Solenopsis invicta,

in road and powerline habitats. Landscape Ecology. 335, 335ñ346.


65

Swihart R. K. & Slade N. A. (1984) Road crossing in Sigmodon hispidus and Microtus

ochrogaster. Journal of Mammalogy. 65, 357ñ360.

Thiel R. P. (1985) Relationship between road densities and habitat suitability in Wisconsin.

American Midland Naturalist. 113, 404ñ407.

Thrasher M. H. (1983) Highway impacts on wetlands: assessment, mitigation, and

enhancement measures. Transportation Research Record. 948, 17ñ20.

Thurber J. M., Peterson R. O., Drummer T. D. & Thomasma S. A. (1994) Gray wolf response

to refuge boundaries and roads in Alaska. Wildlife Society Bulletin. 22, 61ñ68.

Tinker D. B., Resor C. A. C., Beauvais G. P., Kipfmeuller K. F., Fernandes C. I. & Baker W.L.

(1998) Watershed analysis of forest fragmentation by clearcuts and roads in a Wyoming

forest. Landscape Ecology. 13, 149ñ165.

Travis M. D. & Tilsworth T. (1986) Fish passage through Poplar Grove Creek Culvert, Alaska.

Transportation Research Record. 1075, 21ñ26.

Trombulak S. C. & Frissell C. A. (2000) Review of ecological effects of roads on terrestrial

and aquatic communities. Conservation Biology. 14, 18ñ30.

Tyser R. W. & Worley C. A. (1992) Alien flora in grasslands adjacent to road and trail

corridors in Glacier National Park, Montana (USA.). Conservation Biology. 6, 253ñ262.

Tyson R. M. (1980) Road-killed platypus. Tasmanian Naturalist. 60, 8.

U. S. Department of Health, Education, and Welfare. (1971) Sources of air pollution. In:.

Man's Impact on Environment (Ed: T.R. Detwyler) pp. 91ñ97. New York: McGraw-Hill Book

Company.

van Schagen J. J., Hobbs R. J. & Majer J. D. (1992) Defoliation of trees in roadside corridors

and remnant vegetation in the Western Australian wheatbelt. Journal of the Royal Society of

Western Australia 75, 75ñ81.

Vestjens W. J. M. (1973) Wildlife mortality on a road in New South Wales. Emu. 73, 107ñ

112.

Wace N. (1977) Assessment of dispersal of plant species ñ the car-borne flora in Canberra.

Proceedings of the Ecological Society of Australia. 10, 167ñ186.

Weaver V. & Adams R. (1996) Horses as vectors in the dispersal of weeds into native

vegetation. Eleventh Australian Weeds Conference Proceedings, pp. 383ñ397. Weed

Science Society of Australia, Victoria, Australia.


66

Wemple B. C., Jones J. A. & Grant G. E. (1996) Channel network extension by logging roads

in two basins, western Cascades, Oregon. Water Resources Bulletin. 32, 1ñ13.

Weste G. & Taylor P. (1971) The invasion of native forest by Phytophthora cinnamomi I.

Brisbane Ranges. Australian Journal of Botany. 19, 281ñ294.

Whinam J., Cannell E. J., Kirkpatrick J. B. & Comfort M. (1994) Studies on the potential

impact of recreational horseriding on some alpine environments of the Central Plateau,

Tasmania. Journal of Environmental Management. 40, 103ñ117.

Whinam J. & Chilcott N. (1999) Impacts of trampling on alpine environments in central

Tasmania. Journal of Environmental Management. 57(3), 205ñ220.

Whinam J. & Comfort M. (1996) The impact of commercial horse riding on sub-alpine

environments at Cradle Mountain, Tasmania, Australia. Journal of Environmental


Whitford P. C. (1985) Bird behaviour in response to the warmth of blacktop roads.

Transactions of the Wisconsin Academy of Sciences, Arts and Letters. 73, 135ñ143.

Wilkins K. T. (1982) Highways as barriers to rodent dispersal. South West Naturalist. 37,

459ñ460.

Williams N. S. G., Leary E. J., Parris K. M. & McDonnell M. J. (2001) The potential impact of

freeways on native grassland. Victorian Naturalist. 118(1), 4ñ15.

Williams-Linera G. (1990) Vegetation structure and environmental conditions of forest edges

in Panama. Journal of Ecology. 78, 356ñ373.

Willson M. F. & Chrome F. H. J. (1989) Patterns of seed rain at the edge of a tropical

Queensland rain forest. Journal of Tropical Ecology. 5, 301ñ308.

Wilson B. A., Aberton J. & Cahill D. M. (2000) Relationships between site factors and

distribution of Phytophthora cinnamomi in the Eastern Otway Ranges, Victoria. Australian

Journal of Botany. 48(2), 247ñ260.

Wilson J. B., Rapson G. L., Sykes M. T., Watkins A. I. & Williams P. A. (1992) Distributions

and climatic correlations of some exotic species along roadsides in South Island, New

Zealand. Journal of Biogeography. 19, 183ñ194.

Wylie P.B. & Bell L.C. (1973) The effect of automobile emissions on the lead content of soils

and plants in the Brisbane area. Search. 4, 161ñ162.

Yanes M., Velasco J. M. & Suarez F. (1995) Permeability of roads and railways to

vertebrates: the importance of culverts. Biological Conservation. 71(3), 217ñ222.

Parks Victoria is responsible for managing the Victorian protected

area network, which ranges from wilderness areas to metropolitan

parks and includes both marine and terrestrial components.

Our role is to protect the natural and cultural values of the parks

and other assets we manage, while providing a great range of

outdoor opportunities for all Victorians and visitors.

A broad range of environmental research and monitoring activities

supported by Parks Victoria provides information to enhance park

management decisions. This Technical Series highlights some of

the environmental research and monitoring activities done within

Victoria’s protected area network.

Healthy Parks Healthy People

For more information contact the Parks Victoria Information Centre

on 13 1963, or visit www.parkweb.vic.gov.au

Documents

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