A NEW AGENDA FOR
BIOSECURITY
Jeff K. Waage, Rob W. Fraser, John D. Mumford,
David C. Cook and Andy Wilby
Faculty of Life Sciences, Imperial College London
August, 2004
A New Agenda for Biosecurity, August 2004
Contents
CHAPTER 1 – STUDY BACKGROUND AND OBJECTIVES ............................. 8
1.1 GOVERNMENT POLICY ON NON-NATIVE SPECIES RISKS ....................................... 10
1.2 THE NATURE AND SCALE OF NON-NATIVE SPECIES RISK ..................................... 11
1.3 OBJECTIVES OF THE STUDY................................................................................. 12
CHAPTER 2 – AN ECOLOGICAL MODEL FOR NON-NATIVE SPECIES
INTRODUCTIONS ................................................................................................... 14
2.1 CONCEPTUAL MODEL ......................................................................................... 14
2.1.2 Formalisation ......................................................................................... 15
2.1.3 Behaviour and validity of the biological spread model ......................... 16
2.2 BIOLOGICAL PATTERNS OF NON-NATIVE SPECIES INTRODUCTION ....................... 18
2.2.1 Terrestrial invertebrates ........................................................................ 19
2.2.2 Plant diseases......................................................................................... 23
2.2.3 Vertebrates ............................................................................................. 27
2.2.4 Animal diseases ...................................................................................... 28
2.2.5 Terrestrial plants ................................................................................... 32
2.2.6 Aquatic species....................................................................................... 36
2.3 CONCLUSIONS FROM CROSS-TAXON REVIEW ...................................................... 41
2.3.1 Origin of non-native species risks.......................................................... 42
2.3.2 Nature and impact of non-native species ............................................... 43
CHAPTER 3 – AN ECONOMIC MODEL FOR NON-NATIVE SPECIES
INTRODUCTION...................................................................................................... 45
3.1 FORMALISATION ................................................................................................. 45
3.1.1 Graphical representation ....................................................................... 47
3.2 STOCHASTIC SIMULATION .................................................................................. 52
3.3 PARAMETERISATION ........................................................................................... 53
3.4 DEALING WITH NON-MARKET (E.G. ENVIRONMENTAL) FACTORS ........................ 55
CHAPTER 4 – ECONOMIC CASE STUDIES ...................................................... 56
4.1 COLORADO POTATO BEETLE .............................................................................. 57
4.1.2 Control case ........................................................................................... 58
4.1.3 Results .................................................................................................... 60
4.1.4 Conclusion ............................................................................................. 62
4.2 WILD BOAR ........................................................................................................ 63
4.2.1 Affected industries in the United Kingdom ............................................ 64
4.2.2 Control case ........................................................................................... 64
4.2.3 Results .................................................................................................... 66
4.2.4 Conclusion ............................................................................................. 68
4.3 POTATO RING ROT ............................................................................................. 69
4.3.1 Affected industries in the United Kingdom ............................................ 70
4.3.2 Control case ........................................................................................... 70
4.3.3 Results .................................................................................................... 71
4.3.4 Conclusion ............................................................................................. 73
4.4 NEWCASTLE DISEASE ......................................................................................... 74
4.4.1 Affected industries in the United Kingdom ............................................ 75
4.4.2 Control case ........................................................................................... 75
4.4.3 Results .................................................................................................... 78
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A New Agenda for Biosecurity, August 2004
4.4.4 Conclusions ............................................................................................ 81
4.5 GYRODACTYLUS SALARIS................................................................................... 81
4.5.1 Affected aquaculture industries in the United Kingdom ........................ 82
4.5.2 Control case ........................................................................................... 82
4.5.3 Results .................................................................................................... 85
4.5.4 Conclusion ............................................................................................. 87
4.6 CREEPING THISTLE ............................................................................................. 88
4.6.1 Affected industries in the United Kingdom ............................................ 88
4.6.2 Control case ........................................................................................... 89
4.6.3 Results .................................................................................................... 91
4.6.4 Conclusions ............................................................................................ 94
CHAPTER 5 – PATTERNS OF IMPACT OF NON-NATIVE SPECIES ........... 95
5.1 COMPARING IMPACT ESTIMATES BETWEEN SPECIES ........................................... 95
5.2 COMPARING PATTERNS OF IMPACT OVER TIME ................................................... 96
5.2.1 Constant expected impact increments over time .................................... 96
5.2.2 Diminishing expected impact increments over time .............................. 97
5.2.3 Increasing expected impact increments over time ................................. 98
5.3 CROSS-OVER EFFECTS AND VARIABILITY .......................................................... 101
CHAPTER 6 – HORIZON SCANNING AND IMPACT LEVELS .................... 104
6.1 CLIMATE CHANGE ............................................................................................ 106
6.2 TRADE AND MARKETS ...................................................................................... 109
6.2.1 A conceptual model for trade and introduction ................................... 109
6.2.2 The effect of trade on UK agriculture and land use ............................ 112
6.3 SOCIAL ISSUES .................................................................................................. 115
6.3.1 What is a non-native species? .............................................................. 116
6.3.2 New species and societal change ......................................................... 117
6.3.3 Which way will the future go? ............................................................. 117
6.4 CONCLUSIONS .................................................................................................. 118
6.5 A QUANTITATIVE APPROACH TO HORIZON SCANNING..................................... 119
6.5.1 Description ........................................................................................... 120
6.5.2 Affected Industries in the United Kingdom .......................................... 120
6.6.3 Control Case ........................................................................................ 121
6.5.4 A Trade Change Scenario ................................................................... 129
6.5.5 A CAP Reform Scenario....................................................................... 131
6.5.6 Conclusion ........................................................................................... 133
CHAPTER 7 – PREVENTION AND ERADICATION OF NON-NATIVE
SPECIES THREATS ............................................................................................... 134
7.1 PREVENTION AND ERADICATION STRATEGIES – AN OVERVIEW ......................... 135
7.2 ERADICATION ................................................................................................... 135
7.2.1 Eradication, net benefit maximisation and EDcrit .............................. 138
7.2.2 Multiple net benefit maximisation options ........................................... 139
7.3 PREVENTION ..................................................................................................... 140
7.4 EVALUATING PREVENTION VS. ERADICATION POLICY OPTIONS......................... 142
7.5 MULTIPLE TECHNOLOGICAL OPTIONS ............................................................... 145
7.6 TECHNICAL CHANGE AND HOW TO VALUE IT .................................................... 146
CHAPTER 8 – CONCLUSIONS ............................................................................ 149
8.1 ARE BIOSECURITY RISKS INCREASING? ............................................................. 149
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A New Agenda for Biosecurity, August 2004
8.2 CAN WE TAKE A GENERAL APPROACH TO PREDICTING THE ECONOMIC IMPACT OF
FUTURE INTRODUCTIONS? ....................................................................................... 150
8.3 ARE SOME KINDS OF RISK CONSISTENTLY MORE IMPORTANT THAN OTHERS? ... 152
8.4 ARE FUTURE SOCIETAL TRENDS GOING TO CHANGE RISK SUBSTANTIALLY? ..... 152
8.5 CAN WE PRIORITISE INVESTMENT IN CONTROL METHODS? ............................... 153
8.6 HOW CAN POLICY MAKERS USE THIS STUDY? ................................................... 153
REFERENCES ......................................................................................................... 155
APPENDIX 1: FINITE MARKOV CHAINS........................................................ 162
APPENDIX 2: NON-INDIGENOUS SPECIES IN THE UK: EXPLORING
THEIR MEANINGS IN HUMAN AND SOCIAL TERMS ................................ 163
ACKNOWLEDGEMENTS ............................................................................................ 163
EXECUTIVE SUMMARY ............................................................................................ 163
1.0 BACKGROUND TO THE RESEARCH .................................................................... 164
2.0 THE HISTORICAL AND SOCIAL CONTEXT ........................................................... 164
3.0 CONCEPTIONS OF NON INDIGENOUS SPECIES AMONGST DIFFERENT ACTORS .... 166
3.1 Mapping out different discourses............................................................ 166
3.2 What do the different discourses tell us? ................................................ 168
4.0 ISSUES OF DEFINITION: ‘NON-INDIGENOUS’, ‘NON-NATIVE’, ‘ALIEN’, AND
‘INVASIVE’ SPECIES ................................................................................................ 170
4.1 Native/non-native definitions .................................................................. 171
4.2 Invasiveness ............................................................................................ 172
4.3 Alienness ................................................................................................. 173
4.4 Definitions: a need for clarity? ............................................................... 174
5.0 PUBLIC PERCEPTIONS ....................................................................................... 175
5.1 The supposed problem of ‘lay ignorance’ .............................................. 176
5.2 Perceptions of non-indigenous species in the UK .................................. 177
5.3 Public Engagement ................................................................................. 178
5.4 Thinking about the future ........................................................................ 179
6.0 CONCLUSIONS AND RECOMMENDATIONS ......................................................... 181
BIBLIOGRAPHY ....................................................................................................... 183
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A New Agenda for Biosecurity, August 2004
Figures
Figure 1: The annual number of newspaper articles on non-native species problems from UK broadsheets presented
as (a) total numbers and (b) numbers per 10,000 science/environment articles. Analysis used the Factiva database,
searching for keywords combining “non-native, alien, invasive or exotic” with different plant/animal groups.
Newspapers: Guardian, Times, Independent, Independent on Sunday, Financial Times, Observer, Sunday Times. ... 9
Figure 2.1: A conceptual model of non-native species invasion. The main processes are identified: 1) Arrival and
establishment; 2) Local population growth and spread; and 3) Satellite generation, and the biological/geographical
(red) and anthropocentric (blue) controllers............................................................................................................... 15
Figure 2.2: The influence of intrinsic growth rate (r) and satellite generation rate () on total area occupied after 30
years. Diffusion coefficient fixed at 40 ha/yr. ............................................................................................................ 17
Figure 2.3: First recordings of non-native pestiferous arthropod species in Europe (data derived from Smith 1997).
Fitted line represents the best fit poisson model (log [species per decade] = 0.04yr – 75.6; P = 0.004; r2 = 0.63)...... 21
Figure 2.4: First recordings of non-native plant diseases in Europe (including bacteria, fungi and nematodes; data
derived from Smith 1997). Fitted line represents the best fit linear regression model (y=-52.7 + 0.028x, P = 0.006, r2
=0.54). ...................................................................................................................................................................... 25
Figure 2.5: Decade of first introduction (a) and first record in the wild (b), of neophyte taxa in the UK and Ireland
(compiled from Preston et al., 2002). Lines are fit by local non-parametric regression. ............................................ 34
Figure 2.6: Frequency histogram of the lag between year of first import and year of first record in the wild for
established (naturalised) plant taxa in the UK and Ireland (compiled from Preston et al., 2002). ............................... 34
Figure 2.7: Arrival of non-native aquatic species to the UK as collated in the FAO DIAS database. ......................... 38
Figure 2.8: Rate of arrival of non-native species to the Baltic Sea due to deliberate stocking, species associated with
deliberate stocking, and accidental introductions due to shipping. ............................................................................. 39
Figure 2.9: Rate of arrival of species to the Baltic Sea according to taxon. ............................................................... 39
Figure 2.10: Median radial spread rates of the six groups included in this study. Bars represent maxima and minima.
................................................................................................................................................................................. 44
Figure 3.1: The production function with and without a harmful non-native species in the system ............................ 48
Figure 3.2: The economic impact of a harmful non-native species – imported goods ............................................... 49
Figure 3.3: The economic impact of a harmful non-native species – exported goods ............................................... 51
Figure 4.1: Cumulative distribution of the critical level of Expected Damage (ED crit) over 20 years – Colorado Potato
Beetle ....................................................................................................................................................................... 60
Figure 4.2: Area/Time and variability – Colorado Potato Beetle................................................................................ 60
Figure 4.3: Expected Invasion Impact (EI)/Time – Colorado Potato Beetle............................................................... 61
Figure 4.4: Cumulative distribution of the critical level of Expected Damage (EDcrit) over 20 years – Wild Boar ........ 66
Figure 4.5: Area/Time and variability – Wild Boar..................................................................................................... 66
Figure 4.6: Expected Invasion Impact (EI)/Time – Wild Boar .................................................................................... 67
Figure 4.7: Cumulative distribution of the critical level of Expected Damage (ED crit) over 20 years – Potato Ring Rot
................................................................................................................................................................................. 72
Figure 4.8: Area/Time and variability – Potato Ring Rot ........................................................................................... 72
Figure 4.9: Expected Invasion Impact (EI)/Time – Potato Ring Rot .......................................................................... 72
Figure 4.10: Cumulative distribution of the critical level of Expected Damage (EDcrit) over 20 years – Newcastle
Disease .................................................................................................................................................................... 78
Figure 4.11: Incidence/Time and variability – Newcastle Disease ............................................................................ 78
Figure 4.12: Expected Invasion Impact (EI)/Time – Newcastle Disease ................................................................... 79
Figure 4.13: Cumulative distribution of the critical level of Expected Damage (ED crit) over 20 years – Gyrodactylus
salaris ....................................................................................................................................................................... 85
Figure 4.14: Incidence/Time – Gyrodactylus salaries ............................................................................................... 85
Figure 4.15: Expected Invasion Impact (EI)/Time – Gyrodactylus salaris ................................................................. 86
Figure 4.16: Cumulative distribution of the critical level of Expected Damage (EDcrit) over 20 years – Creeping Thistle
................................................................................................................................................................................. 91
Figure 4.17: Area/Time – Creeping Thistle ............................................................................................................... 92
Figure 4.18: Expected Invasion Impact (EI)/Time – Creeping Thistle ....................................................................... 92
Figure 5.1: Constant expected impact increments over time .................................................................................... 97
Figure 5.2: Decreasing expected impact increments over time................................................................................. 98
Figure 5.3: Increasing expected impact increments over time ................................................................................ 100
Figure 5.4: Cross-over effects ................................................................................................................................ 102
Figure 5.5: Cross over effects with variance included. ............................................................................................ 103
Figure 6.1: A species pool model in which the pool of potential non-native species is defined by the action of abiotic,
biotic, trade and transport constraints ..................................................................................................................... 110
Figure 6.2: Patterns of first wild record of naturalised non-native plant species in the UK and Ireland. Curves are
fitted non-parametric cubic B-splines (3 d.f.). .......................................................................................................... 112
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A New Agenda for Biosecurity, August 2004
Figure 6.3: Cumulative distribution of the critical level of Expected Damage (ED crit) over 20 years for the control case
– Foot & Mouth Disease ......................................................................................................................................... 126
Figure 6.4: Incidence/Time and variability – Foot & Mouth Disease......................................................................... 127
Figure 6.5: Expected Invasion Impact (EI)/Time – Foot & Mouth Disease ............................................................... 127
Figure 6.6: Cumulative distribution of the critical level of Expected Damage (ED crit) over 20 years in the trade change
scenario – Foot & Mouth Disease ........................................................................................................................... 130
Figure 6.7: Cumulative distribution of the critical level of Expected Damage (ED crit) differential between the control
case and the trade change scenario over 20 years – Foot & Mouth Disease .......................................................... 130
Figure 6.8: Cumulative distribution of the critical level of Expected Damage (ED crit) over 20 years in the CAP reform
scenario – Foot & Mouth Disease ........................................................................................................................... 132
Figure 6.9: Cumulative distribution of the critical level of Expected Damage (ED crit) differential between the control
case and the CAP reform scenario over 20 years – Foot & Mouth Disease............................................................. 132
Figure 7.1: Benefits of management effort – e.g. an export-limiting disease ........................................................... 137
Figure 7.2: Eradication of an incursion ................................................................................................................... 138
Figure 7.3: Eradication or strategic management? ................................................................................................. 140
Figure 7.4: Pre-invasion biosecurity measures ....................................................................................................... 141
Figure 7.5: No solution through eradication ............................................................................................................ 142
Figure 7.6: Total Expected Benefits of a hypothetical prevention technology for Colorado Potato Beetle ............... 143
Figure 7.7: Expected Net benefit stream for prevention technology ........................................................................ 143
Figure 7.8: Distribution of the present value of net benefits for the prevention option. ........................................... 144
Figure 7.9: Alternative management technologies .................................................................................................. 146
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A New Agenda for Biosecurity, August 2004
Tables
Table 2.1: Non-native species used as case studies for model design ..................................................................... 14
Table 2.2: Parameter values found in the literature for the spread of non-native terrestrial invertebrate species. ...... 21
Table 2.3: Parameter values found in the literature for the spread of plant diseases. Here r refers to the intrinsic rate
of increase of the infected population of the host species. Since these are agricultural host species, this rate of
increase of infection relates readily to spatial extent of infection. ............................................................................... 26
Table 2.4: Parameter values found in the literature for the spread of non-native terrestrial vertebrate species .......... 28
Table 2.5: Parameter values found in the literature for the spread of animal diseases ............................................. 32
Table 2.6: Parameter values obtained form the literature relating to the spread non-native plant species. † - data
concern post-glacial spread of species as estimated from pollen analyses ............................................................... 34
Table 2.7: Data from Grosholz (1996) on the rate of spread of ten marine species. .................................................. 41
Table 3.1: Semi-Quantifiable Risk Categorisation Methodology (AFFA, 2001) .......................................................... 55
Table 4.1: Industries affected by Colorado Potato Beetle .......................................................................................... 58
Table 4.2: Parameterisation – Control Case (Colorado Potato Beetle) ...................................................................... 58
Table 4.3: Sensitivity Analysis – Colorado Potato Beetle........................................................................................... 62
Table 4.4: Industries affected by Wild Boar ............................................................................................................... 64
Table 4.5: Parameterisation – Control Case (Wild Boar) ........................................................................................... 65
Table 4.6: Sensitivity Analysis – Wild Boar................................................................................................................ 68
Table 4.7: Industries affected by Potato Ring Rot ..................................................................................................... 70
Table 4.8: Parameterisation – Control Case (Potato Ring Rot).................................................................................. 71
Table 4.9: Sensitivity Analysis – Potato Ring Rot ...................................................................................................... 73
Table 4.10: Industries affected by Newcastle Disease .............................................................................................. 75
Table 4.11: Parameterisation – Control Case (Newcastle Disease)........................................................................... 76
Table 4.12: Sensitivity Analysis – Newcastle Disease ............................................................................................... 80
Table 4.13: Industries affected by Gyrodactylus salaris ............................................................................................. 82
Table 4.14: Parameterisation – Control Case (Gyrodactylus salaris) ......................................................................... 84
Table 4.15: Sensitivity Analysis – Gyrodactylus salaris ............................................................................................. 87
Table 4.16: Industries affected by Creeping Thistle ................................................................................................... 88
Table 4.17: Parameterisation – Control Case (Creeping Thistle) ............................................................................... 90
Table 4.18: Sensitivity Analysis – Creeping Thistle ................................................................................................... 93
Table 5.1: Estimated impacts on 20 year time horizons for different species (from case studies). ............................. 95
Table 6.1: The relationship between model processes, their drivers and response ................................................. 104
Table 6.2: Principal predictions of the UKCIP02 (Hulme et al. 2002), and their hypothesised influence on model
parameters. Confidence level: High, medium or low, is a qualitative assessment of the reliability of these predictions
given by UKCIP. ..................................................................................................................................................... 107
Table 6.3: Survey of the ISI publications database with search terms: climate change and (species invasions or
species range). The entries don’t represent a review of literature related to the effect of climate change on life-history
parameters of species in general, but in particular those referring to invasive species and changes in geographical
range ...................................................................................................................................................................... 108
Table 6.4: Industries affected by Foot & Mouth Disease ......................................................................................... 120
Table 6.5: British Beef Exports (Tonnes) ................................................................................................................. 123
Table 6.6: Parameterisation – Control Case............................................................................................................ 124
Table 6.7: Sensitivity Analysis – Foot & Mouth Disease .......................................................................................... 128
Table 6.8: Parameterisation – Climate Change Scenario (Foot & Mouth Disease). The arrow indicates parameters
which are increased in the scenario. ....................................................................................................................... 129
Table 6.9: Parameterisation – Trade Liberalisation Scenario (FMD)........................................................................ 131
Table 8.1: Impact Table .......................................................................................................................................... 151
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A New Agenda for Biosecurity, August 2004
Chapter 1 – Study Background and Objectives
There is a long history of the deliberate and accidental movement of animal
and plant species around the world. Introduced, non-native species, which are
sometimes also called alien or exotic species, may not establish permanent
populations and may not be harmful even if they do. A small proportion of
established non-native species become problems, invading managed or
natural ecosystems, causing harm to agriculture and/or the environment.
The UK already has many non-native species problems, some of which are
longstanding and widespread. Public and government concern about future
introductions of new, harmful, non-native species has grown as a result of
recent introductions, including foot and mouth disease (FMD), potato ring rot,
and sudden oak death. There is also new and growing public concern about
established non-native species, e.g. rhododendron, mink, because of their
emerging impact on environmental conservation.
Media coverage gives an indication of how interest in and/or coverage of nonnative species problems has changed in recent years. In Figure 1 we show
the number of newspaper articles on non-native species problems in the UK
broadsheets between 1991 and 2004, presented as (a) total articles and (b)
articles per 10,000 articles in the same science/environment subject area. By
either representation, there has been a dramatic increase in UK press
coverage.
45
40
35
30
(a)
25
20
15
10
5
0
1991 1992 1993 1994 1995 1996 1997 1998 1999 2000 2001 2002 2003
8
A New Agenda for Biosecurity, August 2004
12
10
8
6
(b)
4
2
0
1991 1992 1993 1994 1995 1996 1997 1998 1999 2000 2001 2002 2003
Figure 1: The annual number of newspaper articles on non-native species problems from UK
broadsheets presented as (a) total numbers and (b) numbers per 10,000
science/environment articles. Analysis used the Factiva database, searching for
keywords combining “non-native, alien, invasive or exotic” with different
plant/animal groups. Newspapers: Guardian, Times, Independent, Independent
on Sunday, Financial Times, Observer, Sunday Times.
The biological range of possible new biosecurity risks facing the UK is
considerable – literally from microbes to mammals. The experience of other
countries, specifically with non-native species problems which are not yet in
UK, indicates that there are substantial potential risks for the near future.
Further, there is a widely held belief that increasing global trade and travel is
amplifying risk for all countries, indeed that biological invasion is a potential
“sting in the tail” of a liberalised global economy.
We refer here to the prevention and management of non-native species risks
as a form of biosecurity, while acknowledging that this term has had other
interpretations (e.g. active prevention of animal disease at the farm level).
With every potentially harmful species that may be introduced into UK, the
government has four possible biosecurity options:




do nothing, i.e. allow its introduction and establishment
attempt to prevent its introduction
attempt to eradicate it when it arrives, so that it does not establish or
make a commitment to recurrent, continuing control to suppress it.
When a non-native species problem has become widely established and
“chronic”, its management usually becomes a private responsibility. Chemical
pest control or vaccination of animal herds are two examples where producers
bear the cost of an established non-native species problem. Hence, the
principle role of government is to allocate resources to prevention of new
problems and eradication of problems when they occur, before they become
permanently established.
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A New Agenda for Biosecurity, August 2004
Government will therefore want to know the identity and nature of future nonnative species risks, and to estimate the nature and magnitude of that risk, so
as to anticipate and allocate resources efficiently between them. The sheer
diversity of potential risks, the difficulty of predicting potential harm for any one
species and the speed and stealth of many species in establishing and
spreading makes this quite difficult. This project develops some tools to make
this easier.
1.1 Government policy on non-native species risks
In most countries, different kinds of non-native species problems are dealt
with independently by different government departments. Most effort has been
put into protecting agricultural production from non-native pests and diseases.
For the protection of agricultural crops, a generally efficient and well-tested
national and international regulatory structure has evolved, involving import
restrictions, quarantine, eradication and other measures. National agriculture
departments share information on new threats but make individual decisions,
and formal risk analysis has become a feature of this system in recent years.
A somewhat different system has evolved for the prevention and management
of livestock diseases, based on an internationally agreed list of diseases to be
excluded, and national commitment to eradicate new introductions. Fisheries
and forestry may have their own systems, depending on the country.
Finally, with the recent and growing appreciation of the threats which nonnative species pose to the environment, departments concerned with
conservation may establish programmes of prevention and management.
Most countries, therefore, have a “patchwork” pattern of responsibility for
biosecurity involving several departments. Departmental remits are
sometimes overlapping, and frequently there are historical “gaps” in coverage.
Until recently, for instance, non-native species problems in marine systems
were poorly monitored or managed at the national or international level. The
international trade in garden plants and in pets also represent areas of risk
which have not been well covered because they have fallen outside traditional
departmental remits.
As we understand more about harmful non-native species, we appreciate that
many specific problems are multi-sectoral. Introduced species, for instance,
may have both a measurable agricultural and environmental impact. Grey
squirrel in UK, for instance, causes substantial economic losses to timber
production, but is also responsible for the decline of the native red squirrel, a
wildlife conservation impact (Sheail 1999; Wouters et al. 2000). Non-native
plants may be weeds of both agriculture and natural habitats, in aquatic
environments weeds may affect fishing, navigation, water supply, irrigation,
tourism and biodiversity. Animal diseases may affect livestock and wildlife,
indeed many may be zoonotic, affecting human health as well.
The patchiness of biosecurity coverage between sectors, combined with the
growing multisectoral nature of the problem has led governments today to
consider a more joined-up biosecurity approach, coordinated across all
sectors and departments. In the USA, New Zealand and Australia, interministerial bodies of varying structure and power have been created for this
purpose. Similar national coordination is proposed by the Defra Non-Native
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A New Agenda for Biosecurity, August 2004
Species Review (2003). The recent closer integration of agricultural and
environmental ministries in some countries, as in UK with the creation of
Defra, creates a special opportunity for a coordinated approach.
1.2 The nature and scale of non-native species risk
The threats posed by non-native species are very real and very significant.
From an agricultural perspective, the immediate losses to production from a
new pest or disease are today sometimes vastly compounded by restrictions
placed on export of the affected commodity. This problem is likely to grow
worse. As the WTO liberalises trade, non-tariff trade barriers will become even
more important: countries may use the presence of an unwanted pest in
another country as a basis for excluding certain imports from there. Currently
the WTO is considering a number of challenges relating to risks of non-native
species introductions. Even without a formal process of trade exclusion,
importers may simply react to possible non-native species risks by switching
sources, as we have seen recently in the impact of bird flu outbreaks in Asia
and chicken imports from there.
From an environmental perspective, governments have recognised the threat
of global species exchange in the Convention on Biological Diversity (1992)
and its Article 8h which required parties to “prevent the introduction of, control
or eradicate those alien species which threaten ecosystems, habitats or
species”. In 1996, a UN conference brought together 80 governments and
world experts to consider the nature of this problem, and concluded that it was
“immense, insidious, increasing and irreversible”. This conference, and
subsequent consultations identied non-native species invasions as second
only to habitat loss as the major threat to biodiversity and species extinction.
On an international level, it has been relatively easy to develop broad political
consensus on non-native species problems, as they constitute a shared,
external threat which affects many economies and constituencies in a similar
manner. Unfortunately, this consensus is still underpinned largely by
anecdotal, rather than precise information about non-native species impact
and risk. Much recent literature on non-native species problems (e.g. Bright
1998, Baskin 2002) has had the deliberate and admirable objective of raising
awareness of the global nature of this problem, and it has had the desired
impact, simply because of the sheer number of biological invasion “horror
stories” which can be accumulated. But its anecdotal nature means it is of little
value to building an evidence base and tools for rationale policy decisions.
Policy makers have been influenced by broad-brush economic assessments
which give big numbers but little useful information. Most significant has been
work in the USA by OTA (1993) and Pimentel (2000), the latter famously
estimating that the cot to the USA of non-native species and their control was
$137b. The US Executive Order of 2000, by which President Clinton
established an inter-ministerial Invasive Species Council, was effectively built
around these statistics. These estimates have been constructed by summing
cost data from a series of specific case studies in the literature. Most case
studies were necessarily superficial, involving sometimes arbitrary economic
estimates. Estimates were then scaled up to a national level with simple
multipliers. As costs include both losses and costs of control, and it is not
clear that the problem was one of non-natives per se or inappropriate or
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A New Agenda for Biosecurity, August 2004
ineffective response to them. Pimentel et al. (2002) have since extended their
approach to estimate a global annual loss of $1.5 trillion. Once more, in the
absence of better information, such studies have been important in raising
awareness of the potential magnitude of the problem, which is no doubt
enormous, but they do not constitute a good platform for development of
policy.
Non-native species introductions may cause today economic and social
problems which are different than those anticipated by our current biosecurity
policy, as shown by the 2001 FMD epidemic, where the action taken to control
FMD proved a much greater burden on the economy than the impact of the
disease itself on the livestock industry. The FMD experience has exposed
how much our biosecurity policy is rooted in a post-war food security agenda
which may no longer be relevant.
Finally, there is the impression that the problem of non-native species is
increasing, or getting out of control, perhaps due to liberalised trade. The
press coverage presented in Figure 1 would appear to support the idea of a
growing problem. In fact, the great majority of the approximately 100 species
referred to in these newspaper articles are long-established non-natives. The
growth in press coverage may reflect the spread of these non-natives, but
probably not an increase in new non-native species establishments. The
widespread presumption that non-native species risk is growing remains to be
substantiated.
Let us summarise the policy-relevant issues surrounding non-native species
problems. The problem is probably growing, and probably bigger than
previously thought, if we include emerging evidence of environmental threats.
However, the evidence base for both the magnitude of the problem and its
growth is poor: scientific evidence is often anecdotal and economic evidence
is at best approximate, if nonetheless concerning. Our impressions of real
economic costs may be distorted by “old thinking” (e.g. about national food
security). The public is increasingly interested and aware of the problem, but
not necessarily well informed of the real present and future risk. Finally, the
government is addressing this risk in different ways in different departments,
with overlaps, gaps and a potential for more “joined-upness”.
1.3 Objectives of the study
In this scoping study, we develop an approach which can address many of the
policy issues identified above. The specific objectives of this scoping study
are:
1. To generate a broad understanding of the nature, diversity, rate and
pathways for the introduction of key biological threats.
2. To develop a modelling approach which predicts the potential impact of
new biosecurity threats over specified time horizons, based on a
“government do nothing” premise.
3. To consider, in a horizon scanning context, how likely future changes in
society and the environment will affect predictions of impact, using the
model to generate different scenarios.
12
A New Agenda for Biosecurity, August 2004
4. To develop an approach for evaluating the benefits and costs of
government action against non-native species, including prevention or
eradication, in the context of model predictions on potential costs.
In order to address the policy challenges reviewed above, we have chose to
make this project strongly inter-sectoral and inter-disciplinary. Hence we build
a conceptual approach applicable to all kinds of non-native species threats to
the rural economy and environment: microbial, plant and animal, terrestrial
and aquatic. The model is interdisciplinary, incorporating both ecological and
economic information and theory, and is put into a stochastic framework so as
to reveal both the magnitude and variability of predicted impacts.
13
A New Agenda for Biosecurity, August 2004
Chapter 2 – an Ecological Model for Non-native
Species Introductions
In order to design an ecological model for non-native species introductions,
case studies were compiled for a range of different invasive, non-native
species in the UK or in ecologically similar regions. Species were selected so
as to provide a broad evidence base for the construction of a general
conceptual model. Several species were selected from each of five taxonomic
groups: animal diseases, plant diseases, terrestrial vertebrates, terrestrial
invertebrates and plants. Literature searches provided information on
introduction, the process of spread and ecological and economic impact. The
species studies are listed in Table 2.1, and the results of this analysis are
incorporated into Section 2.2.
Table 2.1: Non-native species used as case studies for model design
Vertebrates
Animal
diseases
Plants
Plant diseases
Terrestrial
invertebrates
Muntjac deer
(Muntiacus
reevesi)
West Nile
Virus
Japanese knotweed
(Fallopia japonica)
Sudden oak death
(Phytophthora
ramorum)
Knopper gall
(Andricus
quercuscalicis)
Canada goose
(Branta
Canadensis)
Newcastle
disease
Himalayan balsam
(Impatiens
glandulifera)
Potato brown rot
(Ralstonia
solanacearum)
New Zealand
flatworm
(Arthurdendyus
triangulates)
Grey squirrel
(Sciurus
carolinensis)
Foot and
Mouth
disease
American Willowherb (Epilobium
ciliatum)
Karnal bunt (Tilletia
indica)
Asian longhorned
beetle (Anoplophora
glabripennis)
Ruddy Duck
(Oxyura
jamaicensis)
Rinderpest
Giant hogweed
(Heracleum
mantegazzianum)
Plum pox potyvirus
(PPV)
Tobacco whitefly
(Bemisia tabaci)
Coypu
(Myocaster
coypus)
Classical
Swine Fever
Rhododendron
(R. ponticum)
Colorado beetle
(Leptinotarsa
decemlineata)
Sika deer
(Cervus nippon)
2.1 Conceptual model
The conceptual model arising from this exercise is used as a framework for
our research into the invasion process (Figure 2.1). It identifies three principle
processes in the biology of non-native species invasions: arrival and
establishment; local population growth and spread; and the seeding of new
satellite populations. It also highlights how biological, geographical and
anthropogenic factors control these processes within an invasion.
This framework forms the basis of a simulation model for species invasions
and impact. The challenges in developing this model were to capture the
general ecological processes of species invasion, while maintaining relevance
to a broad a range of types of invasive organisms. Later, in Chapter 3 we will
14
A New Agenda for Biosecurity, August 2004
integrate this model with an economic model to predict the economic impact
over time of future non-native species introductions.
REGIONAL /
NATIONAL BOUNDARY
Dispersal constraints
(lo
ng
-ra
ng
e
)
Distribution of habitat
(s
ho
r
Land-use change
New sub-population
t-r
a
ng
e

)

Human facilitated
movement
Non-native population
Geographical
factors:
Population
growth and
spread
Land bridges
Air / Ocean
currents
Abiotic constraints

Translocation
of organisms
or vectors
Climate
Climate change
Disturbance
Land management
change
Geology (soils)
Biotic constraints
Predation
Competition
Community
compositional change
Mutualisms
Resource organisms
Figure 2.1: A conceptual model of non-native species invasion. The main processes are
identified: 1) Arrival and establishment; 2) Local population growth and spread;
and 3) Satellite generation, and the biological/geographical (red) and
anthropocentric (blue) controllers.
In this model and chapter we focus on species which do establish and spread,
which we stress are a very small fraction of total introductions (Williamson
1996).
2.1.2 Formalisation
Arrival and establishment are formalised in the simulation model as the simple
probabilities of entry ( Pent ) and establishment ( Pest ) which are combined to give
the probability of invasion Pi 1:
pi  pent  pest
where
0  pi  1
(1).
Modelling of species invasions has been an active sector of ecological theory
for many years. Skellam (1951) employed reaction diffusion models, originally
developed by Fisher (1937) to model the spread of mutant genotypes in
populations, to the spread of muskrat populations in Europe. These models
are of the general form:
 d 2n d 2n 
dn
 f (n)  D 2  2 
dt
dy 
 dx
(2).
1
The quantitative models of following sections use Markov chain models to estimate
transitional probabilities between time periods. See Appendix 1 for and explanation.
15
A New Agenda for Biosecurity, August 2004
where f (n) is the population growth function and D is the diffusion coefficient.
A generic result of these models is that a population diffusing from a point
source will eventually reach a constant asymptotic radial spread rate of 2 rD
in all directions, where r is the population’s intrinsic growth rate. We have
adopted this well-studied model of spread to simulate the local postestablishment population spread of a non-native species. Hence we assume
that once established the population is naturalised and it spreads by a
diffusive process such that area occupied by the population expands following
the function (Hengeweld, 1989; Lewis 1997; Shigesada and Kawasaki 1997):
At  4 Drt 2
(3).
Where At is the area occupied at time t ; D is the population diffusion
coefficient; r is the intrinsic rate of population growth. It is assumed that the
population is in a homogenous environment and expands at an equal rate in
each direction.
Equation (3) allows prediction of the spread of a species on the basis of an
estimated intrinsic rate of population growth, while an estimate of D can be
derived from the Mean Dispersal Distance ( MDD ) (Andow et al., 1990):
i.e.
D
2MDD 2
(4).

Area alone is not sufficient for the purposes of an economic impact
assessment since the density of the invasive species population within that
area influences the control measures required to counter the impacts of the
invasive species. Therefore, the model also assumes that in each unit of area
occupied by the expanding population, the local population density ( N ) grows
logistically to the carrying capacity of the environment ( K ) such that:
N
K
 K

1  
 1e  rt
 N min

(5).
Here, N min is the size of the original influx (usually assumed to be 1).
As the area involved in an initial site expansion and the population density
within that area increases, so too does the likelihood of a random satellite
outbreak some distance from the original site:
Psat  A
(6).
Here  is the rate of satellite generation and A is the occupied area. Satellite
populations grow and expand in the same manner as the original population.
Total occupied area of the original site and satellites grows until A  Amax
(maximum habitable area), at which point total area remains constant.
2.1.3 Behaviour and validity of the biological spread model
The constant rate of increase in radial range predicted by Fisher-Skellam type
reaction diffusion suggests that Area should increase linearly with time
(area increases exponentially). There have been extensive studies to test this
16
A New Agenda for Biosecurity, August 2004
prediction against recorded spread of non-native species. Shigesada and
Kawasaki (1997) discuss the forms that the relationship between Area and t
can take with reference to past non-native species invasions: Type 1 – a
linear increase; Type 2 – biphasic increase with an initial slow slope followed
by a steep linear slope; and Type 3 - an accelerating non-linear curve with the
gradient increasing with time. Our model allows simulation of Type 1 curves (
 = 0) and type 3 expansions (  > 0). We do not explicitly simulate type 2
curves though we do not see this as a major drawback since type 2 and 3
patterns are often difficult to distinguish with real data (Shigesada and
Kawasaki 1997), and both are postulated to result from satellite generation.
Area (1000s km 2)
If we look at the spread and satellite generation parts of the model in isolation,
we can investigate the sensitivity of the rate of spread to the parameters r , D
and  once introduction and establishment stages have been surpassed.
Figure 2.2 shows a response surface of total area occupied at a fixed point in
time after the invasion began in relation to r and  . Note that D can be
substituted for r in these analyses since r and D have equal influence on the
rate of diffusive spread (see equation 3).
200
150
100
50
0
0
0.5
1
1.5
2
2.5
0
2
10
8
6
4
 x 10-6
3
r
Figure 2.2: The influence of intrinsic growth rate (r) and satellite generation rate ( ) on total
area occupied after 30 years. Diffusion coefficient fixed at 40 ha/yr.
This suggests that long distance satellite dispersal can have a marked effect
on the rate of spread of an invading organism, particularly in organisms with a
high intrinsic growth rate or with large diffusion coefficients. Data from
documented spread of non-native species show that long distance dispersal,
and deviation from the predictions of pure reaction – diffusion models, is a
common feature of invasive species spread. Long distance dispersal is
sometimes anthropogenic, involving secondary distribution by human
agencies within the non-native range, and can also be a natural (e.g. when a
species has two-phase dispersal).
General patterns can be observed in empirical studies of range expansion by
non-native species. Firstly, many species of mammal, bird, insect and plant
do exhibit Type 1 dynamics suggesting that that the Fisher-Skellam model is
an appropriate framework (Shigesada and Kawasaki 1997). Agassiz (1994)
reviewed the spread of 25 species of Lepidoptera (butterflies and moths)
which invaded Britain in the last century and found that 21 species exhibited
Type 1 spread dynamics. Secondly, Type 2 or 3 dynamics are sometimes
exhibited by birds (Mundinger and Hope 1982; Hengeweld 1989), insects
17
A New Agenda for Biosecurity, August 2004
(Andow 1993) and plants (Mack 1981) and this is usually attributed to log
distance or ‘jump’ dispersal due to natural or anthropogenic processes. This
effect is simulated in our model by a non-zero value for the satellite generation
parameter  . It has been noted that Type 2 and 3 curves can also be
generated if there is evolutionary adaptation of the non-native species in the
new environment (Hengeweld 1989; Travis and Dytham 2002).
There have been several studies which have aimed to compare spread in
observed invasions with the spread rate predicted by the reaction – diffusion
model. Generally such validation is done by local measurements of r and D
and comparing the model output with larger scale records of the rate of
spread. A good match between theory and observation have been found with
a broad range of species (see Shigesada and Kawasaki 1997 chapter 3)
including muskrats, black death and rabies in Europe (Nobel 1974; Yachi et
al. 1989), the cabbage white butterfly and the sea otter in the United States
(Andow 1990; Lubina and Levin 1989). A failure to account for long-distance
dispersal is the principal cause of under-estimation of the true spread rates,
such as in the case of the cereal leaf beetle (Andow et al. 1990).
One practical advantage of the diffusion modelling framework is that the
spatial spread is a function of the intrinsic population growth rate r and the
diffusion coefficient D . r is the instantaneous rate of population growth that
can be derived from measurements of population size over time (ln[pop size
time t+1 / pop size time t]). In the absence of data of this type, rmax can be
estimated from the fecundity and longevity of the species, which though giving
an indication of potential population growth (and spread) under ideal
conditions only allows us to set an upper limit to r in the field. However,
scaling relationships of rmax to body size may allow estimation of broad
differences2 in r between taxonomic groups.
2.2 Biological patterns of non-native species introduction
The general model for non-native species introductions allows us to compare
different taxa with respect to basic parameters which describe arrival,
establishment and spread. In this chapter, we survey these taxa and examine
the evidence base for their introduction and spread, which might serve as a
basis for future prediction. We chose six groups: the five taxonomic groups
identified in Table 2.1, in which we consider only terrestrial species, and a
sixth group comprising aquatic species. This latter category includes all taxa
in freshwater and marine systems.
For each group, we first give a summary of patterns of introduction, spread
and impact. We then present evidence on introduction and spread, drawn
from a survey of the literature. For data on introduction, we have relied largely
on existing databases which give first records of species in Britain or Europe.
2 Estimates of probability of entry can be obtained from records of interception at ports. However, interceptions are
usually associated with particular commodities or shipment methods, which may not represent all of the pathways of
introduction for a particular species. Other problems identified above also apply to interpretation of interception data,
e.g. variation in sampling and prevention effort. Nonetheless, it may be possible and informative to correlate specific
patterns of entry (interception) and establishment (or outbreak) for species where introduction pathways are limited,
measurable and regularly inspected
18
A New Agenda for Biosecurity, August 2004
These databases are few and of varying quality. They do not permit separate
estimation of entry and establishment, only an estimate of new species
actually established over a time period2. It is also important to note that
interpretation of this data is made difficult where effort has varied over time –
we are generally more aware of new establishments today for all taxa than we
were a century ago. Establishment rates may also reflect the intensity of
preventative measures and how these have changed over time. Hence a
falling rate of establishment may not mean a falling rate of introduction, but
perhaps a growing effectiveness of prevention measures.
2.2.1 Terrestrial invertebrates
Terrestrial invertebrates include particularly arthropod pests of crops and of
livestock, as well as a range of species which may affect the natural
environment. Insect pests of crops have a long history of international spread,
and like plant diseases, national risk assessment and quarantine procedures
for such species are well developed. The International Plant Protection
Convention (IPPC) and regional plant protection organisations, like the
European Plant Protection Organization (EPPO), provide global support to the
development of protocols and sharing of information.
Most new introductions of terrestrial invertebrates are accidental, involving
transport of undetected individuals on agricultural materials (e.g. in soil or as
eggs on plants) or other objects (e.g. as insect larvae in wood used for
packing, as free insects in aircraft cabins). Establishment and spread of
invertebrates is particularly sensitive to abiotic factors such as temperature.
Climate. For instance, Colorado Potato Beetle (see Chapter 4.1) has ravaged
most of continental Europe but spreads to UK only through produce, and may
not thrive in the UK climate, if ever established (Baker et al. 2000).
There is a long historical tradition of invertebrate exchange between Europe
and North America, relating to agricultural trade. While this continues, for
instance in the recent introduction of corn rootworm, Diabrotica virgifera
virgifera, this trend may be declining in favour of other temperate sources (see
Chapter 6.2.1). It is also noteworthy than many new crop pests in UK in recent
decades have been of subtropical origin, as pests of protected cultivation,
including mites, leafminers, aphids, thrips and moths. Climate change
(Chapter 6.1) may increase the impact of such pests outside glasshouses.
Trade in horticultural plants, as opposed to food crops, have been major
pathways of recent introduction, According to the Royal Horticultural Society,
a record five new garden insect pests arrived in Britain in the one year 2002
(The Independent 20 Jan 2003). Many of these organisms will now enter via
continental Europe, where horticultural propagation for UK is done. There has
also been recently a distinctive pattern of Eurasian species extending their
range across Europe into Britain, including tree pests like oak knopper gall
(Andricus quercuscalicus (Walker et al. 2002), spruce beetle (Dendroctonus
micans) and horse chestnut leafminer (Cameraria ohridella).
Outside of direct agricultural impact, non-native invertebrates may have a
substantial environmental impact. For instance, the knopper gall wasp can
greatly reduce acorn and oak production (Hails and Crawley 1991), while the
19
A New Agenda for Biosecurity, August 2004
New Zealand flatworm eliminates earthworms from soils, with their important
decomposition function (Blackshaw 1997).
From the perspective of animal and human disease, the movement of disease
vectors, e.g. mosquitoes and midges, is of growing concern, particularly with
global warming. Aquatic larval stages permit accidental transport in
containers: the mosquito vector of introduced West Nile virus in USA, Aedes
albopictus, was itself introduced on shipments of tires from Asia (Craig 1993).
New crop pests and new species of insects in groups commonly recorded by
naturalists, e.g. moths, are probably well reported and capable of providing an
evidence base on the pattern of introduction. The Central Science
Laboratories have recently completed a database of insect introductions into
UK (R. Baker, pers. comm.). Some, at least partial, databases, exist from
other countries. However, there are probably a large number of nonagricultural invertebrates species whose history and pattern of introduction is
unknown, including many soil inhabiting invertebrates.
2.2.1.1 Patterns of introduction
Our analysis on the arrival of patterns of terrestrial invertebrate species is
derived from the EPPO/CABI database (Smith et al. 1997). These data
include pestiferous species which on the quarantine lists of EPPO which are
likely to cause economic damage to European nations. This therefore
represents a biased sample of terrestrial invertebrates, but we know of no
available database which includes non-pestiferous or environmental pest
species
In order to give a sufficient sample size we tabulated year of first
record of species entering any European country, rather than just the UK.
The data indicate that the rate of new incidence of pestiferous arthropods into
Europe increased exponentially to the end of the last century. However, the
rate in the final decade of the last century was much reduced, possibly
indicative of saturation or improved prevention, or simply of delayed reporting.
Predicted rates for the current decade, assuming that the fitted trend
continues suggest establishment of between five and ten new species (Figure
2.3).
20
4
3
2
0
1
Arthropod species per decade
5
6
A New Agenda for Biosecurity, August 2004
1900
1920
1940
1960
1980
2000
Figure 2.3: First recordings of non-native pestiferous arthropod species in Europe (data
derived from Smith 1997). Fitted line represents the best fit poisson model (log
[species per decade] = 0.04yr – 75.6; P = 0.004; r2 = 0.63).
A recent study by the National Audit Office (2003) of plant pest and disease
prevention in England and Wales suggests that outbreaks of new pest and
disease species are rising in recent years. Outbreaks represent occurrences
in the country of local populations of non-native pest species, many of which
will be eradicated, or may not establish anyway. Hence, outbreaks will be
more frequent than establishments, but the pattern is nonetheless striking:
from 1993 to 2001, outbreaks per year fluctuated around 150, but in 2001 they
leapt to 350. NAO notes that, in recent years, however, only one insect pest
has established, the Western Flower Thrip, Frankliniella occidentalis. A rising
number of outbreaks does not reflect rising establishment, particularly if
prevention and eradication are effective. However, putting all this information
together, it is likely that rates of introduction are rising and rates of new
species establishment are either static or rising.
2.2.1.2 Patterns of spread
Terrestrial invertebrates are one of the best-studied groups of invasive
species, and we have a relatively large data collection of data on their rates
and patterns of spread following naturalisation (Table 2.2). Even so, relatively
few studies have measured individual population parameters of the invading
organism in addition to the overall rate of spread. For many of this group the
Area increases linearly with time, suggesting that the diffusion model fits the
data adequately (satellite creation can be assumed to be zero).
Table 2.2: Parameter values found in the literature for the spread of non-native terrestrial
invertebrate species.
Common
name
Name
r
(year)

D
2
(km
yr-1)
-
(km
2 -1
yr )
Predicted
Radial
expansion
(km yr-1)
Observed
Radial
expansion
(km yr-1)
A
(km2
yr-1)
Source
21
A New Agenda for Biosecurity, August 2004
Japanese
beetle
Popillia japonica
Gypsy
moth
Lymantria dispar
Rice
water
weevil
Lissorhoptrus
oryzophilus
Small
white
butterfly
Pieris rapae
Tetse fly
Glossina sp.
4.6
9-32
2.464
9.3-90
5.5, 27.5
(Shigesada et
al.
1995;
Shigesada
and Kawasaki
1997)
9.45
(Liebhold
al. 1992)
28
(Andow et al.
1993)
15-170
(Andow et al.
1993)
2-25
et
(Hargrove
2000)
Etainia decentella
3.05
(Agassiz
1994)
Stigmella
suberivora
2.2
{Agassiz,
1994 #147
Psychoides
filicivora
1.69
(Agassiz
1994)
Caloptilia
rufipennella
6.71
(Agassiz
1994)
Parocystola
acroxantha
7.33
(Agassiz
1994)
Agrolamprotes
micella
0.056
(Agassiz
1994)
Teleiodes
alburnella
3.89
(Agassiz
1994)
Blastobasis lignea
3.72
(Agassiz
1994)
Blastobasis
decolorella
2.2
(Agassiz
1994)
Cacoecimorpha
pronubana
2.26
(Agassiz
1994)
Ptycholomoides
aeriferanus
2.31
(Agassiz
1994)
Epiphyas
postvittana
2.54
(Agassiz
1994)
Adoxophes orana
1.07
(Agassiz
1994)
Lozotaeniodes
formosanus
2.65
(Agassiz
1994)
Acleris abietana
4.12
(Agassiz
1994)
Pammene
aurantiana
2.43
(Agassiz
1994)
Phlyctaenia
perlucidalis
4.68
(Agassiz
1994)
Dioryctria
schuetzeella
5.36
(Agassiz
1994)
Xanthorhoe
biriviata
1.81
(Agassiz
1994)
Spargania luctuata
1.24
(Agassiz
1994)
22
A New Agenda for Biosecurity, August 2004
Cereal
leaf
beetle
Eupithecia
phoeniceata
2.14
(Agassiz
1994)
Peribatodes
secundaria
3.46
(Agassiz
1994)
Hadena compta
2.54
(Agassiz
1994)
Lithophane
leautieri
3.78
(Agassiz
1994)
Polychrysia.
Moneta
2.99
(Agassiz
1994)
Phyllonorycter
leucographella
10.3
(Agassiz
1994)
Phyllonorycter
platani
8.63
(Agassiz
1994)
14.7-170
(Andow et al.
1990)
Oulema melanopus
1.61.9
0.4
13-127
The agricultural species cited in this table, towards the top, can show very
rapid spread, making eradication difficult. Both flight and movement on
transported crops or other materials can result in rapid production of satellite
populations. Fortunately, most current agricultural outbreaks are local, often in
protected habitats where control is relatively easy. The long list of Agassiz
(1994) are for moths invading more natural habitats. Movement is slower, but
populations are probably more difficult to find and contain.
2.2.2 Plant diseases
Plant diseases are caused by a range of organisms, including bacteria,
nematodes, fungi, viruses and mycoplasm-like organisms. The introduction of
plant diseases into UK is, as in most countries, principally associated with the
accidental importation of infected plant material or infected soil. Like insect
pests, diseases of agricultural crops and their movements are covered by
established national inspection systems and risk analysis procedures. EU
regulation applies increasingly, and IPPC and EPPO contribution to
international protocols and monitoring. Arrival of contaminated plant material
is frequent, and therefore inspection and interception is extremely important.
Improved capacity to trace origin and movement of infected material, as
applied in the recent potato ring rot outbreak, greatly facilitates the prevention
and early eradication of outbreaks.
Establishment of plant diseases in a country may restrict exports of the
affected crop to other countries, depending upon the nature of the disease.
Hence, plant disease and trade are often entangled, and plant diseases may
be used to justify non-tariff trade barriers.
Plant diseases and their host range are very labile, and new strains with new
properties and host ranges evolve continuously through mutation and
hybridisation. There is an agricultural tradition of dealing with the appearance
and global movement of new plant disease strains through biosecurity
measures and plant breeding for resistance. Governments are generally less
well prepared for new plant diseases which threaten native flora. Dutch elm
disease, Ophiostoma ulmi, is a classic case of such an “environmental plant
disease”, where movement and evolution led to the emergence of a new,
23
A New Agenda for Biosecurity, August 2004
devastating strain on native trees in different continents (Brasier 2001). More
recently, sudden oak death, Phytophthora ramorum, which affects both
horticultural plants (e.g. rhododendron) and native trees (Quercus spp.) has
emerged in a similar manner. In the case of plant diseases, movement of
infected plant material and inoculum not only spreads a disease but puts a
pathogen in an environment where it may evolve new forms. For instance, it is
believed now that a hybrid between a strawberry and woodland Phytophthora
spp. has created a new, deadly canker disease of alder (C. Prior, pers
comm.). Recent flooding in Europe may have encouraged the mixing The
emergence of such “new diseases” of non-agricultural plants poses particular
challenges for rapid detection, diagnosis and management in the extensive
system of natural forests or amenity plantings. The rapid European spread of
the alder Phytophthora, for instance, occured in part because of a lack of
inspection and of sufficient research into control and measures to quickly
miminise dispersal. (C. Prior, pers comm.).
The emergence and spread of devastating diseases of native and amenity
trees in North America, Europe and other regions is a continuing phenomenon
which appears to be related both to the introduction of non-native
species/strains and to the way in which this has stimulated local evolution of
new strains.
2.2.2.1 Patterns of introduction
Like invertebrates, our information on the arrival of plant diseases is derived
from the EPPO/CABI database (Smith et al. 1997). These data include
species which on the quarantine lists of EPPO which are likely to cause
economic damage to European nations. Again, year of first record in Europe
was analysed for fungal, bacterial and nematode diseases. The data indicate
that the rate of establishment of new diseases of crops in Europe has
increased steadily over the past century. Predicted rates for the current
decade, equate to approximately three to five new species per decade (Figure
2.4). The comments made above in Section 2.2.1 for agricultural insect pest
outbreaks apply as well to plant diseases (NAO 2003), outbreaks are
increasing, but not necessarily establishments.
24
3
2
1
0
species per decade
4
A New Agenda for Biosecurity, August 2004
1900
1920
1940
1960
1980
2000
year
Figure 2.4: First recordings of non-native plant diseases in Europe (including bacteria, fungi
and nematodes; data derived from Smith 1997). Fitted line represents the best
fit linear regression model (y=-52.7 + 0.028x, P = 0.006, r2 =0.54).
2.2.2.2 Patterns of spread
Data on the spread rates and patterns of spread for plant diseases were less
readily available than some of the other taxa. The data available show that
these organisms can spread very rapidly, sometimes across continental
scales in individual years, though it is often difficult to determine whether the
organisms are dispersing, or whether the apparent population spread actually
represents wave of local infection through space caused by the emergence
from local resting stages in response to climatic or other variables.
Experimental tests of disease propagation at small temporal and spatial
scales do, however, highlight fast spatial dynamics (Table 2.3). There
appears to be a distinction in these data between pathogens of woody species
and those of herbaceous species, the latter being orders of magnitude faster.
25
A New Agenda for Biosecurity, August 2004
Table 2.3: Parameter values found in the literature for the spread of plant diseases. Here r
refers to the intrinsic rate of increase of the infected population of the host
species. Since these are agricultural host species, this rate of increase of
infection relates readily to spatial extent of infection.
Common
name
Name
r
(year)

D
2
(km
yr-1)
(km2 -1
yr )
Predicted
Radial
expansion
(km yr-1)
Observed
Radial
expansion
(km yr-1)
A
(km2
yr-1)
Source
herbaceous
hosts
potato blight*
Phytophthora
infestans
58-153
(Zadocks
and
Shein
1979)
yellow rust*
Puccinia
striiformis
36-99
(Zadocks
and
Shein
1979)
36
(Zadocks
and
Shein
1979)
tomato
mosaic virus*
woody hosts
oak wilt
disease*
Ceratocystis
fagacearum
0.77
(Zadocks
and
Shein
1979)
leaf rust*
Cronartium
fusiforme
0.4
(Zadocks
and
Shein
1979)
wilt fungus*
Fusarium
oxsporum
0.5
(Zadocks
and
Shein
1979)
root rot
fungus*
Phytophthora
cinnamomi
1.54
(Zadocks
and
Shein
1979)
sudden
death*
Valsa
eugeniae
0.34
(Zadocks
and
Shein
1979)
tobacco blue
mold*
13.9 km/d
(Aylor 2003)
wheat stem
rust*
35 km/d
(Aylor 2003)
Dispersal mechanisms are highly variable, and this may account for some of
the variability observed, e.g. between fungal spores transmitted by water and
wind. Aerial dispersal of spores and movement by insect vectors must
certainly contribute to the rapid spread of relevant species, but satellite
creation by movement of infected plant material is also important, as has been
postulated recently in US and Europe for the movement of Phytophthora
ramorum via ericaceous shrubs distributed widely to garden centres.
26
A New Agenda for Biosecurity, August 2004
2.2.3 Vertebrates
Terrestrial vertebrates, including mammals, birds, reptiles and amphibians,
have a long history of introduction into UK and integration into natural and
agricultural ecosystems. The great majority of vertebrate introductions have
resulted from the intentional importation of non-native species, though not
necessarily their intentional release. Historically, introductions for food,
hunting and fur production have predominated. More recently, introductions
have been associated with the growth of trade in exotic pets and of zoological
parks. Not surprisingly, many non-native vertebrates are held to be desirable,
even if they may cause harm locally when in dense numbers. Non-native
deer, rabbits, pheasants and other species are valued by many as an addition
to our biodiversity and considered part of our countryside culture. Reintroduction programmes of birds and mammals, and range extensions of
species from Europe, like collared dove and little egret, probably serve to
increase public acceptance of changes in our vertebrate fauna, and blur the
native/non-native divide (see Chapter 6.3.1).
High local populations of non-native mammals and birds can cause
substantial agricultural impact (e.g. Mountford 1997; White and Harris 2002
and see Chapter 4.2). However, most attention today is directed towards the
environmental impact of introduced vertebrates. Predatory species, such as
mink, can reduce populations of native species, while both predators and
herbivores can displace native species, through competition for food, breeding
sites or other resources. The potential for indirect competition is increasingly
apparent – e.g. where an aliens species brings with it a disease to which it is
resistance but which affects a native species, as has been suggested for grey
and red squirrel.
Predation and competition by non-natives have demonstrable potential to
drive native vertebrate species to extinction, but this is much more likely on
isolated or small island habitats than at a national level. The genetic erosion of
native species through interbreeding with introduced species is a quite
different biodiversity effect of considerable importance in UK. Examples
include interbreeding of red and sika deer, of domestic and wild cat, and of
ruddy and white-headed duck.
Non-native vertebrates will also have effects on ecosystem services,
particularly grazing species like deer, wild boar and rabbits, which modify flora
and affect habitat succession. Coypu, an aquatic rodent introduced from
South America for its fur, had a serious effect on river bank structure and
hence waterway services, and was eradicated through an substantial
campaign in the 1980s (White and Harris 2002).
2.2.3.1 Patterns of introduction
A review of the introduction of terrestrial non-native vertebrates into the UK
reveals that current introduction rates are near zero (White and Harris 2002).
Although 23 mammal, 21 bird, 8 amphibian and 3 reptile species are listed,
only 1 mammal, 2 bird, 2 reptile and 3 amphibian species entered the UK
between 1950 and 2000, no species from these groups were listed as
entering between 1990 and 2000. That introductions are declining overall may
reflect widespread awareness of the risks posed by non-native mammals and
27
A New Agenda for Biosecurity, August 2004
birds to native biodiversity and ecosystem. Terrestrial vertebrate introductions
in future are likely, therefore, to be fewer.
2.2.3.2 Patterns of spread
Good data are available on the spread rates and population parameters for
several invading vertebrate species (Table 2.3). There is an apparent
distinction between the spread rates of mammal species and bird species, the
latter usually being more rapid. Predicted spread rates based on a simple
diffusion model often a close to observed spread rates, particularly in mammal
species.
Table 2.4: Parameter values found in the literature for the spread of non-native terrestrial
vertebrate species
Common
name
r
Name
(year)

D
2
(km
yr-1)
(km2 -1
yr )
Predicted
Radial
expansion
(km yr-1)
Observed
Radial
expansion
(km yr-1)
6-32
1- 25
(Andow et al.
1990)
A
(km2
yr-1)
Source
mammals
muskrat
Ondatra
zibethica
0.21.1
51230
red deer
Cervus elaphus
1-1.6
(Clarke 1971)
Himalayan
thar
Hemitragus
jemlaticus
0.68
(Caughley
1970)
Californian
sea otter
Enhydra lutris
0.056
13.5,
54.7
1.74, 3.5
1.4, 3.1
(Lubina and
Levin 1988)
Wild boar
Sus scrofa
0.020.3
0.63.0
0.5-1.1
(MAFF 1998)
Birds
European
starling
Sturnus vulgaris
11.2
(Shigesada
and Kawasaki
1997)
“
“
91.6
(Caswell
al. 2003)
house finch
Passer
domesticus
35
(Mundinger
and
Hope
1982)
collard dove
Streptopelia
decaocto
56.3
43.7
(Hengeweld
1989)
Pied
flycatcher
Ficedula
hypoleuca
0.33
4.74
1.1
1.5
(Caswell
al. 2003)
et
sparrowhawk
Accipter nisus
0.038,
0.145
317,
373
2, 4.9
2.4, 3.1
(Caswell
al. 2003)
et
Grey squirrel
Sciurus
carolinensis
0.82
17.9
7.7
7.7
(Okubo et al.
1989)
0.01
770.3
61.2
et
2.2.4 Animal diseases
Animal disease threats to UK are created by a range of viruses, bacteria,
protozoa and fungi, as well as by parasitic arthropods and nematodes. Insect
28
A New Agenda for Biosecurity, August 2004
vectors of disease must also be considered, as their introduction or spread in
UK may increase transmission of new and endemic agents. Animal diseases,
are often highly contagious, particularly in the high density environments
typical of modern livestock and poultry production, and can be lethal or
chronic, causing reductions in fitness and productivity. They are usually
introduced through movement of infected animals (pets and wild animals may
be carriers, as well as livestock) or animal products (e.g. contaminated meat).
The prevention and control of livestock diseases has a long history, and is
managed today by substantial national veterinary programmes, backed up by
international agreements on movement and control, including for UK, EU
regulation and agreements under the Office International des Epizootiques
(OIE). Under the OIE, a list of priority diseases is recognised which must be
monitored by all governments “stamped out” when it appears. Presence of
disease has an immediate and dramatic effect on trading in relevant animal
products, hence stamping out usually involves rapid culling campaign,
followed by an assessment period and restoration of disease free status and
trade. Recent examples in Europe of foot and mouth disease (FDM), swine
fever and bird flu affecting poultry illustrate these features of animal disease
invasions. Disease “scares” may close markets even without any official
procedure.
So distinct has been the management of animal diseases that it is often set
apart from efforts to integrate non-native species activity (e.g. animal diseases
were specifically excluded from the Defra Non-Native Species Review (2003).
We conclude that any differences lie more with the distinctive history of
veterinary practices and institutions than with distinctive biological features of
animal diseases. However, as with plant diseases, concepts of establishment
and spread may have a different interpretation than they have for animals and
plants.
There have been a number of detailed reviews of animal disease threats to
UK following recent outbreaks, notably those produced by the Royal Society
(2002), Veterinary Laboratories Agency (2003) and the Institute of Animal
Health (draft report for Defra, 2003). We summarise here key points relevant
to our study.
Local disease outbreaks are usually identified and controlled so quickly that
the concept of establishment in a population ecological sense is difficult to
apply – the entire FMD outbreak of 2001 would not be considered by some
animal health experts as “establishment” of that disease, as it did not become
endemic. A typical historical pattern for a new, non-native animal disease is a
series of outbreaks, followed perhaps by an epidemic, which are successively
stamped out. Some outbreaks may lead to the disease becoming endemic,
but then it still may die out. The critical factor here is the number of new
infections caused by one infective host (Ro in epidemiological modelling
terms) – if this falls below one, due to a failure of survival or transmission in
the host population, the disease will decline to extinction. Further outbreaks
continue this process and opportunity for endemism.
In this context, the concern today that animal disease risks will continue and
increase relates to three factors:
29
A New Agenda for Biosecurity, August 2004

global growth in animal production systems (the “livestock revolution”
FAO 2002) and increased international movement of disease by
different pathways may increase frequency of outbreaks,

an increase in those conditions which improve disease transmission
and survival (e.g. vectors, alternate hosts, intensive production
systems) may worsen outbreaks and increase the risk of endemism,

in addition to outbreaks of known diseases, there is a greater possibility
for new, emerging diseases resulting from contact with new host
sources or regions, as well as from disease evolution, where mutation,;
selection and hybridisation may generate new virulent strains.
Opportunities for new disease emergence and evolution are also
improved by growing international movement of animals and animal
products.
Recent studies suggest that major animal disease threats to UK include new
outbreaks of mostly viral diseases, including as rabies, FMD, classical swine
fever, Newcastle Disease, avian influenza, enzootic bovine leucosis, and
equine viral arteritis, all of which have appeared before. Amongst diseases
which would be entirely new to UK, the viruses responsible for blue tongue,
West Nile virus, African swine fever, swine vesicular disease, and yetunknown emergent diseases are among the most likely threats in the near
future.
Biological features shared by many of these new disease threats include the
existence of wild animal reservoirs (and usually a degree of wild animal
impact, see below) and their zoonotic potential. From an economic
perspective, we will see in Chapter 5.1 that these features could make future
impact of such threats enormous: export income, conservation and human
health effects are all associated with high levels of cost to the UK economy.
Beyond threats to animal production, there is considerable current concern
about the spread of wildlife diseases. The introduction of new, non-native
species/strains is one of several factors causing this, others include
anthropogenic changes in production, land use and climate which affect
disease spread and survival, the evolution of new diseases in new conditions
and hosts (Williams et al. 2002). The introduction of new wildlife diseases to
new countries is closely linked to livestock production. Diseases of wild
animals are not easily transmitted across continental barriers, and movement
domestic animals provide one of the few pathways for this. Further, domestic
animals constitute, in a new area, a substantial host reservoir for a new
pathogen to build up and infect, or even evolve onto, local native species. The
growing proximity of human habitation and activity and natural habitats and
wild species (e.g. through patterns of housing development and
“suburbanisation” of wildlife like deer, foxes, etc.) will increase this risk.
Two recent trends in introduced wildlife diseases deserve particular mention.
Firstly, the growing international trade in pets has encouraged the introduction
and mixing of a range of pathogens. A recent example is the appearance of
an African monkeypox virus in prairie dogs, a wild rodent now traded as a pet
in the USA (The Independent, 9 June 2003). Secondly, while many
domesticated animals pose disease risks for related native species, the
30
A New Agenda for Biosecurity, August 2004
greatest threat appears to come from newly domesticated or commercialised
species and their international production. Relatively recent initiatives in game
farming, fish farming, shrimp and crayfish farming and frog farming which
address public demand for novelty foods or international demand for new
protein sources, are also sources of recent new wildlife diseases. This
problem must reflect a poor understanding of disease risks in newly
domesticated species, but it points clearly to how entire taxa of native species,
such as amphibians, now face hitherto unimaginable levels of risk from new
diseases.
2.2.4.1 Patterns of introduction
Because diseases of domestic animals are so often “stamped out” as local
outbreaks, the concept of “first establishment” is not so useful to determine
current baseline rates of introduction of new diseases. Even data on the rate
of outbreaks is poor – while all governments report to OIE on these, we are
advised that these records are not dependable, given the strong trade
implications of reporting a new outbreak. Hence, we have not identified or
analysed databases of introduction to extract changes over time. IAH (draft
report for Defra 2003) predict that, over the near future, the number of new
disease outbreaks (including re-appearance of eradicated diseases and
entirely new diseases) will rise, largely as a result of trade and travel. Human
transport of contaminated produce and movement of animals will continue to
be the key pathways for introduction. Over a longer period, climate change will
increase the frequency of disease outbreaks by creating more favourable
conditions for survival of pathogens and vectors. In the next 10-30 years, IAH
predicts that about 30 entirely new animal diseases will appear, some of
which will reach UK.
2.2.4.2 Patterns of spread
There is only limited, historical information available on the unconstrained rate
of spread of animal diseases, because of the widespread policy of stamping
out. The spread of disease in human populations may provide a useful
comparison. We present just two examples in Table 2.5.
31
A New Agenda for Biosecurity, August 2004
Table 2.5: Parameter values found in the literature for the spread of animal diseases
Common
name
r
(year)

D
2
(km
yr-1)
(km2 -1
yr )
Predicted
Radial
expansion
(km yr-1)
Observed
Radial
expansion
(km yr-1)
A
(km2
yr-1)
Source
Black death
19
25000
720
320-650
(Nobel 1974)
Rabies
66
40-50
70
30-60
(Yachi et al.
1989)
The recent outbreaks of FMD have generated considerable information and
modelling on rapid disease spread which highlight the role of aerial spread
and satellite creation through long distance movement of animals. But the
extremely rapid spread of this disease is not necessarily typical for new
introductions into UK. Other existing or potential UK animal diseases, such as
bovine TB, TSE and rabies, are known to move more slowly.
2.2.5 Terrestrial plants
Plants are perhaps the best understood taxon of non-native species, largely
because of their ease with which they can be identified, monitored and
studied. In UK, the record of plant introductions is long: specialists use the
term “neophytes” to identify species introduced, usually by human activity
after 1500, and archaeophytes as non-native species introduced prior to this
period (Preston et al. 2002).
The great majority of the thousands of anthropogenic plant introductions into
Britain have been intentional, associated with agriculture, forestry and
gardening. Few have become invasive: Williamson’s (1996) tens rule applies
here: as a rule of thumb, one in ten introduced plants escape cultivation, one
in ten of them establishes and one in ten of them spreads in an invasive
manner. Nonetheless, as we shall see, even this small proportion is
potentially a substantial number of invasive plant species, given the high
historical levels of introduction.
Non-native plant species adapt to many habitats but are particularly common
in disturbed and successional situations. Amongst non-native plant species,
agricultural crops are usually not invasive in UK. In other parts of the world,
non-native pasture grasses and trees species selected for rapid growth have
proven invasive, hence new crops, particularly species selected for biomass
production, deserve future attention. Most invasive non-native plants in UK
were introduced as garden ornamentals, like Rhododendron ponticum,
Japanese knotweed (Fallopia japonica), Himalayan Balsam (Heracleum
mantegazzianum) and New Zealand Pigmyweed (Crassula helmsii), one of
several garden ornamental pond plants. These invasive “garden escapes”
invade particularly aquatic or riparian habitats and some invade woodlands.
Successional habitats, such as those in conversion from agricultural to
“natural” land are particularly invasible by non-native species.
The impact of invasive, non-native plants is varied. There are few species in
UK that cause major losses to agricultural production, mostly pasture: the
case study on Cirsium arvense in Chapter 4.6, gives a “hypothetical” example
32
A New Agenda for Biosecurity, August 2004
of such a situation. Most impact appears to be environmental, involving
displacement of native flora, which affects appearance of, and access to,
natural habitats, and reduced abundance of local species. Extinction of native
species is unlikely, but “species loss” does arise through hybridisation of
native and non-native species. This “genetic erosion” of native species is
occurring, for instance, with Spartina anglica, a hybrid between native and
alien cordgrass, and the introgression of Spanish bluebell, Hyacinthoides
hispanica, into our native species.
Non-native plants can also have severe impacts on ecosystem processes and
services. In some parts of the world, for instance, invasive grasses change fire
regimes, encouraging fires which eliminate diverse grasslands and forests,
and facilitate the further spread of the non-native grass species (Mack et al.
2000). In other regions, non-native tree introductions have the opposite effect
of creating woodlands from native grassland. These profound impacts on
ecosystem function are not typical of UK plant invasions, except perhaps in
aquatic systems, where extensive infestations of non-native water weeds can
change levels of light and water chemistry, with impacts on the entire aquatic
food chain.
2.2.5.1 Patterns of introduction
There are relatively good records on the patterns of plant species introduction
into the UK. We will concern ourselves with recent introductions, i.e. in the
past few centuries. The New Atlas of the British and Irish Flora (Preston et al.
2002) provides interesting data on the pattern of plant introduction. Dates of
introduction of plant species, usually as ornamentals, and first record in the
wild (naturalisation, or what we will call establishment) were extracted for all
neophyte3 species, sub-species, hybrids and aggregates present in the
database (n=1304). This data is presented in Figure 2.5.
Consider first Fig. 2.5b, which shows that establishment of non-native species
has risen steadily over the past century, no doubt drawing from a large
accumulated pool of introduced non-natives in gardens. There is no sign of
establishment rates declining and about 70-80 establishments are predicted
for the current decade. Applying the “tens rule” to these establishments, less
that ten invasive species are likely to establish in this period.
Fig. 2.5a shows the pattern of introduction over the past few centuries – the
fact that it declines after the 1800s is an artefact – the Atlas contains only
species established in natural habitats, hence it will not contain species which
are introduced in the 1900s but not yet established in the wild. According to
experts at the Royal Horticultural Society, the rate of introduction of new
garden plant species has probably increased over recent decades due to
growing interest in gardening and the demand for novelty (S. Thornton Wood
and C. Prior, pers. comm.).
The discrepancy between Figs. 2.5a and b is revealing: a plant may take
many years to move from introduction in a garden to establishment in natural
3
Neophytes are non-native taxa in which the introduction resulting in naturalisation occurred
after 1500. To be designated as naturalised, a population must be self-sustaining (not reliant
on further introduction) for more than five years.
33
A New Agenda for Biosecurity, August 2004
b)
0
20
20
40
40
60
80
60
100
80
a)
0
Number of taxa per decade
120
habitats. In Fig. 2.6 we show the distribution of this time lag, whose modal
period is 100 years. Consider that an established plant which is invasive will
then take some decades to spread (Table suggests a maximum linear/radial
spread of at most, tens of km per year and usually much less), and it will be
clear to see that a known invasive plant introduced to UK today may not have
a substantial impact for more than a century.
1500
1600
1700
1800
1900
1500
2000
1600
1700
1800
1900
2000
Figure 2.5: Decade of first introduction (a) and first record in the wild (b), of neophyte taxa in
the UK and Ireland (compiled from Preston et al., 2002). Lines are fit by local
non-parametric regression.
160
Number of taxa
140
120
100
80
60
40
20
0
0 - 25 - 50
- 75 - 100 - 125 - 150 - 175 - 200 - 225 - 250 - 275 - 300 - 325 - 350 - 375 - 400 4001000
Lag period (years)
Figure 2.6: Frequency histogram of the lag between year of first import and year of first
record in the wild for established (naturalised) plant taxa in the UK and Ireland
(compiled from Preston et al., 2002).
2.2.5.2 Patterns of spread
Not surprisingly, plants are usually slow to spread, and this is clear from Table
2.6. Data on woody plants (Hengeveld 1989), when compared to herbaceous
species at the top the table shows that trees spread even more slowly.
Table 2.6: Parameter values obtained form the literature relating to the spread non-native
plant species. † - data concern post-glacial spread of species as estimated from
pollen analyses
Common
name
Name
r
(year)
D

(km2 y
r1)
(km2 -1
yr )
Predicted
Radial
expansion
Observed
Radial
expansio
A
(km2
yr-1)
Source
34
A New Agenda for Biosecurity, August 2004
(km yr-1)
n (km yr-1)
Solidago
altissima
741
(Weber 1998)
Solidago
gigantea
910
(Weber 1998)
Solidago
graminifolia
128
(Weber 1998)
Alliaria petiolata
366528
0
(Nuzzo 1993)
645
(Perrins
1993)
100
(Thompson 1987)
Impatiens
gladulifera
0.06
2.6-3.8
44541
9.4-32.9
Lathyrum
salcicaria
Cheatgrass
Bromus
tectorum
Japanese
sedge
Carex
kobomungi
Rhamnus
frangula
3.5 x
10-5
5
(initial
phase)
et
al.
(Shigesada et al.
1995)
0.01
6
0.197
0.0063
0.0067
(Frappier
2003)
et
al.
Cirsium arvense
0.006
†
Abies
0.04-0.3
(Hengeweld 1989)
†
Acer
0.5-1
(Hengeweld 1989)
†
Alnus
0.5-2
(Hengeweld 1989)
†
Carpinus betulus
0.05-1
(Hengeweld 1989)
†
Carpinus sativa
0.2-0.3
(Hengeweld 1989)
†
Corylus-type
1.5
(Hengeweld 1989)
†
Fagus
0.2-0.3
(Hengeweld 1989)
†
Fraxinus
excelsior - type
0.2-0.5
(Hengeweld 1989)
†
Fraxinus ornus
0.03-0.2
(Hengeweld 1989)
†
Juglans
0.4
(Hengeweld 1989)
†
Picea
0.08-0.5
(Hengeweld 1989)
†
Pinus
1.5
(Hengeweld 1989)
†
Pistacia
0.2-0.3
(Hengeweld 1989)
†
Quercus(decidu
ous)
0.08-0.5
(Hengeweld 1989)
†
Tilia
0.05-0.5
(Hengeweld 1989)
†
Ulmus
0.1-1
(Hengeweld 1989)
Median (5,
95 %iles)
0.29 (0.01,
1.85)
Under such circumstances, the creation of satellite populations becomes very
relevant. There is only one study where satellite formation is considered, on
the British grass, Bromus tectorum in USA, and here it is clear that  is very
large and significant relative to normal, vegetative spread. Highly invasive
non-native weeds often show strong satellite formation: the spread of Bromus
is thought to have been by road as seeds; new water weeds spread rapidly as
widely sold commercial ornamentals, subsequently discarded or washed out
into local water bodies; weeds of disturbed areas travel as seeds or plant
fragments in transported soil. For this and purely demographic reasons, the
spread of an invasive weed when it finally begins to “take off” can be dramatic
and newsworthy.
35
A New Agenda for Biosecurity, August 2004
2.2.6 Aquatic species
To include all aquatic non-native species for consideration in one section is
problematic. This grouping cuts across taxa described separately above with
their clear differences in properties relevant to introduction, spread and
impact. Further, aquatic systems comprise two very different ecosystems:
freshwater and marine. However, we have consolidated this area for two
reasons:
Firstly, many aquatic invasions are poorly understood, relative to terrestrial
systems, and it is therefore useful to pool limited information across this
ecosystem. This lack of understanding reflects in part a historical lack of
national and international oversight of non-native species introductions,
particularly in marine environments. While IPPC and OIE have ensured
international “coverage” of non-native terrestrial animal and plant problems,
aquatic systems have fallen “between regulatory stools” and with the
exception of fisheries management, it was not until the 1990s, through interest
generated by the Convention on Biological Diversity and, particularly, the
International Maritime Organization, that the full extent of damaging aquatic
invasion has been addressed.
Secondly aquatic systems have distinctive properties relevant to biological
invasion. New species can disperse freely and rapidly, often through
planktonic larval stages, mixing with local fauna and flora. Aquatic systems
also constitute relatively self contained food chains which can be profoundly
disrupted by single invasive species, with resultant impacts on ecosystem
function and services. UK scientists working on freshwater invasions told our
study that “things happen more quickly and dramatically in aquatic systems”,
in contrast to terrestrial ones, perhaps for this reason.
These easily disrupted ecosystems can be as small as a local lake choked by
a non-native water plant, affecting light intensity, oxygen, shelter and other
factors which then induce changes in local plant and animal populations. Or
the ecosystem can be as large as the Black Sea. This water body was
invaded in the 1980s by a non-native predatory comb jelly which effectively
shut down the food chain between zooplankton and fish, resulting in a
collapse of fish populations, algal blooms and its own fluctuation population
dynamics. At peak populations, this one species comprised most of the Sea’s
biomass (Williamson 1996).
The following comments are focused particularly at freshwater ecosystems,
but many apply to marine systems. Marine systems will be considered at the
end.
Most information on non-native species problems exists for aquatic production
systems, particularly fish production. Historically, introduction of new fish
species, for game fishing and farming, has generated a range of problems, in
particularly predation or displacement of native fish and aquatic invertebrate
species. However, new fish introductions are declining, and most of these are
into fish farms and angling ponds where further spread is unlikely, except in
the event of flooding. Introduction of new species through the aquarium trade,
as discarded “pets” is now, perhaps the greater non-native species risk,
36
A New Agenda for Biosecurity, August 2004
particularly as water temperatures rise, but substantial negative effects have
not yet emerged.
By far the major risk today to both commercial fishing and conservation of
game and other fish species is posed by non-native fish diseases and
parasites (CEFAS, pers. comm.). Outbreaks of introduced viral diseases of
native and farmed trout have been frequent over recent decades. Parasitic
worms, such as Gyrodactylus salaris, have caused fatal infections of Atlantic
salmon in Norway and now threatens UK systems (See Chapter 4.5).. New
fish diseases and parasites may enter regions through aquaculture,
subsequently spreading to native species. The Diseases of Fish Act 1937,
and subsequent UK and EU regulation provides good protection, although an
expanding, open Europe places strains on its application. Known disease
threats can be anticipated, but it is the new, unknown diseases, as well as
known diseases which behave differently in new areas which are of greatest
concern (CEFAS, pers. comm.).
A parallel situation can be seen in other aquatic production systems. The
introduction of the North American signal crayfish, Pacifistachus leniusculus,
into UK has led to the decline of native crayfish, largely through the spread of
a crayfish disease to which the native species was more susceptible. In
tropical regions, the spread of shrimp farming has brought with it disastrous
outbreaks of introduced viruses and other diseases.
As with animal diseases and wildlife, it appears that “domesticated” aquatic
species in aquaculture systems create direct threats to related, native species
by creating channels for the movement of disease and reservoirs for its
maintenance and continuous re-introduction.
Other environmental impacts of non-native species in aquatic systems include
loss of biodiversity (e.g. through predation by introduced species) and loss of
ecosystem function and services. In this latter category, we may include:

changes in water quality – introduced carp and translocated barbel
increases turbidity, with impacts on plants and animals, creating
entirely different water systems with different sediment and chemical
properties.

changes in structure of waterways – large infestations of non-native
water weeds can contribute to flooding and changes in water flow,
while bank-burrowing species like Chinese mitten crab (and the
eradicated coypu) can weaken the sides of water channels.

changes in food chains – planktonic filterers like zebra mussel, or
scavengers like Chinese mitten crab may, by virtue of their sheer
abundance, starve food chains of resources, while top predators may
eliminate keystone species important to water quality.
In marine environments, there is a similar pattern of introduced species
affecting both production systems (e.g. as parasites and pathogens carried
with new species for mariculture), local biodiversity and ecosystem processes.
(GISP 2003).
37
A New Agenda for Biosecurity, August 2004
2.2.6.1 Patterns of introduction
As indicated earlier, many freshwater introductions have been both deliberate,
in the form of game species, and accidental, in the form of escaped farm
species and water plants and the diseases and parasites inadvertently
introduced with various non-natives. In marine systems, it is clear that that
movement of pelagic larval stages in ballast water has been a major pathway
for the introduction of invertebrates to near-shore environments in distant
parts of the world. International action has now been initiative to stop this
movement (National Research Council 1996). Hull fowling by sessile species
is also thought to have been a major pathway of introduction. Rates of
introduction in particular countries have been high: in the San Francisco Bay
in California, it has been estimated that a new non-native species establishes
every 12 weeks (Cohen and Carlton 1995; Bright 1998).
There are now a number of extensive data sources on the introduction of nonnative aquatic species, particularly in Australia and USA. We have analysed
some UK-relevant data from two sources, and cite a third below:

FAO Database on Introductions of Aquatic Species (DIAS:
http://www.fao.org/waicent/faoinfo/fishery/statist/fisoft/dias/)
which
contains entries for the UK and many other countries worldwide. The
database was initially developed in the 1980s for freshwater fish
species, but has since expanded to include molluscs and crustaceans
and marine species. It was compiled largely by questionnaire
responses by experts, and contains deliberate introductions and those
accidental introductions which have been identified as having severe
impact on native species. Thus the database contains a biased subset
of the non-native species in UK waters. Figure 2.7 shows the number
of species per decade arriving to the UK over the last 120 yrs. There
is little that can be concluded from these data on temporal trends in
the rate of species entry, but it is clear that until relatively recently
species were being introduced into the UK.
Species per decade
6
5
4
3
2
1
0
1880s 1890s 1900s 1910s 1920s 1930s 1940s 1950s 1960s 1970s 1980s 1990s
Figure 2.7: Arrival of non-native aquatic species to the UK as collated in the FAO DIAS
database.
 Baltic Sea Alien Species Database. (http://www.ku.lt/nemo/alien_
species_search.htm). This database covers marine and estuarine species
38
A New Agenda for Biosecurity, August 2004
in the Baltic Sea and was compiled from the literature and expert opinion.
Here there is evidence of an increasing rate of entry of non-native species
through the last century and no evidence to date of a slow-down in
introduction (Figure 2.8). Whilst there is evidence that the introduction of
species for deliberate stocking is starting to be controlled, and with it the
introduction of parasites/symbionts associated with these species, the rate
of species inadvertently introduced through shipping continues to increase.
When we break these species down to individual taxa, markedly different
patterns are observed, but high variability over the last several decades
makes prediction of likely future rates extremely difficult (Figure 2.9).
16
Associated
Species per decade
14
Shipping
12
Stocking
10
8
6
4
2
Figure 2.8: Rate of arrival of non-native species to the Baltic Sea due to deliberate stocking,
species associated with deliberate stocking, and accidental introductions due to
0
shipping.
1880s 1890s 1900s 1910s 1920s 1930s 1940s 1950s 1960s 1970s 1980s 1990s
Species per decade
16
14
Crustacea
12
Mullusca
Pisces
10
total
8
6
4
2
Figure 2.9: Rate of arrival of species to the Baltic Sea according to taxon.
0

1900s 1910s
1930s 1940s a1950s
1960sof1970s
1980s 1990s
Eno1880s
et 1890s
al. (1997)
have1920s
undertaken
study
non-native
marine
species (plant and animal) in British waters. Their analysis of new
established species recordings by decade from the late 1800s show
no distinctive pattern, except for a boom in establishments recorded in
the 1970s. No non-native marine fish were identified, and most
introduced species were red algae, polychaete worms, crustacea and
mollusks. In more recent decades, ballast water, ship fouling and
39
A New Agenda for Biosecurity, August 2004
mariculture appear to be the more important and emerging pathways
of introduction.
2.2.6.2 Patterns of spread
Spread of non-native species through catchments can be very rapid. Spread
between isolated water bodies and catchments is constrained by isolation.
Nonetheless, where there is a mechanism facilitating movement between
water bodies, spread can be rapid, as is the case in US waterways with the
movement of waterweeds by recreational boating. There have been a number
of studies of dispersal in marine systems, particularly for sessile or bottomdwelling species with pelagic larvae. These are illustrated in Table 2.7.
40
A New Agenda for Biosecurity, August 2004
Table 2.7: Data from Grosholz (1996) on the rate of spread of ten marine species.
Common name
or class
Name
r
D
2
(year)
(km y
r1)
Predicted
Radial
expansion
(km yr-1)
Observed
Radial
expansion
(km yr-1)
Tunicate
Botrylloides
leachi
1.8
1800
114
16
Crab
Carcinus
maenas
3.3
925
111
55
Barnacle
Eliminius
modestus
3.5
265
61
30
Crab
Hemigrapsus
sanguineus
2.5
761
87
12
Snail
Littorina
littorea
3.9
242
62
34
Bryozoan
Membranipora
membranacea
9
50
42
20
Mussel
Mytilus
galloprovinciali
s
3.2
1352
131
115
Mussel
Perna perna
8.4
338
107
95
Opisthobranch
mollusc
Philine
auriformis
8.9
365
114
80
Opisthobranch
mollusc
Tritonia
plebeian
7.5
288
93
50
Median
%iles)
(5,
95
100 (50, 123)
Grosholz (1996) compared spread rates, based on diffusion models, of these
ten marine species with ten terrestrial species, including insects, birds, plants,
mammals and animal diseases. Contrary to the prediction that marine
dispersal would be more rapid, due to pelagic larval stages, it was in fact
significantly slower on average. High variance suggested that long-range
dispersal of larvae of marine species occurred, but principally as a rare event.
Spread of marine species was not as well described by diffusion equations
than spread of terrestrial species.
2.3 Conclusions from cross-taxon review
This cross-taxon review has been necessarily superficial. Each taxon would
benefit from more detailed research, particularly as the evidence base for
introduction, spread and impact is so poor. However, it is clear from this initial
analysis that, at least on a European level, non-native species introductions
have been increasing across a number of taxa in recent decades, and are
likely to continue doing so. Data available are subject to two main errors which
would act in opposite ways – growing awareness will lead to growing reporting
of establishments, while improving prevention will reduced reporting of
establishment.
41
A New Agenda for Biosecurity, August 2004
Very rough estimates from the survey of taxa above suggest that rates of
establishment of harmful non-native, terrestrial invertebrates, plant diseases,
animal diseases and weeds may be on the order of 5-10 species per taxon
per decade. As a taxon, only vertebrates appear to be declining in rates of
establishment in recent decades, although information on aquatic
introductions is still too poor to identify a trend. Overall, we conclude that there
is a growing risk of non-native species introduction.
What can we say about the origin, nature and potential impact of this risk?
2.3.1 Origin of non-native species risks
Whereas it is traditional to think of biosecurity risks being associated with
accidental contamination of agricultural (animal and plant) imports, our
analysis suggests that these imports account for only a fraction of the
problem, and possibly a declining one. A large source of risk is deliberate
importation of plant and animal species, for the garden trade, the pet trade,
new food production systems (e.g. fish, crayfish production). This risk is
twofold – species intentionally introduced into controlled habitats may escape
into natural ecosystems, and those escaped species may be contaminated
with diseases that will affect native or agriculturally important species. When
we come to consider prevention and control, it will be clear that intentional
introductions offer substantial opportunities for risk reduction, relative to purely
accidental introductions.
Another distinctive aspect of the origin of risk is that introduced by evolution
and adaptation of introduced species. Across the taxa surveyed are examples
of non-native species which hybridise with native species or evolve to adapt to
local conditions, thereby creating new problems for both agriculture and the
environment. These evolutionary effects are even less predictable than those
caused by introduction alone. Richardson et al. (2000) and Simberloff and
Von Holle (cited in Baskin, 2002, p 144) illustrate a number of other ways in
which non-native species interact with each other or native species to
accelerate or exacerbate invasions. These include new associations of
species which facilitate reproduction, spread or survival of a non-native, such
as dispersal of non-native plant seeds by native or introduced birds. Note that
these “second order effects” are not incorporated into the ecological model,
and hence not in the economic model to follow. This means that impact of
non-natives may be underestimated for this reason.
Many authors have associated a growing risk from non-native species with the
rapid growth of international trade. Trade statistics cited include the value of
traded goods, which have risen from $192b in 1965 to $6.2 trillion in 2000.
Upward trend statistics are also seen in commodity imports and container
shipments. Often, however, it is hard to link the introduction of particular taxa
with such broad pathways, and we must ask whether positive correlations of
general trade and increase in introductions are really evidence of causation?
Many of our new problems appear associated with pathways other than the
bulk import of traded or even agricultural commodities. Speciality importers, of
pets, plants, certain foodstuffs (e.g. bushmeat), game fish as well as individual
travellers bringing such species into the country, may pose a greater risk from
many taxa than large-scale commodity movements. For some taxa, like
42
A New Agenda for Biosecurity, August 2004
plants, tomorrow’s non-native species problems are already here, and the rate
limiting process is not introduction but establishment and spread in nature.
A much more refined analysis of changes in trade, travel, transport and
tourism will be needed to firmly demonstrate the link between non-native
species risks and recent global trade liberalisation. However, all institutions
consulted for this study have observed that the complete removal of trade
restrictions within Europe will have major implications for risk. Recent UK
biosecurity crises have shown the ease with which animal (FMD, BSE) and
plant (potato ring rot, sudden oak death) diseases may move between UK and
continental Europe via trade, and pan-European movement of new
environmental problems (e.g. oak knopper gall, horse chestnut leaf miner),
probably by other routes, is now a regular phenomenon. With Europe’s new
members, and even more distance and permeable borders, non-native
species risk may increase substantially. On the plus side, such risk may now
be detectable earlier and at some distance from the UK, and EU regulation
may provide new muscle to biosecurity measures.
2.3.2 Nature and impact of non-native species
This study suggests that there are three important ways in which non-native
species can affect the UK economy and society. Firstly, non-native species
may directly affect production systems, like agriculture, as pests, diseases
and weeds. Secondly, due to international biosecurity and trade structures,
the appearance of a new non-native species may affect trade and the
economy, even without having any direct harmful effect. This phenomenon is
seen with animal, fish and plant diseases, but may grow in future.
Thirdly, all taxa of non-native species pose some kind of risk to the
environment. Indeed, this is the most rapidly emerging kind of impact amongst
all these taxa. Environmental impact of non-native species can be of two
kinds.
1.
Non-native species can directly affect biodiversity, reducing the
abundance of native species through predation, competition
(including the introduction of diseases which have more impact on
natives) or parasitism (in the case of introduced parasites or
pathogens). Extinction has rarely been a consequence of nonnative species impact in UK, but the loss of native plants and
animals through hybridisation with non-native species is likely.
2.
Non-native species can affect ecosystem processes and services,
such as the supply of clean water and air, or the functioning of
ecosystems to provide resources which support animal and plant
communities in food chains, or ecological succession. Obvious
examples include non-native species which affect the physical
environment, such as the structure of waterways and channels, or
the turbidity and quality of water. Invasive plants may affect ground
cover and soil structure, while non-native herbivores may do the
same by removing vegetation.
43
A New Agenda for Biosecurity, August 2004
3.
While impact on biodiversity is likely to be more obvious and more
press- worthy, the impact of non-native species on ecosystem
services are likely to be more severe in the long term, more cryptic
and slower to develop. Domesticate animals and crops/garden
plants provide both pathways and reservoirs for diseases of
environmental importance, thereby linking agricultural and
environmental risks across a number of taxa.
Radial expansion (km/yr)
100000
10000
1000
100
10
1
0.1
0.01
Figure 2.10:
Aquatic
Invertebrates
Vertebrates
Plants
Plant diseases
Animal diseases
Terrestrial
Invertebrates
0.001
Median radial spread rates of the six groups included in this study.
represent maxima and minima.
Bars
A second aspect of impact involves the rate of spread of non-native species.
In Figure 2.10 we summarise the limited information we have collected on
rates of spread for different species. Not surprisingly diseases emerge as
rapidly spreading non-natives, while plants and vertebrates are slowly
spreading. However there is a surprisingly large range for some taxa, which
cautions against generalisation. The other important feature to arise from this
analysis is the potential for long lag-times in the emergence of non-native
species problems. This is clear for plants, which we suggest may take as long
as 100 to 200 years from introduction to harmful, invasive status. When put in
contrast to a plant or animal disease, where impact is substantial in less than
a year, it is easy to see that policy decisions about where to focus prevention
and eradication will be influenced strongly by the relevant timescale
envisioned by policy makers.
44
A New Agenda for Biosecurity, August 2004
Chapter 3 – an Economic Model for Non-Native
Species Introduction
We now use the biological model illustrated in Fig. 2.1 to construct an
economic model. We will do this in two ways. Firstly, we model the probability
of entry and establishment of a new species as a stochastic process.
Secondly, we assume that once a non-native species becomes established it
grows and spreads, according to a biological model, and this growth and
spread is converted to a cost as the species affect larger areas of specific
resource. We model these resources as marketable commodities, e.g. crops
or herds of livestock. As we have seen in the previous chapter, effects on
marketable commodities are only a part of the impact on non-native species.
Environmental goods are also affected. Some of these, e.g. supply of water,
or biodiversity-driven tourism, may be treated as commodities with a market
value, but others will not. We will revisit this issue of non-market values in
Chapter 5.2.3. For the moment, we will refer to affected goods, agricultural
and environmental, and to their producers, who may be farmers or, for
instance, landowners who are conserving watersheds and producing water or
recreation as an environmental good.
3.1 Formalisation
The model assumes that producers receive no assistance from public
institutions and incur all impact and control costs themselves. The model
therefore generates a predicted impact of a non-native species based on a
“what if the government did nothing to prevent or control this species?”
scenario. The risk of invasive species incursions is then, in effect, simply a
risky production parameter. It is not suggested that such a situation will
eventuate, but it is necessary to determine what the true benefits to producers
are from maintaining freedom from the harmful affects of non-native species),
and therefore how much effort should be expended on this maintenance.
Other economic assumptions are that there is one invasive organism of
concern that is not native to a particular country or region, and that this
organism has an impact on one known agricultural or environmental good in a
homogenous environment. Secondly, assume the domestic market for the
potentially affected commodity is perfectly competitive, implying product
homogeneity. Thirdly, assume that the contribution of domestic producers of
that affected commodity to the total world supply is insufficient to exert
influence on the world price, the exchange rate and domestic markets for
other commodities. On this basis there are three economic parameters used
in determining invasive species-induced producer surplus losses:
1. Total management cost increments – Production cost increases will result
from the need for additional management activities necessary to minimise
damage to or loss of the commodity. Depending on the nature of the nonnative species concerned this may involve chemical pesticide applications
(including additional vehicle and labour costs), the destruction of
45
A New Agenda for Biosecurity, August 2004
infected/infested hosts, habitat manipulation and/or biological control
techniques4.
2. Revenue losses – This will comprise firstly a direct loss of marketable
product. Despite incorporating a management programme for the new
species into normal management practice, a certain amount of production
loss may still occur through the effects of an introduced organism. This
effect may be as high as 100 per cent in some cases, while in others it
may be negligible.
Secondly, revenue losses may include the loss of export sales. In many
cases the loss of “pest-free area” status can have a profound impact on
export revenue since the ability to sell products to markets around the
world is compromised. This does not necessarily mean that all exports of
an affected commodity are lost. Although high-priced markets may be lost,
the good can often be sold to ‘second-best’ markets where a lower price is
received. The subsequent loss of earnings represents a cost associated
with establishment of the non-native species.
The timing of these costs will depend on the organism concerned. For
instance, in the case of an animal disease such as FMD all exports of
cloven-hoofed animal and animal product exports are stopped as soon as
one case is diagnosed in a country or trading region. In other cases, the
export of susceptible products is only banned from the immediate area of
infection (or areas in close proximity to an infected site), as in the case of
plant diseases such as black sigatoka of bananas. Where exported
products have been processed or refined, there may be no loss of export
revenue resulting from a pest outbreak.
3. Indirect effects - Due to their use as inputs into the production processes
of other industries, changing production environments for some
commodities can have indirect as well as direct consequences. If these
indirect effects are taken into account the impact of invasives can be far
greater than indicated by primary production losses. Consequential flowon effects from exogenous supply shocks may be captured using inputoutput tables, but are ignored here. In the case of many agricultural pests,
flow-on effects also include environmental damage sustained through pest
damage. These are perhaps more correctly termed externalities, rather
than indirect effects. They too are temporarily ignored in this theoretical
discussion, and considered later in case studies (Chapter 4).
The total area affected is the sum area predicted by the ecological model.
Total expected damage cost of an original site in time period t (ED t) is given
by:
ED t  Pi  (MDCi  Nt  At )
(7).
where:
4
No attempt is made to predict the development and availability of new and improved control
agents for resistant pests, the likely cost of these products and the capacity of pest species to
develop resistance to them.
46
A New Agenda for Biosecurity, August 2004
MDC i  marginal damage cost for non - native species i;
N t  species density at time t ;
At  area affected at time t ;
Here, the average total cost increment and total revenue loss comprises of the
factors explained above, i.e.:
MDCi  Ci  Ri
(8).
where:
Ci  increase in average total cost of production attributab le to species i;
Ri  decrease in total revenue attributab le to species i.
A constant marginal damage cost (or average damage cost) is assumed, that
can then be combined with a biological spread model.
3.1.1 Graphical representation
A static, partial equilibrium model can be used to examine the economic
implications of invasive species. For simplicity, this discussion centres on a
species that is host-specific, affecting a commodity, q. Once again, assume
the following:
(1) The species can be controlled by additional local activities, the costs of
which are borne by producers (i.e. raising the Average Total Cost (ATC) of
q production);
(2) The domestic market for q is perfectly competitive;
(3) The domestic price for q is above the ‘landed’ price of imported (identical)
product;
(4) The contribution of the UK to the total supply of q is insufficient to exert
influence on the world price, exchange rate or domestic markets for other
goods.
Consider an enterprise producing q. The production function describes the
relationship between physical quantities of factor inputs (I) and the physical
quantities of output involved in producing q given the state of technological
knowledge possessed by the producer. So, the level of output he/she
produces is some function, call it f, of I:
q  f (I )
(9).
For the moment, assume any risky factors in the production process simply
take on their average values.
Figure 3.1 provides a graphical representation of a possible production
function with and without a harmful non-native species, denoted x. Generally,
to be of biosecurity significance, x must have a negative impact on output
when established in a production area. An exception may occur where there
are human health and/or environmental implications to non-native species
introductions, as mentioned above. This will be discussed at length below,
but for now assume the only host of x is the commodity q.
47
A New Agenda for Biosecurity, August 2004
Output
(q )
f (I ) - Without Invasive
f (I )* - With Invasive
q0
Figure 3.1: The production function with and without a harmful non-native species in the
system
I
If this is 0 the case, the production Ifunction can be
seen to move
Inputs (Ito
) the right
since the quantity of inputs required to produce any given level of output
increases due to the presence of the organism. For instance, should a
producer of q have to use an additional chemical treatment to those already
used for other pest species control to produce qo, the quantity of inputs
required will increase from I0 to I1 (as Figure 3.1 has been constructed). Thus,
non-native species impact can be seen in much the same light as a negative
technological change5.
0
1
To examine the economic welfare implications of non-native species-induced
change requires some discussion about cost and revenue functions. In short,
Total Revenue (TR) for any producer supplying the market for q depends on
the quantity sold and the price (p) at which it is sold (i.e. TR = pq), while Total
Costs (TC) are a function (call it c) of output (i.e. TC = c(q)). Profit () is
simply stated as TR minus TC. Given that the price facing a competitive,
profit-maximising producer of q is dictated by the market as a whole, their
profit maximisation decision can be stated as:
max.   pq  c(q)
(10).
q
To simplify the following discussion c(q) will not be divided into its fixed and
variable components. Hence, assume fixed costs of production are zero, so
ATC equal average variable costs.
It should be noted that is not necessarily the case that the producer’s choice
of output of q will be positive. Where the minimum value of ATC exceeds the
prevailing market price it is in the interests of a profit-maximising producer to
produce no output in order to minimise losses. At prices above the minimum
5
An equivalent means of explaining Figure 3.1 is the amount of q produced by a given set of
inputs is reduced in the presence of an invasive species.
48
A New Agenda for Biosecurity, August 2004
value of ATC the Marginal Cost (MC) curve relates the grower’s profitmaximising output to price, and thus represents their supply curve, q(p).6
The supply curve for the collective industry can simply be found by
horizontally summing the supply curves of all producers supplying the market
for q. If there are n producers and the supply curve for the ith producer is
denoted qi(p), then the supply curve for the industry (Q(p)) is given by:
n
Q( p )   qi ( p)
(11).
i 1
So, this industry supply schedule, which formalises the relationship between
industry output and collective marginal costs of production, can be used to
calculate industry profit under different production conditions.
Returning now to the production functions of Figure 3.1 (with and without an
invasive species in the system), the implications of an introduction for a
grower’s profit-maximising output decision become clear. As the level of
inputs needed to produce each unit of q increases in response to costly efforts
to keep a new non-native species at bay, or at least subdued, so too must MC
and ATC. The extent of this change is represented by MDCi in equation (7).
Recalling the characteristics of c(q), the ATC curve will be U-shaped, as
depicted in the left frame of Figure 3.2. Here, two sets of cost curves are
shown dealing with both a ‘with invasive species’ (MC* and ATC*) and ‘without
invasive species’ scenario (MC and ATC).
Figure 3.2: The economic impact of a harmful non-native species – imported goods
A profit-maximising producer will choose to produce a level of output
corresponding to the point where p equals the MC of production. At this point,
6
Hence, q(p) must identically satisfy the first-order condition p  c[ q ( p )] and the and
second order condition c[ q ( p )]  0 .
49
A New Agenda for Biosecurity, August 2004
the differential between total cost and total revenue is maximised. Assuming
the prevailing domestic market price, p, is below a closed market equilibrium
price (shown here as pD in the right hand frame of the diagram), a producer
characterised by the cost curves MC and ATC would choose to produce
quantity q0 (i.e. where p = MC) and earn a profit of ABCp in the absence of
the non-native species. Once again, note that output will be positive so long
as the price received by the producer remains above the minimum value of
the ATC of production.
If all producers in the industry behave in a similar manner, the industry supply
schedule produced by the horizontal summation of each producer’s output at
different prices would resemble the curve S in the right hand frame of Figure
2.4. According to the industry demand schedule (DI) domestic consumers will
demand the quantity Q1 at price p. Of this, Q0 will be supplied by domestic
growers, and Q1 - Q0 by imports. In this situation, producer surplus is given
by the shaded area HIJ, and consumer surplus by JMN. Note that under a
domestic closed-economy equilibrium scenario (i.e. ED) producer surplus
would be the larger area HEDpD, and consumer surplus the smaller area
pDEDN. Hence, the ‘traditional’ gains from trade is shown as EDMI.
If a harmful non-native species, x, were to now enter the production region
and become established, the effect at the producer level will be rising ATC
(and MC), recalling assumption (1) above. A greater cost is now involved in
producing each unit of q after the outbreak than before it (so MDCx > 0 in
equation (7)). At the prevailing market price p the increased costs of
production would lower producer output from q0 to q* where producer surplus
is the heavily shaded area EFGp.
If the probability of x’s entry and establishment is P, then the expected loss of
producer surplus at the farm level (EDF) associated with the organism can be
expressed as:
EDF = P × (ABCp - EFGp)
(12).
At an industry level, the domestic supply curve will contract (from S to S * in
the right frame of Figure 2.4) in the face of added growing costs. Domestic
producer surplus will decline to the heavily shaded area KLJ, representing a
loss of HILK. So, the expected damage to the collective industry from x (EDI)
can be expressed as:
EDI = P × HILK
(13).
Assumption (3) above specifies that the domestic price of x is above a world
price, but what if we now reverse this assumption? If the world price is now
assumed to be above a domestic market equilibrium price, growers can earn
more revenue by selling q on the world market. The effect of a pest like x on
an exported commodity is illustrated in Figure 3.3.
50
A New Agenda for Biosecurity, August 2004
Figure 3.3: The economic impact of a harmful non-native species – exported goods
Here, the prevailing world price for q is pw. Consider the pre-invasion supply
schedule, S0. At price pw, the domestic demand schedule in the right hand
frame of the diagram reveals the industry is willing to supply Q 0, while the
domestic demand for q is only Q1. The industry can sell the residual Q0 – Q1
and earn a total producer surplus of ABC (shaded). Consumer surplus is the
area MNC. A producer within the industry characterised by the cost curves
ATC0 and MC0 in the left frame of the diagram earns a profit of DEFp by
producing and selling q0 and the price pw.
Now consider the impact of the non-native species x on the industry. Once
again, necessary changes to the production process to deal with x raise the
ATC and MC curves of a typical producer up to ATC1 and MC1. They still
receive the world price pw, but it is now only economic to produce q1, at which
they accrue the producer surplus IJKpw. Therefore, if the probability of entry
and establishment of x is denoted P, EDF can be expressed as:
EDF = P × (DEFpw – IJKpw)
(14).
The aggregate effect of x across the industry is a contraction of the supply
curve in the right hand frame of the diagram to S 1. In a closed market
situation this would result in a domestic market price of p1. But, as this is
below pw the industry can continue to supply the world market and earn a
higher amount than it would in a closed market. The heavily shaded area
GHC indicates total producer surplus. Consumer surplus is unaffected since
the price remains at pw (recalling assumption (4)), and remains MNC. Hence,
in terms of the diagram EDI can be expressed as:
EDI = P × ABHG
(15).
Note that had the contraction in supply induced by the entry of the non-native
species been much worse, it could have spelled the end for all exports of the
commodity q. If, for instance, the post-invasion supply curve resembles S2, all
exports would cease. The industry could still supply Q 1 to the domestic
market, but only earn a producer surplus of LMC. Sales of Q0 – Q1 would
effectively be lost to the effects of x. Note also that at the level of individual
producers, such a dramatic cost increase may be sufficient to push producers
out of the market if the minimum value of their ATC function were to exceed
pw .
By describing how a harmful, non-native species impacts on the behaviour of
economic agents, its strategic significance to the economy can be measured.
Using the assumption of introduction and establishment allows us to measure
the true benefit to the economy of keeping a species out, and therefore its
biosecurity significance.
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A New Agenda for Biosecurity, August 2004
3.2 Stochastic simulation
The use of stochastic simulation is becoming common in risk analyses
modelling, where there is a great deal of parameter uncertainty and variability.
Based on the graphical analysis developed in the previous sub-section, the
spreadsheet model used in the case studies to follow is designed to estimate
the expected change in producer surplus at the producer level across all
potentially affected hectares or, in the case of animal diseases, animals. By
adding the effects, an aggregate supply curve shift is estimated, from which
the change in total producer surplus can be deduced. Note that because
economic effects are measured on a per unit of time basis, and because UK
data on the value of agricultural production is collected on the conventional
basis of annual values, the standard unit of time measurement for the
stochastic simulations in the case studies is one year. It follows that in
situations where the time frame of reference for our analysis is longer than
one year then the annual results need to be aggregated over the appropriate
time period of years. Ideally, the model would also have the capacity to
explore interactions between variables that go into determining the extent of
the supply curve shift, as well as their uncertainty and variability. But to do so
would involve dynamic modelling techniques requiring considerably more time
and data for any case examined. As a practical solution to this problem,
Monte Carlo simulation is used to sample from distributions defined in a
spreadsheet model using the @Risk software package7. Each parameter is
specified as a probability distribution rather than a point estimate, and then
multiple iterations of the model run in which one value is randomly sampled
across the range of each distribution. The number of iterations used in each
of the case studies to follow is 10,000.
A lack of quantitative information means that it is not possible to fit specific
distributions for each parameter to relevant data. Parameters are therefore
specified as one of three different types of distributions:
1. Uniform – a distribution where every value across a range has an equal
probability of occurrence. Sometimes referred to as a rectangular
distribution, the uniform distribution is specified using a minimum and a
maximum value. These are only used to describe variables with extreme
amounts of variability, such as the probabilities of entry and establishment
(in the absence of a quantitative risk assessment).
2. Discrete – a distribution where only a specified number of discrete
outcomes are possible between a minimum and maximum, each of which
has a certain probability of occurrence. Each outcome in the distribution
has a value and a weight indicating the value’s probability of occurrence.
For instance, this distribution can be used to describe the number of
additional chemical treatments required to suppress an invasive species in
an affected area. If the minimum number of additional applications is 1
and the maximum 3, and all outcomes have an equal likelihood of
occurrence from year to year, the number of sprays used in any one year
can be estimated as Discrete({1,2,3},{1,1,1}). Note that the probability
7
Palisade Corporation.
52
A New Agenda for Biosecurity, August 2004
weights can sum to any number since they are normalised to probabilities
by @Risk.
3. Pert – a form of beta distribution specified using minimum, most likely and
maximum values. The range of the distribution is dependent on the
minimum and maximum values, while the most likely value determines
skewness. This is the most frequently used distribution in the case studies
to follow since it can be used to represent a range of expert opinion. For
example, a number of different scientific publications may contain a range
of values for the intrinsic rate of population increase (r) for a particular
invasive species between 3 and 7. If the most frequently cited value is 6, r
can be specified as Pert(3,6,7).
The impact of the choice of distributions used is not explored in this report
beyond conducting sensitivity analyses.
3.3 Parameterisation
Ideally, the model should be parameterised with data from the literature. As
we have seen in Chapter 2, data for biological parameters are not always
easy to obtain. For economic parameters, we may have problems estimating
the cost of non-native species management by producers, whereas market
values may be estimated from national agricultural statistics, Frequently,
therefore, we will have insufficient data to propose a particular value or
distribution of values for a parameter. In this case, a system of semiquantitative categorisation can be used to parameterise the model. This
simple process requires relevant experts to choose from a set of alternatives
to indicate that which best describes a model parameter pertaining to a
particular pest. This alternative then effectively describes a probability
distribution that can then be used in Monte Carlo simulation.
For instance, take the probability of pest entry, pent. Although an economic
analysis of a potential agricultural pest threat could be accompanied by a
comprehensive risk analysis designed to determine likely probabilities of entry
(and establishment for that matter), there is not always the evidence base to
do this. An alternative is presented in Table 2.1. Here, the pent in both the
base case and scenarios are estimated using the semi-quantitative risk
categorisation methodology outlined in AFFA (2001), presented in the table.
Consider the example of invasive species x. Assuming Britain does not
import the host of this species from areas where it is established, and is
physically located a reasonable distance from known populations, the
likelihood of x’s entry into Britain might be considered Negligible. As Table
3.1 indicates, this would mean an entry probability (in the control case) of
between 0.00 and 1.00  10-6, which can be specified quantitatively as a
uniform distribution for modelling purposes (i.e. Uniform(0,0.000001)). If a
consensus of relevant trade, climate and ecological experts believed that
trade liberalisation will open up new and efficient pathways for x to travel to
Britain, the likelihood of it entering might be re-categorised as Low. The
impact of this scenario can be estimated using the corresponding Uniform
distribution with a minimum value of 0.05 and a maximum of 0.30 (i.e.
Uniform(0.05,0.30)). So by using the model to compare a control case and a
scenario it is possible to demonstrate the extent to which the x’s biosecurity
significance is set to change over time.
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A New Agenda for Biosecurity, August 2004
54
A New Agenda for Biosecurity, August 2004
Table 3.1: Semi-Quantifiable Risk Categorisation Methodology (AFFA, 2001)
Likelihood
Descriptive Definition
Probability Range
High
Very likely to occur
0.7 - 1.0
Moderate
Occurs with even probability
0.3 - 0.7
Low
Unlikely to occur
0.05 - 0.3
Very Low
Very unlikely to occur
0.001 - 0.05
Extremely Low
Extremely unlikely to occur
0.000001 - 0.001
Negligible
Almost certainly will not occur
0 - 0.000001
3.4 Dealing with non-market (e.g. environmental) factors
As mentioned earlier, this model is built around a presumption that all impact
affects goods that can be expressed as marketable commodities. This poses
problems with the evaluation of impact on the environment, which we suspect
is significant (see Chapter 2). Some environmental goods can be given
market values. Thus, invasive tree species in rural areas may extract water
with direct measurable effects on water supplies to human populations (e.g.
van Wilgen et al., 2001). More often, however, environmental goods may not
have a clear market value. There has been, to date only limited success in
quantifying impacts of invasive species on environmental goods such as
biodiversity (U.S. Congress, Office of Technology Assessment, 1993).
Moreover, not only may an environmental good like biodiversity have a nonmarket value in terms of use, it may also have existence, bequest or moral
values which are dependant on its continued existence, and which could
extend over generations in time (Mumford, 2001).
It would therefore seem imperative to provide a mechanism to present both
market and non-market effects of harmful non-native species in comparable
For the case studies to follow, we include where possible estimates of use
value for environmental goods, and make an additional analysis of possible
“social effects”, in which we include these non-market effects relating to the
environment, as well as any other market related effects in different sectors,
such as health and social welfare. This allows us at least to identify where
commodity based effects may be most of or only part of an overall impact of a
non-native species, without presenting a consolidated impact value. In
Chapter 6 we propose a semi-quantitative approach which would encompass
non-market valuation.
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A New Agenda for Biosecurity, August 2004
Chapter 4 – Economic Case Studies
Using the bioeconomic modelling framework developed in Chapters 2 and 3,
we can examine a series of case studies to show how biosecurity significance
can be assessed. One species has been chosen from each taxonomic group,
and is not intended to be representative of that group. Each merely serves as
an example of how the methodology developed earlier in this chapter can be
applied. A mix of both established and non-established species has been
used to demonstrate the large variability between taxa and the versatility of
the model.
For each case study, a brief introduction to the organism is provided along
with the vital statistics of the UK industries it is affecting (in the case of
established non-natives) or could affect (non-established non-natives), such
as total area, gross value of production and export value. The assumptions
used in forming a control case are then outlined in detail. These refer to all
the parameters of the model, and suggest values that are representative of
current circumstances. That is, the future is assumed to reflect the present in
terms of affected industry values, the probability of entry and establishment,
average total cost increments, biological parameters, and so forth. Future
scenarios are not considered at this point. Note that introduction and spread
rates identified in Chapter 3 are of little direct relevance to models of particular
species here: in two cases we model species which are already established
(P(Entry) = P(Establishment) = 1).
In addition to parameter estimates for the economic and biological variables in
the model, a statement on social impact is also provided for each pest. This is
divided into human health, environmental and socioeconomic categories, and
a scoring derived as in Table 3.1 to indicate the severity of impact, where
there is one. Anecdotal evidence is provided where applicable to supplement
this categorisation of social impacts for different species.
The results of Monte Carlo simulation for each of the case studies are also
presented in terms of three different outputs:
i.
Cumulative distribution of the critical level of Expected Damage (ED crit)
over 20 years - represents the damage to be expected from an invasive
(and therefore the benefits of excluding that invasive) as a yearly
average over 20 years. This indicates the mean and variability
associated with the organism’s impact over time. The significance of
EDcrit will be demonstrated in Chapter 7 where it is used as a ceiling (or
critical level) of total cost for pest management projects. Here it will be
shown that if the costs of managing a pest exceed EDcrit (which
represents the expected benefit of keeping the invasive out), then a net
loss could result over time.
ii.
Area/Time or Number/Time and variability – plots the expected area
affected or expected number of animals affected over ten year intervals
beginning at year 08.
8
A larger number of reference points could have been used in these plots. In fact, yearly
estimates are possible using the simulation model, but with a large number of iterations this
slows simulation time considerably. The points 0, 10, 20 and 30 years were arbitrarily chosen
by the authors.
56
A New Agenda for Biosecurity, August 2004
iii.
Expected Invasion Impact (EI)/Time – represents the expected cost/time
relationship for an organism. As with the area/time relationship, EI is
plotted over ten year intervals beginning at year 0. So, (i) above
indicates the distribution about the 20-year value for EI. Section 2.5
discusses EI patterns in detail.
Following the results for each case study, a sensitivity table is used to indicate
the relative significance of each parameter in the model to the expected
average annual damage over 20 years (i.e. the mean value represented in (i)
above). The value of each parameter is altered by 50 per cent, and resultant
change in the EDcrit over 20 years measured as a percentage. The higher the
percentage change induced by changes in the parameter, the more critical it
is for the model. It follows that the rigor with which the parameter value is
estimated should be highest for those parameters of a high sensitivity.
4.1 Colorado Potato Beetle
Colorado Potato Beetle (Leptinotarsa decemlineata) is a serious insect pest of
potatoes, and has also been known to affect tomatoes and eggplants. It also
affects many Solanum species growing wild across Britain. Adult insects and
larvae feed on the foliage of host plants and are capable of stripping young
plants in a short period of time if uncontrolled. A characteristic black, sticky
excrement is left on the stem and leaves of affected plants. In some cases,
the insect also eats the tubers. Their distinctive orange and black appearance
makes L. decemlineata a favourite of children who can transport them from
one region to another. However, besides human-aided dispersal, the insect
relies on prevailing winds to transport it to new areas. Generally, insects are
capable of flying up to 3km, but under favourable conditions distances over
100km can be reached (Botha, 2001).
L. decemlineata is not found in Britain, but has become established
throughout the United States, Costa Rica, Cuba, Guatemala, Mexico, Canada,
and parts of Europe, Asia and Africa. If L. decemlineata were to become
naturalised in Britain it is expected to spread relatively slowly, and only affect
the potato and tomato industries.
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A New Agenda for Biosecurity, August 2004
4.1.1 Affected industries in the United Kingdom
Table 4.1: Industries affected by Colorado Potato Beetle
*
**
Affected Industries*
Gross Value of
Production (5yr
Avg.)**
Gross Value of
Exports (5yr Avg.)**
No. Ha (5yr Avg.)**
Potato
£596,800,000
£275,000
159,000
CABI (2003).9
Defra (2002B).
4.1.2 Control case
Assume that no eradication campaign is to be mounted against a
L. decemlineata outbreak in Britain in future. Instead, assume the domestic
potato industry simply chooses to live with the insect on a permanent basis.
As a result, the costs of production will rise, a small proportion of export
markets for affected products will be permanently lost, and yield losses
(despite control) will increase.
Changes in Average Total Cost
The chemical of choice for L. decemlineata control is lambda-cyhalothrin
which costs £2.25 per hectare (Defra, 2000)10. Application and monitoring
costs are assumed to be around £13.50/ha. The number of applications is
required is represented as Discrete({0,1,2}{0.5,1,0.5}). The absence of
L. decemlineata in a crop does not mean a grower is saved of additional costs
since the crop must be inspected for the insect to determine if chemical
spraying is necessary.
Induced Changes in Average Total Revenue
Yield Loss
Although chemical treatment for L. decemlineata is expected to be effective, a
small proportion (between 2% and 6%) of the population is expected to
survive to adulthood. Their effect on overall crop yields is likely to be minimal
(Harding et al., 2002). Yield losses are represented here as Pert(0%,1%,2%).
Export Revenue Loss Attributable to Loss of Pest Freedom Status
Given the distribution of L. decemlineata around the world and the relative
ease of interceptions using inspection and SPS measures, the impact of the
insect on export sales in the long run is likely to be small. Annual export
losses for affected growers are estimated as Pert(0.0%,2.5%,5.0%).
Biological Model Parameters
Table 4.2 lists the parameter values characterising the control scenario.
Table 4.2: Parameterisation – Control Case (Colorado Potato Beetle)
Tomatoes are also described as a “secondary host”, with incidence of severe attack by the
beetle is relatively rare (CABI, 2003; Defra, 2000).
10This discounts the effects of incidental L. decemlineata control from other pest management
activities.
9
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A New Agenda for Biosecurity, August 2004
Parameter
Assumed Parameter Value
P(Entry)
Uniform(0.001,0.05) (see AFFA, 2001)
P(Establishment)
Uniform(0.30,0.70) (see AFFA, 2001)
Amin (ha)
Pert(1.0,1.5,2.0)
Amax (ha)
159,000 (Defra, 2002B)
R
Pert(0,0.025,0.05)
Nmin
Pert(1,2,3)
K (Nmax)
Pert(10000,55000,100000)
Smax
Pert(70,85,100)

Pert(1.010-5,5.9510-4,1.010-3)
D
Pert(50,60,70)
Social Effects of Naturalisation
Human Health: Nil.
Environmental: Nil.
Socio-Economic: Nil.
59
A New Agenda for Biosecurity, August 2004
4.1.3 Results
X <=£0
X <=£1,039,500
95%
1.0 5%
0.8
Mean = £135,000
0.6
0.4
0.2
Figure 4.1: Cumulative distribution of the critical level of Expected Damage (ED crit) over 20
years – Colorado Potato Beetle11
0.0
0.0
0.5
1.0
1.5
2.0
2.5
3.0
3.5
4.0
4.5
Values in £ Millions
100,000
90,000
95%
Area Affected (ha)
80,000
70,000
60,000
50,000
40,000
30,000
20,000
Mean
10,000
Figure 4.2: Area/Time and variability – Colorado Potato Beetle
5%
0
10
20
30
Year
£2,000,000
£1,800,000
95%
£1,600,000
£1,400,000
£1,200,000
£
£1,000,000
£800,000
11
To reiterate, this is an average annual expected damage cost over 20 years, and is used to
£600,000
represent the biosecurity significance of individual species in all of the case studies presented
£400,000
in this section.
It represents the net present value of the damage to the economy
Mean avoided by
maintaining
the exclusion of an organism.
£200,000
5%
£0
10
20
Year
30
60
A New Agenda for Biosecurity, August 2004
Figure 4.3: Expected Invasion Impact (EI)/Time – Colorado Potato Beetle
61
A New Agenda for Biosecurity, August 2004
Sensitivity Analysis
Table 4.3: Sensitivity Analysis – Colorado Potato Beetle
Parameter
P(Entry)
P(Establishment)
Average Total Cost –
Chemical Costs per ha
Average Total Revenue Loss
– Yield Loss
Average Total Revenue Loss
– Export Losses
Area Affected Upon
Introduction (Amin)
Maximum Affected Area
(Amax)
Intrinsic Rate of Spread*
Pest Density Immediately
Upon Introduction (Nmin)
Maximum Attainable Pest
Density (K)
Maximum Number of Satellite
Infestations (Smax)
Intrinsic Rate of Satellite
Generation ()
Population Diffusion
Coefficient (D)
Change in Parameter Value
(%)
Resultant Change in
Expected Damage (%)
- 50.0
- 99.2
+ 50.0
+ 90.4
- 50.0
- 57.3
+ 50.0
+ 74.2
- 50.0
- 11.5
+ 50.0
+ 1.6
- 50.0
- 29.6
+ 50.0
+ 35.2
- 50.0
- 8.8
+ 50.0
+ 6.2
- 50.0
- 7.9
+ 50.0
+ 4.1
- 50.0
- 17.7
+ 50.0*
+ 12.8
- 50.0
- 72.6
+ 50.0
+ 51.0
- 50.0
- 17.9
+ 50.0
+ 0.5
- 50.0
- 9.4
+ 50.0
+ 2.9
- 50.0
- 8.2
+ 50.0
+ 4.9
- 50.0
- 60.3
+ 50.0
+ 35.0
- 50.0
- 67.9
+ 50.0
+ 51.6
* Sensitivity test value beyond a maximum attainable value, and is therefore purely for illustration.
4.1.4 Conclusion
British potato growers are likely to benefit from the exclusion of
L. decemlineata from the country by around £135,000 per year. This makes
the insect a species of relatively minor significance from a national biosecurity
62
A New Agenda for Biosecurity, August 2004
perspective. This assessment is based on the production cost increments
and revenue losses avoided by absence of the pest. While the insect is
excluded, potato growers are spared the expense of chemical treatments and
crop monitoring, but gain little (if anything) in terms of export sales. The most
significant variables in determining the expected benefits of exclusion are the
probabilities of entry and establishment, the average total cost increment
associated with the insect, the intrinsic rate of spread, the intrinsic rate of
satellite generation, and the population diffusion coefficient. Further research
carried out to determine likely values of these parameters will enhance the
results of this analysis, and reduce the level of uncertainty inherent in the
model. Note that these primarily relate to biological characteristics of the
insect rather than the markets it is likely to affect.
4.2 Wild boar
Wild boar (Sus scrofa) was native to the British Isles before its extinction in
the 17th century from hunting and loss of habitat. However, in the past 20
years there have been a number of escapees from wildlife parks and farms
throughout Britain, and a self-perpetuating population now exists in wooded
areas in the south east of the country. On the European continent, S. scrofa
range freely in substantial numbers and are known to cause serious
agricultural damage to some crops. They are a hardy, omnivorous species
that can survive and flourish in a diverse range of habitats (including coastal
swamps, fresh or brackish marshland, riparian environments, woodlands and
forested areas). When food supplies are low, feeding activity extends on to
agricultural land. They are a favoured hunting quarry, particularly in European
countries, where wild boar hunting is a well-regulated, prestigious and
expensive sport (Defra, 1998).
If S. scrofa continues to spread throughout suitable habitats in Britain it is
expected to have a relatively minor impact on cereal and horticultural
production. It is also expected to damage fences, and may cause human
health and environmental impacts. The animal may also pose a problem for
livestock industries by acting as a reservoir for pests and diseases, and may
therefore prolong future eradication campaigns. However, this impact has
been ignored here due to the presence of native deer populations that could
impose the same cost.
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A New Agenda for Biosecurity, August 2004
4.2.1 Affected industries in the United Kingdom
Table 4.4: Industries affected by Wild Boar
*
**
Affected Industries*
Gross Value of
Production (5yr
Avg.)**
Gross Value of
Exports (5yr Avg.)**
No. Ha (5yr Avg.)**
Cereals
£2,652,800,000
£264,000,000
32,450
Potato
£596,800,000
£275,000
159,000
CABI (2003).
Defra (2002B).
4.2.2 Control case
S. scrofa is currently estimated to occupy between 25 and 40 square
kilometres in the south east of England. Assume that no eradication
campaign is to be mounted against the existing or future population. Instead,
assume landholders take independent action on their properties to minimise
damage to production. As a result, the costs of production will rise and yield
loss will increase in the long term.
Induced Changes in Average Total Cost of Production
Production cost increases will result because of the need to repair fencing
damaged by S. scrofa. Assume that in every square kilometre occupied by
the animals there is between 5 and 10 meters of fencing in need of repair per
year at a cost of around £3.00 per metre (Forestry Commission, 2002).
Assume each metre of damage takes an average of one hour to locate and
repair per year, and that the opportunity cost of labour is £13.50 per hour.
Induced Changes in Average Total Revenue
Yield Loss
The impact of S. scrofa is expected to be relatively small in terms of overall
crop yields, particularly when food supplies in the natural environment are
plentiful (Defra, 1998). In cereal crops yield losses are specified as
Pert(0.0%,0.5%,1.0%), and in potato crops as Pert(0%,1%,2%). These
losses persist despite the on-farm efforts of affected farmers to minimise
damage.
Export Revenue Loss Attributable to Loss of Pest Freedom Status: Nil.
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A New Agenda for Biosecurity, August 2004
Biological Model Parameters
Table 4.5: Parameterisation – Control Case (Wild Boar)
Parameter
Assumed Parameter Value
P(Entry)
1
P(Establishment)
1
Amin (km2)
Pert(25.0,32.5,40.0)
Amax (km2)
20,000
R
Pert(0.02,0.11,0.27) (Defra, 1998)
Nmin
Pert(1,3,5) (inferred from Defra, 1998)
K (Nmax)
Pert(3,5,7)
Smax
Pert(70,85,100)

Pert(0.0,5.9510-6,1.010-5)
D
Pert(1,2,3) (inferred from Defra, 1998)
Social Effects of Naturalisation
Human Health: All Suidae (old world pigs), except any domestic form of S.
scrofa, are listed as dangerous wild animals under the Dangerous Wild
Animals Act 1976, as amended in 1984 and there is a risk of injury to
members of the public as a result of defensive animal behaviour. The wooded
areas that free-living wild boar live in often include public footpaths and are
used, particularly in the summer months, by camping groups, tourists and
people walking their dogs. Dog owners in particular may be more likely to
come into contact with free-living S. scrofa as a result of dog-boar interactions
(Defra, 1998). Free-living S. scrofa also present a traffic hazard, particularly
where a road dissects two areas of woodland (Defra, 1998). The human
health implications of S. scrofa are considered to be Low.
Environmental: Wild boar are a former native species but their impact after an
absence of 300 years, on current native flora and fauna is unknown (Defra,
1998). The broad diet is likely to cause negative effects on native plant and
bird populations (particularly species that nest on the ground). Their
environmental impact is considered Low.
Socio-Economic: The free-living S. scrofa in southern England may be
considered a species of biodiversity value and a reintroduction (albeit
accidentally) of a once native species or as a potential economic resource, to
generate revenue from the sale of meat and from organised hunting fees, as
is the case on the continent. The opposing argument states that they are now,
after an absence of several centuries, an invasive species and a potential pest
of agriculture, a threat to the health of domestic farm stock and a potential
65
A New Agenda for Biosecurity, August 2004
danger to people in the countryside (Defra, 1998). The net socio-economic
impact associated with S. scrofa is considered to be Very Low.
4.2.3 Results
1.0
X <=£770,000
5%
X <=£3,280,000
95%
Mean = £1,823,000
0.8
0.6
0.4
0.2
Figure 4.4: Cumulative
distribution of the critical level of Expected Damage (ED crit) over 20
0.0
years
0 – Wild Boar
1
2
3
4
5
6
Values in £ Millions
16,000
95%
Area Affected (km sq.)
14,000
12,000
10,000
8,000
Mean
6,000
4,000
5%
2,000
Figure 4.5: Area/Time and variability – Wild Boar
0
10
20
30
Year
£7,000
95%
£6,000
£'000
£5,000
£4,000
Mean
£3,000
£2,000
5%
£1,000
66
A New Agenda for Biosecurity, August 2004
Figure 4.6: Expected Invasion Impact (EI)/Time – Wild Boar
67
A New Agenda for Biosecurity, August 2004
Sensitivity Analysis
Table 4.6: Sensitivity Analysis – Wild Boar
Parameter
P(Entry)
P(Establishment)
Average Total Cost – Fence
Repairs
Average Total Revenue Loss
– Yield Loss
Average Total Revenue Loss
– Export Losses
Area Affected Upon
Introduction (Amin)
Maximum Affected Area
(Amax)
Intrinsic Rate of Spread (r)
Pest Density Immediately
Upon Introduction (Nmin)
Maximum Attainable Pest
Density (K)
Maximum Number of Satellite
Infestations (Smax)
Intrinsic Rate of Satellite
Generation ()
Population Diffusion
Coefficient (D)
*
Change in Parameter Value
(%)
Resultant Change in
Expected Damage (%)
- 50.0
na
+ 50.0
na
- 50.0
na
+ 50.0
na
- 50.0
- 5.5
+ 50.0
+ 5.2
- 50.0
- 44.6
+ 50.0
+ 43.3
- 50.0
na
+ 50.0
na
- 50.0
- 5.1
+ 50.0
+ 4.4
- 50.0
- 87.7
+ 50.0*
+ 29.4
- 50.0
- 44.5
+ 50.0
+ 44.1
- 50.0
- 0.3
+ 50.0
+ 0.6
- 50.0
- 8.4
+ 50.0
+ 4.8
- 50.0
- 0.7
+ 50.0
+ 1.2
- 50.0
- 0.2
+ 50.0
+ 0.2
- 50.0
- 22.7
+ 50.0
+ 22.6
Sensitivity test value beyond a maximum attainable value, and is therefore purely for illustration.
4.2.4 Conclusion
British cereal and potato growers are expected to benefit from the removal of
the S. scrofa population to the tune of £1.8 million per year. The animal is
therefore a pest of moderate economic significance from a national biosecurity
68
A New Agenda for Biosecurity, August 2004
perspective. In the absence of the animal, cereal and potato growers benefit
from reduced crop loss as a result of the feeding habits of S. scrofa, and
would avoid the costs associated with fence repair. The social benefits
associated with the animals’ removal are likely to be positive, but of a
relatively low magnitude. The most significant variables in determining the
expected benefits of exclusion are the average total cost increment, the
maximum infested area, the intrinsic rate of spread, the intrinsic rate of
satellite generation, and (to a lesser extent) the population diffusion
coefficient.
4.3 Potato Ring Rot
Ring Rot of potato is caused by the bacterium Corynebacterium sepedonicum,
and is one of the most serious potato diseases in Asia, North America, and
central and northern European countries. Tomatoes are also affected. The
disease causes the early death of plants, rotting of progeny tubers and
extensive yield reduction. A yellowing of the lower leaves on one or more
stems is followed by a progressive wilt and eventual death of plant stems.
The infection spreads to tubers by way of the stolons causing a cheesy,
odourless rot of the vascular ring. Tangible losses from C. sepedonicum
result from a loss of seed certification and requirements for disinfection of
equipment and stores. The disease is spread to seed tubers by mechanical
planters, elevators, dressers and other handling machinery, and so is
generally a more severe problem in highly mechanised potato-growing
operations (Stansbury et al., 2001). If C. sepedonicum were to spread to
Britain it is expected to severely damage the domestic host industries through
the need for disinfection procedures, yield losses and export market losses.
Currently around 5% of total potato sales revenue is generated by export
sales, 1.1% of which comprises of seed exports.12 13
12
Destinations for seed and ware potatoes include Spain, the Canary Islands, Sweden,
Portugal, Sri Lanka, Egypt, Israel, Algeria, Saudi Arabia, Morocco, France, the Philippines,
Oman, Georgia, Hungary, Holland, Belgium, Cyprus, Denmark and Ireland (Greenvale AP
plc, 2003).
13 At present the Netherlands is the largest international supplier of seed potatoes, exporting
approximately 750,000 tonnes of seed potatoes per annum (FAO, 2003). It is interesting to
note that C. sepedonicum is present in the Netherlands (CABI, 2003).
69
A New Agenda for Biosecurity, August 2004
4.3.1 Affected industries in the United Kingdom
Table 4.7: Industries affected by Potato Ring Rot
Affected Industries*
Gross Value of
Production (5yr
Avg.)**
Gross Value of
Exports (5yr Avg.)**
No. Ha (5yr Avg.)**
Potato
£596,800,000
£275,000
159,000
CABI (2003). 14
** Defra (2002B).
*
4.3.2 Control case
Assume that no eradication campaign is to be mounted against a
C. sepedonicum outbreak in Britain in future. Instead, assume the domestic
potato industry simply chooses to live with the disease on a permanent basis.
As a result, the costs of production will rise, a proportion of export markets for
affected products will be permanently lost, and yield losses (despite control)
will increase.
Induced Changes in Average Total Cost of Production
It is difficult to speculate as to the likely costs of necessary implementation of
crop rotation, disinfection and other sanitation practices. Disinfectants such
as quaternary ammonia, chlorine, iodine or phenol-containing compounds
applied to equipment and other contaminated surfaces for a minimum of 10
minutes under low organic load are effective against C. sepedonicum (CABI,
2003). However, estimating a likely cost on a per hectare basis is somewhat
difficult.
It is therefore estimated in relatively broad terms
(Pert(£9/ha,£18/ha,£27/ha)). The number of applications is equally difficult to
estimate, and is represented as Discrete({0,1,2,3}{0.5,1,1,1}).
Induced Changes in Average Total Revenue
Yield Loss
Yield losses in affected areas are expected to be high. This is represented as
a pert distribution with a minimum value of 10%, a maximum value of 30%
and a most-likely value of 20% per annum (CABI, 2003).
Export Revenue Loss Attributable to Loss of Pest Freedom Status
Exports of susceptible commodities account for around £25.0 million per
annum, of which £0.3 million are attributable to sales of seed potatoes.
Between 10% and 20% of these markets are expected to be lost in the long
term. However, ware potato markets may also be affected to the same
degree since tubers represent a vector for disease spread15. Processed
14
Sugarbeet has been described as a natural symptomless host. C. michiganensis subsp.
sepedonicus has been isolated from sugarbeet seed and roots (Bugbee and Gudmestad,
1988; CABI, 2003).
15 Short-term losses may be considerably higher, but to assume this will persist in the longterm may overstate losses. A seed certification scheme may successfully contain the disease
and cause minimal impact to the export of seed potatoes, as has been the case in the
Netherlands.
70
A New Agenda for Biosecurity, August 2004
potato exports will not be affected. Collectively, export losses (expressed as a
percentage of total market losses) are represented as Pert(10%,15%,20%).
Biological Model Parameters
Table 4.8: Parameterisation – Control Case (Potato Ring Rot)
Parameter
Assumed Parameter Value
P(Entry)
Uniform(0.001,0.05) (see AFFA, 2001)
P(Establishment)
Uniform(0.7,1.0) (see AFFA, 2001)
Amin (ha)
Pert(1,3,5)
Amax (ha)
159,000 (Defra, 2002b)
r
Pert(3,5,7)
Nmin
Pert(1,2,3)
K (Nmax)
Pert(10000,55000,100000)
Smax
Pert(70,85,100)

Pert(1.010-5,5.9510-4,1.010-3)
D
Pert(0,0.25,0.5)
Social Effects of Naturalisation
Human Health :Nil.
Environmental: Nil
Socio-Economic: VEERU (2003) estimate £170 million tourism industry
earnings were lost from domestic tourists who chose to travel overseas rather
than in the British countryside during the 2001 FMD outbreak, and £425 value
added forfeited from international visitors forced to go elsewhere due to
movement restrictions. Quarantine measures for C. sepedonicum would be
far less severe than those imposed for FMD, and it is doubtful international
visitors would be forced to change their travel plans. Nevertheless, public
footpaths may be subjected to periodical closures and/or decontamination
measures that may detract from the utility gained by users of these walkways.
The extent to which this occurs is unclear since the movement of
contaminated soil on clothing or footwear is not recognised as a mode of
disease dispersal (CABI, 2003).
Assuming the costs associated with
recreational inconvenience is 0.01% of the lost domestic tourist revenue lost
in the FMD outbreak, it would still constitute £42,500 per year. Socioeconomic losses attributable to the disease can generally be regarded as
Extremely Low.
4.3.3 Results
71
A New Agenda for Biosecurity, August 2004
X <=£0
1.0 5%
0.8
X <=£2,922,700
95%
Mean = £644,700
0.6
0.4
0.2
Figure
0.0 4.7: Cumulative distribution of the critical level of Expected Damage (ED crit) over 20
years
Ring6Rot
0
2 – Potato
4
8
10
12
14
16
18
Values in £ Millions
70,000
95%
Area Affected (ha)
60,000
50,000
40,000
30,000
Mean
20,000
10,000
5%
Figure -4.8: Area/Time and variability – Potato Ring Rot
0
10
20
30
Year
£12,000
95%
£10,000
£'000
£8,000
£6,000
£4,000
Mean
£2,000
Figure£-4.9: Expected Invasion Impact (EI)/Time – Potato Ring Rot
0
10
20
5%
30
Year
72
A New Agenda for Biosecurity, August 2004
Sensitivity Analysis
Table 4.9: Sensitivity Analysis – Potato Ring Rot
Parameter
P(Entry)
P(Establishment)
Average Total Cost –
Chemical Costs per ha
Average Total Revenue Loss
– Yield Loss
Average Total Revenue Loss
– Export Losses
Area Affected Upon
Introduction (Amin)
Maximum Affected Area
(Amax)
Intrinsic Rate of Spread (r)
Pest Density Immediately
Upon Introduction (Nmin)
Maximum Attainable Pest
Density (Nmax)
Maximum Number of Satellite
Infestations (Smax)
Intrinsic Rate of Satellite
Generation ()
Infection Diffusion Coefficient
(D)
*
Change in Parameter Value
(%)
Resultant Change in
Expected Damage (%)
- 50.0
- 93.5
+ 50.0
+ 58.0
- 50.0
- 35.9
+ 50.0
+ 27.5
- 50.0
- 18.0
+ 50.0
+ 11.2
- 50.0
- 31.7
+ 50.0
+ 31.7
- 50.0
- 6.1
50.0*
+ 8.8
- 50.0
- 3.5
+ 50.0
+ 5.2
- 50.0
- 14.4
+ 50.0*
+ 30.4
- 50.0
- 71.8
+ 50.0
+ 72.2
- 50.0
- 7.9
+ 50.0
+ 9.0
- 50.0
- 18.1
+ 50.0
+ 20.0
- 50.0
- 8.6
+ 50.0
+ 6.4
- 50.0
- 32.2
+ 50.0
+ 16.0
- 50.0
- 94.7
+ 50.0
+ 43.4
+
Sensitivity test value beyond a maximum attainable value, and is therefore purely for illustration.
4.3.4 Conclusion
The UK potato industry is expected to benefit from the exclusion
C. sepedonicum by around £0.6 million per year. The disease is therefore a
73
A New Agenda for Biosecurity, August 2004
pest of relatively minor importance in terms of national biosecurity
significance. This is not to say that it is of minor significance to potato
producers (particularly those exporting seed potatoes), but in terms of overall
economic benefits production gains from exclusion are expected to be
reasonably modest. In the absence of the disease, potato growers would
benefit from reductions in yield and export losses, and the costs associated
with disinfection procedures. The most significant variables in determining the
expected benefits of exclusion are the probabilities of entry and
establishment, yield loss, the intrinsic rate of spread, and the infection
diffusion coefficient. Research conducted on these parameters is expected to
reduce the uncertainty surrounding the results of this analysis.
4.4 Newcastle Disease
Newcastle Disease (ND) is a highly contagious viral disease of domestic
poultry and other birds, which is also known to cause conjunctivitis in humans.
The virus, A/PMV 1, belongs to the paramyxovirus genus of the family
Paramyxoviridae, and there are three distinct strains: Velogenic (highly
virulent), Mesogenic (moderately virulent), and Lentogenic (mildly virulent).
While velogenic ND remains exotic to Britain, avirulent strains are endemic, as
indeed they are in most countries. The virus is transmitted mainly through
direct contact between deceased or infected animals, although trade in dayold chicks and frozen carcases is also a possible vector. Mortality rates for
the virulent strain in susceptible flocks may exceed 90%, and there is a
tendency for it to appear quickly and spread rapidly. The disease forms a
major constraint to international trade, being a notifiable disease under the
OIE Agreement (AHA, 1998).
Prior to 1976, outbreaks of ND were common throughout Europe. However,
in the period between 1976 and the present day only 2 major outbreaks have
occurred. The first of these was witnessed in 1984 and involved the
destruction of 817,000 chickens across 15 countries. The second occurred in
1997 involving the destruction 388,000 chickens and 260,000 turkeys across
6 countries. This outbreak involved 11 cases in Britain, four in broiler
chickens and seven in turkey rearing flocks (Defra, 2003b). A relatively minor
outbreak of the disease occurred in Denmark in 2002.
74
A New Agenda for Biosecurity, August 2004
4.4.1 Affected industries in the United Kingdom
Table 4.10: Industries affected by Newcastle Disease
Affected Industries*
Gross Value of
Production (5yr
Avg.)**
Gross Value of
Exports (5yr Avg.)**
No. (5yr Avg.)**
Poultry and Poultry
Meat
£1,487,128,400
£163,755,700
155,745,000
Hen’s Eggs
£513,952,100
£8,360,700
35,400,000
*
OIE (2002).
**
Defra (2002B).
4.4.2 Control case
Assume that no eradication campaign is to be mounted against an ND
outbreak in Britain in future. Instead, assume the arrival of the virus triggers a
system of widespread vaccination in non-affected poultry, and the destruction
of infected poultry. As a result, the costs of production will rise and a
significant proportion of export markets for affected meat products will be
permanently. Livestock losses will be 100% when infection first occurs, after
which vaccination is assumed to keep subsequent losses to love levels. The
existence of the EU and the OIE is not assumed to hinder a vaccination
campaign16.
Induced Changes in Average Total Cost of Production
Conventional vaccines are expensive, only available in large quantities and
can be difficult to transport since they are affected by heat. However, new
vaccines, developed through projects sponsored by the Australian Centre for
International Agricultural Research (ACIAR), are temperature tolerant (or
thermostable), and have proved effective in trials under laboratory conditions
and in villages in places like Malaysia. They are also relatively cheap, safe to
both chicken and handler (i.e. overdosing causes no ill effect) and can spread
from vaccinated to non-vaccinated birds (Alders and Spradbrow, 1998).
Broiler chickens are usually vaccinated when seven to ten days of age.
Chickens kept for egg production are usually vaccinated at least three times.
The vaccine is given when birds are approximately seven days old, again at
about four weeks and a third time at about four months of age. Revaccination
while in lay is commonly practiced (Joey Farms Game Fowl, 2003).
Vaccination is assumed to cost around £5.00/100 birds (Cyril Bason Ltd.,
2003). The number of vaccinations required per year is specified as
Discrete({3,4,5,6,7}{1,1,1,1,1}).
Under ‘normal’ circumstances, these groups require a “stamping out” policy whereby
infected and possibly infected animals are slaughtered and animal movement restricted to
eliminate the virus. See OIE (2003).
16
75
A New Agenda for Biosecurity, August 2004
Induced Changes in Average Total Revenue
Yield Loss
In the first year of infect the yield loss caused by ND is expected to be 100%.
Hence, the gross value of stock is lost. This is approximately £955/100hd for
broilers, and £1,452/100hd layers. Infected livestock are replaced in the
following years by vaccinated chickens, of which between 0% and 2% perish
(i.e. Pert(0%,1%,2%)).
Export Revenue Loss Attributable to Loss of Pest Freedom Status
Assume only live poultry sales are lost as a result of ND becoming
permanently established. This is probably an overestimate of long-term
losses since vaccination will eventually lead to the eradication of the disease
in the UK. However, in absence of adequate detection of poultry that have
been vaccinated but are infected with the disease, it is assumed all live bird
sales (representing around 40% of total gross value of production) are lost in
the long term. The percentage of total poultry exports lost is therefore
specified as Pert(30%,40%,50%). No egg export losses are expected to be
lost in the long term.
Biological Model Parameters
Table 4.11: Parameterisation – Control Case (Newcastle Disease)
Parameter
Assumed Parameter Value
P(Entry)
Uniform(0.3,0.7) (see AFFA, 2001)
P(Establishment)
Uniform(0.3,0.7) (see AFFA, 2001)
Amin (hd)
Pert(100,200,300)
Amax (hd)
191,145,000 (Defra, 2002B)
r
Pert(3,5,7)
Nmin
Pert(1,3,5)
K (Nmax)
Pert(10000,55000,100000)
Smax
Pert(50,75,100)

Pert(5.010-2,7.510-2,1.010-1)
D
Pert(1,2,3)
Social Effects of Naturalisation
Human Health: ND infections in humans seldom occur, and when they do
symptoms tend to manifest themselves in the form of a mild to moderate and
sometimes-painful conjunctivitis. Very rarely have respiratory symptoms been
reported. The eyes are particularly vulnerable to infection from airborne
infective particles or rubbing eyes with hands after handling infective material.
76
A New Agenda for Biosecurity, August 2004
Infection of other systems, e.g. lungs, is most likely to occur through inhalation
of dust from faecal material (Defra, 2003).
The risk of human infection following exposure to ND is likely to be very low.
While it is possible for humans to become infected by the routes described,
such infections are very unusual. There is no evidence of any cases among
individuals involved in dealing with the disease in recent outbreaks in the UK
or elsewhere (Defra, 2003).
The Human Health implications of ND
naturalisation are considered to be Negligible.
Environmental: Wild birds are also susceptible to ND. Concern for wild bird
populations is high in Britain, and population counts are often used as a
means of assessing the state of environmental health of regions (e.g. Defra,
2002B). Any apparent impact of wild bird numbers will be valued highly.
Hence, environmental effects are considered to be Moderate.
Socio-Economic: Captive birds may also be affected by ND if the disease
were to become naturalised in Britain, but the extent of these losses is not
expected to be severe. Zoo and wildlife parks and sanctuaries will nee to
vaccinate susceptible species on a regular basis. The socio-economic
implications of the disease are assumed to be Very Low.
77
A New Agenda for Biosecurity, August 2004
4.4.3 Results
X <=£47,335,000
5%
1.0
X <=£101,920,000
95%
Mean = £80,175,000
0.8
0.6
0.4
0.2
Figure 4.10: Cumulative distribution of the critical level of Expected Damage (ED crit) over 20
years – Newcastle Disease
0.0
0
20
40
60
80
100
120
Values in Millions
Animals Affected ('000hd)
7,000
95%
6,000
5,000
4,000
Mean
3,000
2,000
1,000
Figure 4.11:
Incidence/Time and variability – Newcastle Disease
0
10
20
5%
30
Year
78
A New Agenda for Biosecurity, August 2004
£120,000
95%
£100,000
Mean
£'000
£80,000
£60,000
£40,000
£20,000
5%
£Figure 4.12:
Expected Invasion Impact (EI)/Time – Newcastle Disease
0
10
20
30
Year
79
A New Agenda for Biosecurity, August 2004
Sensitivity Analysis
Table 4.12: Sensitivity Analysis – Newcastle Disease
Parameter
P(Entry)
Change in Parameter Value
(%)
Resultant Change in
Expected Damage (%)
- 50.0
- 24.5
+ 50.0
+ 16.4
- 50.0
- 23.5
+ 50.0**
+ 17.7
- 50.0
- 5.4
+ 50.0
+ 4.0
- 50.0
- 0.1
+ 50.0
+ 0.2
- 50.0
- 44.2
+ 50.0
+ 45.5
- 50.0
- 0.9
+ 50.0
+ 1.0
- 50.0
- 3.6
+ 50.0*
+ 4.1
- 50.0
- 4.1
+ 50.0
+ 4.1
- 50.0
- 2.0
+ 50.0
+ 1.5
- 50.0
- 1.0
+ 50.0
+ 0.9
- 50.0
- 0.9
+ 50.0
+ 1.6
- 50.0
- 11.5
+ 50.0
+ 9.3
- 50.0
- 1.4
+ 50.0
+ 0.7
**
P(Establishment)
Average Total Cost –
Vaccination
Average Total Revenue Loss
– Yield Loss
Average Total Revenue Loss
– Export Losses
Animals Infected Upon
Introduction (Amin)
Maximum Number of
Affected Animals (Amax)
Intrinsic Rate of Spread (r)
Pest Density Immediately
Upon Introduction (Nmin)
Maximum Attainable Pest
Density (K)
Maximum Number of Satellite
Infestations (Smax)
Intrinsic Rate of Satellite
Generation ()
Infection Diffusion Coefficient
(D)
*
Sensitivity test value beyond a maximum attainable value, and is therefore purely for illustration.
**
Since the parameter is a probability, maximum possible test value used is one.
80
A New Agenda for Biosecurity, August 2004
4.4.4 Conclusions
The poultry industry is expected to benefit from the exclusion ND by around
£80.2 million per year, making the disease of high importance in terms of its
national biosecurity significance. By maintaining area freedom from the
disease, broiler and layer chicken farmers are spared the costs of
vaccinations, and increased mortality rates. More significantly, export losses
(particularly in terms of meat and meat products) are avoided by maintaining
the exotic status of the disease. Being of importance to the OIE, a loss of ND
area freedom may cause export losses of up to 50 per cent. Predictably, the
probabilities of entry and establishment and expected export losses were the
most significant factors in determining the biosecurity significance of the
disease in this analysis.
4.5 Gyrodactylus salaris
Gyrodactylus salaris is a small, leech-like parasite of salmonids. Only Atlantic
salmon (Salmo salar) are severely affected by this parasite, although it has
been reported to affect rainbow trout (Oncorhynchus mykiss), Arctic char
(Salvelinus alpinus), North American brook trout (Salvelinus fontinalis),
grayling (Thymallus thymallus), North American lake trout (Salvelinus
namaycush) and brown trout (Salmo trutta). Salmon from Scottish rivers have
also been shown to be susceptible to G. salaris. The parasite attaches itself
to its host by an opisthaptor at one end of the body and feeds using glands at
the other end. Attachment can cause large wounds and feeding can damage
the epidermis and allow secondary infections, particularly in severe infections
where several thousand parasites may be attached to a single fish
(FRS, 2004).
There are believed to be in excess of four hundred individual Gyrodactylus
species affecting fish and frogs, in both fresh and salt water. These parasites
reproduce in a remarkable way in that they have evolved a 'Russian doll'
arrangement. They give birth to live young, with a daughter parasite being the
same size as the mother. Inside this newborn daughter there is already a
developing granddaughter, and so the process continues (FRS, 2004).
Experiences in Norway have demonstrated that G. salaris is a particularly
severe species. Following its introduction to the country in the 1970s
catastrophic losses of Atlantic salmon were witnessed. Over 40 Norwegian
rivers have now been infected and their native salmon populations effectively
exterminated (FRS, 2004).
Regrettably, it has been extremely difficult to estimate parameter values in this
assessment, and it should therefore be treated as a rough approximation of
G. salaris’ impact on the UK economy.
81
A New Agenda for Biosecurity, August 2004
4.5.1 Affected aquaculture industries in the United Kingdom
Table 4.13: Industries affected by Gyrodactylus salaris
Affected Industries*
Gross Value of
Production (5yr
Avg.)**
Gross Value of
Exports (5yr Avg.)**
No. farms**
Salmon Farming
£265,000,000
£150,000,000
340
Trout Farming
£36,000,000
£0
265
(FRS, 2004). 17
** The Scottish Parliament (1999).
*
4.5.2 Control case
Forming a control case for G. salaris naturalisation is extremely difficult.
Assume that no eradication campaign is to be mounted against the parasite,
and that the Scottish salmon and English trout industries are the only
commercially significant industries affected18. G. salaris can not survive full
strength sea water, so is not expected to affect marine fisheries in the UK.
Farmed species such as trout are not affected by the parasite, but are
potential spread vectors. The parasite is spread by contact, implying that the
only means of a wild Salmo salar population becoming infected is through
mechanical transfer or escapees19. Once detected, assume one rotenone
treatment (or equivalent) is used on affected fisheries deemed to present a
risk to wild S. salar to eliminate all potential hosts before re-stocking20.
Rotenone is a naturally occurring pesticide obtained from leguminous plants
such as Lonchocarpus, Derris and other species. Its application to the control
of fisheries pests is not new, having been employed by fisheries managers for
over 50 years. Its persistence in the water has been shown to be lengthy,
sometimes taking two to three months to dissipate to tolerable levels in some
water courses (Bruas et al., 2002). It is assumed here that re-stocking can
only take place 6 months after treatment.
17
As mentioned above, only Salmo salar are severely affected by G. salaris, for which a
recreational value exists (in addition to other non-market values associated with its ecological
status). Trout species are the primary market affected by the pest, although the impact on the
species is generally much less severe.
18 This represents a highly simplistic method of measuring the economic impact of a
complicated invasive species problem. The results should therefore be viewed in context.
Ideally, a more rigorous ecological and economic analysis would be used, although there is
no guarantee the results would contain a higher degree of accuracy. It may be the case that
the market effects of the pest are negligible since only an environmental conscience will
provoke treatment of infected farm fish species showing little or no sign of distress as a result
of G. salaris infection.
19 Although containment techniques for farm fish have improved, about a million salmon are
believed to have escaped from farms in Scotland since 1998 (McDowell, 2002).
20 It is in fact doubtful that detection will occur early in a G. salaris outbreak in farmed trout.
Although it does occur on trout species and may be spread by them, the parasite exists in
trout populations seemingly without harm (Bellona, 2003). Moreover, the removal of affected
farmed trout will probably achieve very little in terms of the removal of G. salaris. The use of
rotenone is merely used as a proxy for change in farm management behaviour, assuming
there is one.
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A New Agenda for Biosecurity, August 2004
Salmon are farmed on 340 sites in Scotland, all of which are assumed to
present a risk to wild S. salar populations (The Scottish Parliament, 1999)21.
Trout are farmed on 265 sights on 78 river catchments in England and Wales.
Of these catchments, 49 contain wild S. salar populations and have trout
farms (Peeler et al., 2003). Hence, assume that only farms in these 49
catchments (approximately 60 per cent of the total industry in England and
Wales) would treat with rotenone upon detection of G. salaris. There are 63
trout farming sites in Scotland (SERAD, 2002). Due to the abundance of
S. salmo throughout the rivers of Scotland, assume all of these farms would
treat with rotenone immediately upon detection.
Induced Changes in Average Total Cost of Production
Assume affected farms undertake one rotenone treatment immediately upon
detection. Further assume the average volume of water to be treated per
farm is 4 hectares by 6 foot deep22. Rotenone (liquid) costs around £208 per
litre, and assume 75 litres are needed per farm. This pushes cost per farm up
to around £15,600 per treatment.
Induced Changes in Average Total Revenue
Yield Loss
Affected farms experience 100 per cent yield loss due to rotenone treatment
for a period of 6 months, after which time re-stocking can commence.
Thereafter, assume yield losses attributable to G. salaries (or indeed
rotenone) are negligible.
Export Revenue Loss Attributable to Loss of Pest Freedom Status
During rotenone treatment, affected fisheries experience a 100 per cent loss
of export revenue for a period of 6 months.
21
Around 30 of the sites are affected by Infectious Salmon Anaemia, a contagious viral
disease of salmon transmitted through water (The Scottish Parliament, 1999).
22 Obviously this is assumption involves a great deal of speculation.
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A New Agenda for Biosecurity, August 2004
Biological Model Parameters
Table 4.14 lists the parameter values characterising the control case.
Table 4.14: Parameterisation – Control Case (Gyrodactylus salaris)
Parameter
Assumed Parameter Value
P(Entry)
Uniform(0.05, 0.3) (see AFFA, 2001)
P(Establishment)
Uniform(0.7,1.0) (see AFFA, 2001)
Amin (no. fisheries)
1
Amax (no. fisheries)
340
R
Pert(0.20,0.35,0.50)
Nmin
Pert(1,2,3)
K (Nmax)
Pert(1.0M,5.5M,10.0M)
Smax
Pert(0,3,5)

Pert(0.01,0.03,0.05)
D
Pert(0.00,0.25,0.50)
Social Effects of Naturalisation
Human Health: Nil.
Environmental: Wild S. salar populations can be devastated by G. salaries, as
has been graphically illustrated in Norway where the parasite has all but
wiped out native salmon populations in affected waters. First attempts to
clear some affected Norwegian rivers were made in 1993 using a rotenone
treatment. This drastic approach eliminates all fish species in the river, after
which restocking may be carried out using eggs and juveniles collected prior
to treatment. Not all rotenone treatments have been successful. It has been
shown that the technique is only feasible where relatively short rivers with
favourable biological and geographical conditions are concerned (Bruas et al.,
2002). The environmental damage is therefore categorised as High.
Socio-Economic: Nil.
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A New Agenda for Biosecurity, August 2004
4.5.3 Results
X <=£11,520,100
5%
1.0
X <=£28,003,000
95%
Mean = £20,522,000
0.8
0.6
0.4
0.2
Figure 4.13: Cumulative distribution of the critical level of Expected Damage (ED crit) over 20
0.0
years – Gyrodactylus salaris
0
5
10
15
20
25
30
35
Values in £ Millions
600
No. Fisheries Affected
95%
500
Mean
400
300
5%
200
100
Figure 4.14:
Incidence/Time – Gyrodactylus salaries
0
10
20
30
Year
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A New Agenda for Biosecurity, August 2004
£30,000
£25,000
95%
£'000
£20,000
Mean
£15,000
5%
£10,000
£5,000
Figure£-4.15: Expected Invasion Impact (EI)/Time – Gyrodactylus salaris
0
10
20
30
Year
86
A New Agenda for Biosecurity, August 2004
Sensitivity Analysis
Table 4.15: Sensitivity Analysis – Gyrodactylus salaris
Parameter
P(Entry)
P(Establishment)
Average Total Cost –
Chemical Costs per ha
Average Total Revenue Loss
– Yield Loss
Average Total Revenue Loss
– Export Losses
Area Affected Upon
Introduction (Amin)
Maximum Affected Area
(Amax)
Intrinsic Rate of Spread (r)
Pest Density Immediately
Upon Introduction (Nmin)
Maximum Attainable Pest
Density (K)
Maximum Number of Satellite
Infestations (Smax)
Intrinsic Rate of Satellite
Generation ()
Infection Diffusion Coefficient
(D)
Change in Parameter Value
(%)
Resultant Change in
Expected Damage (%)
- 50.0
- 10.7
+ 50.0
+ 15.1
- 50.0
- 16.2
+ 50.0
+ 12.2
- 50.0
- 1.1
+ 50.0
+ 1.5
- 50.0
- 12.5
+ 50.0
na
- 50.0
- 10.5
+ 50.0
+ 10.8
- 50.0
- 0.1
+ 50.0
+ 0.1
- 50.0
+ 9.3
+ 50.0
- 6.4
- 50.0
- 15.0
+ 50.0
+ 7.8
- 50.0
- 4.7
+ 50.0
+ 2.1
- 50.0
- 4.0
+ 50.0
+ 4.2
- 50.0
- 2.1
+ 50.0
+ 1.3
- 50.0
- 1.4
+ 50.0
+ 1.5
- 50.0
- 14.3
+ 50.0
+ 8.8
4.5.4 Conclusion
G. salaris has been shown to be a severe threat to the UK salmon industry.
The expected gains from retaining area freedom from the disease are
estimated to be around £20.5 million per year, making the pest of high
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A New Agenda for Biosecurity, August 2004
importance in terms of its national biosecurity significance. By maintaining
area freedom from this parasite, costly and environmentally destructive
removal methods in fresh water systems are avoided, as are the
consequential losses in domestic and export fish sales. Curiously, none of the
parameters for which a sensitivity analysis was conducted stands out as being
highly significant in determining the biosecurity significance of G. salaris. This
would indicate that the modelling framework may not be the most appropriate
for aquatic organisms. Further research on these types of invasive organisms
is required, particularly in terms of its spread between fisheries.
4.6 Creeping Thistle
It is difficult to model a case study for plant invasions, because of their very
slow development (see Section 2.3.5). We therefore do this in an approximate
manner by taking a native weed similar to potential invasive species. Thistles
are one of the most important non-native invasive species affecting
agricultural worldwide, particularly pasture. with major government and private
sector control and eradication efforts in areas of introduction, particularly
Australia, New Zealand, USA and Canada. Creeping Thistle (Cirsium
arvense) is a native perennial in UK, which is often considered 'noxious' (in a
legal sense), and has therefore been of concern to farmers who grow cereals,
oilseeds and forage products. As it is also a major invasive weed in other
continents, we use it as case study. Because it is native, we model this as an
established non-native. Therefore, this case study should be thought of as
examining the impact, over the next 20 years of a plant species which is
already established and widely spread. However C. arvense is much more of
a problem to agriculture than most of our plant invaders, hence it is more easy
to evaluate in economic impact terms.
The weed can infest a broad range of temperate agricultural crops, and is
found in both disturbed (tilled) and no-tillage agricultural fields used for
producing most annual, winter annual, and perennial agronomic and
horticultural crops, as well as adjacent sites, (including non-cropped
undisturbed roadsides). Although the centre of origin of the weed is unknown,
it is present throughout much of the world, including Europe, western Asia,
northern Africa, North America and Oceania, as well as the UK. The seeds of
C. arvense can be dispersed by transport in contaminated crop seed, feed,
packing straw and manure, as well as by irrigation water and wind (CABI,
2003).
The abundance of C. arvense in the UK causes additional production costs
across a range of plant industries. Herbicide use is higher than it otherwise
would be, as are yield losses. In addition to affects on agriculture, the weed
has a negative effect on the environment by displacing other native and highly
favoured plant species.
4.6.1 Affected industries in the United Kingdom
Table 4.16: Industries affected by Creeping Thistle
Affected Industries*
Gross Value of
Production (5yr
Avg.)**
Gross Value of
Exports (5yr Avg.)**
No. Ha (5yr Avg.)**
88
A New Agenda for Biosecurity, August 2004
*
**
Wheat
£1,220,100,000
£93,650,800
1,996,000
Oats
£48,100,000
£5,000,000
126,000
Barley
£640,450,800
£96,086,100
1101,000
Canola
£235,498,000
£33,550,000
432,000
Cabbage
£60,900,000
£8,820,000
100,000
Carrots
£94,900,000
£22,680,000
25,720
Cauliflowers
£50,100,000
£8,820,000
10,000
Field Beans
£41,800,000
£0
164,000
Flowers and Bulbs
£43,300,000
£0
850
Lettuce
£128,526,100
£22,680,000
25,720
Potatoes
£596,800,000
£25,000,000
159,000
Raspberries
£30,900,000
£6,000,000
3,000
Strawberries
£78,400,000
£21,003,500
6,000
Tomatoes
£75,705,300
£18,994,000
20,010
(CABI, 2003).
Defra (2002B).
4.6.2 Control case
Assume that no eradication campaign is to be mounted against C. arvense
due to the sheer enormity of the task. Instead, assume domestic agricultural
industries continue to live with the weed on a permanent basis. As a result,
the costs of production will continue rise and yield losses (despite control) will
slightly increase over time.
Induced Changes in Average Total Cost of Production
Chemical costs for additional C. arvense control in broad acre crops are
based on Trifluralin (@ 0.5-1.5L/ha, approximately £2.60/L) (DAWA, 2000),
and application costs are assumed to be around £1.15 per hectare. In
intensive horticultural crops (including orchard fruit) chemical costs are based
on Ally (@ 5-8g/ha, approximately £52.20/200g) (DAWA, 2000), and
application costs are estimated at £11.50 per hectare23. The number of
applications is required is represented as Discrete({0,1,2,3}{0.5,1,1,0.5}).
Induced Changes in Average Total Revenue
Yield Loss
Although chemical treatment for C. arvense is expected to be effective, a
small proportion of the population in cultivated crops is expected to survive
and produce seed, and reinfestation from adjoining land areas that were not
23
Note the chemical cost is critical here rather than the type of chemical used.
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A New Agenda for Biosecurity, August 2004
treated continually takes place. The effect on overall crop yields is likely to be
minimal. Yield losses are represented here as Pert(0.0%,0.5%,1.0%).
Export Revenue Loss Attributable to Loss of Pest Freedom Status
Assuming widespread distribution of the weed throughout the EU and the
trading world, the impact of the weed on export sales in the long run is likely to
be negligible.
Biological Model Parameters
Table 4.17: Parameterisation – Control Case (Creeping Thistle)
Parameter
Assumed Parameter Value
P(Entry)
1
P(Establishment)
1
Amin (ha)
3,614,000 (Preston et al., 2003)
Amax (ha)
4,546,025 (Defra, 2002B)
r
Pert(1,2,3)
Nmin
Pert(1.0,1.5,2.0)
K (Nmax)
Pert(10000,55000,100000)
Smax
Pert(100,200,300)

Pert(0.050,0.075,0.100)
D
Pert(0.0000,0.0015,0.0020)
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A New Agenda for Biosecurity, August 2004
Social Effects of Naturalisation
Human Health: Nil.
Environmental: The environmental impact of C. arvense may be severe in
places through the displacement of native plant species. There may also be
positive environment impacts related to additional sources of pollen and
nectar for insect and bird populations. The net environmental impact is
therefore ambiguous, but is generally regarded as negative. Environmental
impact is categorised here as Very Low.
Socio-Economic: Nil
4.6.3 Results
X <=£7,217,600
5%
1.0
X <=£61,958,000
95%
Mean = £30,354,200
0.8
0.6
0.4
0.2
Figure 4.16: Cumulative distribution of the critical level of Expected Damage (ED crit) over 20
0.0
years – Creeping Thistle
0
20
40
60
80
Values in £ Millions
5,000
Area Affected ('000ha)
4,500
4,000
3,500
3,000
2,500
2,000
1,500
1,000
500
91
0
10
Year
20
30
A New Agenda for Biosecurity, August 2004
Figure 4.17: Area/Time – Creeping Thistle
£90,000
£80,000
£70,000
£60,000
£'000
95%
£50,000
£40,000
£30,000
Mean
£20,000
£10,000
5%
£-
Figure 4.18: 0Expected Invasion Impact
(EI)/Time – Creeping Thistle
10
20
30
Year
Note that C. arvense is widely established across Britain, so the area and
costs in year zero are greater than zero.
92
A New Agenda for Biosecurity, August 2004
Sensitivity Analysis
Table 4.18: Sensitivity Analysis – Creeping Thistle
Parameter
P(Entry)
P(Establishment)
Average Total Cost –
Chemical Costs per ha
Average Total Revenue Loss
– Yield Loss
Average Total Revenue Loss
– Export Losses
Area Affected Upon
Introduction (Amin)
Maximum Affected Area
(Amax)
Intrinsic Rate of Spread (r)
Pest Density Immediately
Upon Introduction (Nmin)
Maximum Attainable Pest
Density (K)
Maximum Number of Satellite
Infestations (Smax)
Intrinsic Rate of Satellite
Generation ()
Infection Diffusion Coefficient
(D)
Change in Parameter Value
(%)
Resultant Change in
Expected Damage (%)
- 50.0
na
+ 50.0
na
- 50.0
na
+ 50.0**
na
- 50.0
- 42.6
+ 50.0
+ 24.3
- 50.0
- 2.8
+ 50.0
+ 3.1
- 50.0
- 1.4
+ 50.0
+ 1.1
- 50.0
- 6.9
+ 50.0
+ 4.0
- 50.0
- 29.6
+ 50.0*
+ 20.5
- 50.0
- 78.2
+ 50.0
+ 76.9
- 50.0
- 10.8
+ 50.0
+ 1.1
- 50.0
- 8.0
+ 50.0
+ 2.9
- 50.0
- 1.6
+ 50.0
+ 0.1
- 50.0
- 11.9
+ 50.0
+ 1.8
- 50.0
- 5.3
+ 50.0
+ 2.4
*
Sensitivity test value beyond a maximum attainable value, and is therefore purely for illustration.
**
Since the parameter is a probability, maximum possible test value used is one.
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A New Agenda for Biosecurity, August 2004
4.6.4 Conclusions
A wide range of agricultural industries are affected by the presence of C.
arvense in the UK, and therefore stand to benefit from its removal if it were
possible. We have modelled C. arvense as a “non-native super weed of
agriculture”. If spread continues unabated from the present level of infestation,
it is estimated that the economy will be worse off by around ₤30.4 million per
year, making it a weed of high biosecurity significance. In addition to
agricultural damage, there are also social costs inflicted by the weed which
should be considered, but which are ambiguous. There are both positive and
negative environmental impacts resulting from the weed’s presence in
ecosystems. The average total cost of control and the intrinsic rate of spread
are the parameters of highest significance to the results presented here. The
costs of C. arvense are spread widely across the economy and the
environment.
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A New Agenda for Biosecurity, August 2004
Chapter 5 – Patterns of Impact of Non-native Species
The case studies in the previous sections are intended to demonstrate the
potential to take a single approach to the evaluation and comparison of quite
different invasion threats, by reducing each threat to its basic biological and
economic properties. Because many parameters have been guessed, it is
easy to challenge the predictions of each case study. It is quite likely that
Defra agencies can parameterise models for these organisms much better
than we have done here, in which case these models may have some direct
value. However, we have generated these case studies for the purpose of
comparison and analysis. We will first focus on comparing the magnitude of
impact of different biosecurity risks, and then the pattern of impact over
different future time horizons.
5.1 Comparing impact estimates between species
While these case studies are few and parameterisation difficult, there are still
some striking difference in the cost to UK of a “government do nothing”
approach for different kinds of non-native species risks. In Table 5.1, we
summarise some predictions of the case studies. Note that, as shown in the
case studies, predicted impact has broad confidence intervals, and averages
may be misleading. However, large differences can be seen between different
kinds of non-native, invasive species. Information on Foot and Mouth Disease
(FMD) is included from a case study to appear later in Chapter 6.
Table 5.1: Estimated impacts on 20 year time horizons for different species (from case
studies).
Species
Colorado Beetle
Wild Boar
Potato Root Rot
Newcastle Disease
FMD
Average Impact (£)
135,000
1,823,000
844,700
Environmental Effect
Nil
Low
Nil
80,175,000
Moderate
1,030,000,000
Moderate
Gyrodactylus salaries
20,522,000
High
Creeping Thistle
30,350,200
Very Low
Species causing losses that strictly proportional to the area they infest, like
many crop pests, have impacts ranging from 103 - 106, whereas species which
cause export losses have impacts from 106 to 109. (in the case studies;
Newcastle Disease, Gyrodactylus and FMD: potato ring rot export losses are
very small). Export market issues effects dominate non-native species impact
wherever they occur. This supports, as far as pure market economics is
95
A New Agenda for Biosecurity, August 2004
concerned, the current policy in Defra of higher investment in reducing nonnative threats to animal health, relative to plant and environmental health.
Environmental impact is not quantified in the model, but Table 5.1 shows that
it is variable. Where it is high, it will increase dramatically the overall impact of
the species. In our case studies, high environmental impact is associated with
species affecting exports, thus increasing their impact even further. It is a
feature of many non-native diseases that they will affect domestic animals and
crops as well as wild species (Section 2.3.4), generating these environmental
costs. Animal diseases may further affect human health, as zoonoses, which
will further increase impact estimates. However, other taxa may have
substantial, even dominant environmental effects, such as weeds of native
vegetation.
5.2 Comparing patterns of impact over time
So far, we have compared the possible impact of different species over the
same time horizon of 20 years. The model may also be used to consider total
impact on the UK economy over different future time horizons. Repeating
Monte Carlo simulations for separate, increasing time horizons allows
construction of an impact vs time function. We have done this in Chapter 4 for
each case study, with time intervals of 10, 20 and 30 years. Please note that
these trends are not a continuous function generated by the model, but a
series of separate, independent model predictions for discrete time intervals of
increasing length, joined up to produce a pattern for analysis. Examining
these patterns reveals important economic features of invasions which may
influence decisions about management priorities. On the basis of theoretical
arguments and case study observations, we now examine three possible
impact vs time patterns, and compare them.
5.2.1 Constant expected impact increments over time
Firstly, consider the examples of Colorado Potato Beetle (Leptinotarsa
decemlineata), Wild Boar (Sus scrofa) and Potato Ring Rot (Corynebacterium
sepedonicum). Associated with these three pests is an almost linear
relationship between cost/impact and time. That is, between the 10, 20 and
30-year time intervals the degree to which expected invasion impacts increase
remains relatively constant.
Expected Total Invasion Impact
Figure 5.1 provides an abstract view of the flow of expected impact over time
for pests like these. Clearly, the change in impact is identical between years
0-10 (EI10), 10-20 (EI20) and 20-30 (EI30). The resultant expected impact
curve is linear (EIA).
EIA
EI30
96
EI20
A New Agenda for Biosecurity, August 2004
Figure 5.1: Constant expected impact increments over time
Interestingly, all three of these pests can be classed principally as crop pests
in that their major economic impact is associated with their effect on crops of
one type or another.
In addition to the probabilities of entry and
establishment, (with the exception of wild boar, already introduced in our
model) and yield loss, the parameters with the highest sensitivities in the
quantitative analyses were biological in nature for all three cases. In particular
the intrinsic rate of spread (r) and diffusion coefficients (D) were of a high
sensitivity.
It should be noted that a linear EI curve does not imply a linear biological
spread pattern. Recall from Section 2.1 that the spread model used in each of
the case studies to generate expected prevalence over time is non-linear.
The reason why we observe linear EI curves is attributable to the process of
discounting future impacts.
Both private and social impacts are captured by the EI curve, and both of
these impact categories are discounted, but for different purposes. A discount
rate is applied to private contexts to reflect the opportunity cost of investment
decisions. Assume that a farmer is spared £100 worth of crop damage due to
the exclusion of an invasive species in one time period. In the following time
period, the farmer has the option of putting this additional revenue back into
cropping, or to invest it in something else. For instance, he/she may choose
to put the money into stocks, shares or bonds and earn an interest rate of, say
10 per cent per annum. This rate of interest that could be earned by the £100
is an opportunity cost of reinvesting in cropping. So, £100 worth of crop
damage prevented in the second time period is only worth £90 in the current
time period due to discounting with this private discount rate.
The social discount rate is harder to define. In the social or government
context discounting reflects the view that future generations will be "better off"
than the current generation. Technological progress is making the production
of goods and services cheaper over time, and real incomes are rising. So,
£100 to the average person in the society of 20 years time is worth less than
£100 is to the average person in society today. It follows that the social
benefits of biosecurity policies accruing in the future should be discounted.
The problem is how to choose a social discount rate.
5.2.2 Diminishing expected impact increments over time
In the Newcastle Disease (ND) case study, a distinctly different expected
impact relationship was observed. Between the 10, 20 and 30-year time
intervals expected impact rises at a steadily decreasing rate. Figure 5.2
provides an abstract view of the flow of expected impact over time for pests
like ND. Here, the change in expected impact between years 0-10 (EI10) is
97
A New Agenda for Biosecurity, August 2004
Expected Total Invasion Impact
larger than in the 10-20 year interval (EI20), which in turn is larger than the
change in expected impact over the 20-30 year interval (EI30). The resultant
expected impact curve is labelled (EIB).
EIB
EI30
EI20
EI10
Figure 5.2: Decreasing expected impact increments over time
0
10
20
30
Year
The variables of the highest significance in the ND case study in addition to
the probabilities of entry and establishment was total loss of exports. These
revenue losses are felt immediately upon ND detection, and so have their
impact on the economy as soon as the disease appears. Remember, as this
is a “government do nothing” model, of maximal impact, once the non-native
disease is established, it is not eliminated, only mitigated by private (i.e.
producer) action. Hence, the explosive growth in the expected impact occurs
following first establishment, with the loss of export earnings, while thereafter,
the change in impact is associated largely with spread, and ultimately the
impact is discounted, as in our first model. Thus, impact increases at a slower
and slower rate, resulting in an expected impact curve the shape of EIB in
Figure 5.2. A similar, but less dramatic pattern of impact over time is found in
case studies for the fish parasite, Gyrodactylus salaris, and FMD (see Chapter
6). For both of these, sensitivity analysis has also shown that loss of exports is
an important variable in determining impact, relative to other model variables.
5.2.3 Increasing expected impact increments over time
Recall that non-market value information was not included in the quantitative
case study analyses24. Non-market effects, such as effects on the
environment, may change the magnitude of predicted impact, as we have
discussed in Section 5.1. From Chapter 2, we have seen that such
environmental effects may be associated with a wide range of taxa which may
pose non-native species risks by affecting biodiversity and ecosystem
services. In our case studies, we have identified possible non-market
environmental effects in case studies on Wild Boar, Newcastle Disease, FMD,
24
See section 8.2 for an explanation of how semi-quantitative non-market information used in
resource allocation decision-making.
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Creeping Thistle and, particularly, Gyrodactylus salaris. But, will non-market,
environmental effects influence the pattern as well as the magnitude of
economic impact over time? We suggest that it will, and further that it will
generate a third type of impact curve, that of increasing expected total
invasion impact costs over time. Our argument is based on two elements:
1. Income elasticity of demand with respect to environmental
benefits. The income elasticities for environmental goods are thought to
be large and positive. That is, the value of specific environmental goods is
expected to increase as incomes rise25. Hence, the benefits of avoiding a
given amount of environmental impact from a non-native species will be
greater in 20 years than now, and greater still in another ten years.
Comprehensive empirical evidence for such a pattern of income elasticity
is currently lacking (Whitby, 2000), and two reasons have been forward as
to why this might be the case. Firstly, there is a tendency for a strategic
misrepresentation of preference when expressing utility derived from
environmental goods, and secondly demand for specific environmental
assets (or marginal changes in the health of an environmental asset) tend
to be embedded within stated or revealed preferences for much broader
environmental issues (Whitby, 2000; Kristrom and Riera, 1996). By
contract, agricultural goods may show opposite elasticities26.
2. Supply and demand. As invasive species spread, their impact reduces
environmental goods and benefits which are finite, e.g. the abundance of a
rare species or habitat. As supply of these goods and benefits declines
relative to demand, the price of environmental asset will have a tendency
to rise27. For instance, as red squirrel becomes increasingly rare with the
spread of alien grey squirrel, the value to recreation and tourism of each
red squirrel colony or site to society grows. If the supply of these benefits
falls short of demand, prices will rise. Using stated preference techniques,
Costa and Kahn (2003) present evidence suggesting this may be the case
for non-market goods, including environmental amenities. In addition to the
supply-sensitive “use value” of environmental goods are non-use values,
including “existence values” which also increase as species and habitats
become rare and threatened with extinction by invading species. The
reason that such supply and demand arguments do not apply as strongly
25
This is not to say that there is a tendency for people in lower income brackets to be
ignorant of the natural environment. On the contrary, lower income groups tend to recognise
the individual value of specific environmental assets, and consequently the opportunity cost of
investing in the preservation/rehabilitation of one at the expense of others (Kristrom and
Riera, 1996).
26 It is likely that the expenditure share of market goods such as agricultural commodities will
decrease over time as income rises. Consider the example of chicken meat, as in the ND
example above. As income levels rise and the range of substitute meat products increases
through the effects of global trade, there will be a tendency for many consumers to switch
from chicken to alternatives, such as aquaculture products. Note, however, that this this
possible fall in the economic significance of agricultural products has not been factored in to
the quantitative analysis of section 2 due to perceived increases in the general level of
population. We have chosen, arbitrarily, to offset any increased discounting of agricultural
goods by an increase in the number of consumers of the product in a global marketplace.
27 By the same token, if the supply of environmental benefits outstrips demand, prices will
have a tendency to fall.
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A New Agenda for Biosecurity, August 2004
to agricultural commodities is their greater capacity for substitution. For
instance, while a fall in lamb production in Snowdonia resulting from nonnative weed invasion may have little effect on local lamb prices, due to
importation of meat from elsewhere, the disappearance of a rare native
mountain flower from Snowdonia as a result of that same invasion cannot
be substituted externally.
Expected Total Invasion Impact
Figure 5.3 demonstrates the impact of these effects conceptually. Consider
the case of an invasive species with a negative effect on environmental
goods, like Gyrodactylis salaris. The change in the expected impact of such
species between years 0-10 (EI10) is less that that occurring between years
10-20 (EI20), which in turn is less than that occurring between years 20-30
(EI30). The slope of the resultant total impact curve (EIC) is positive and
increasing at an increasing rate. Note that this curve shape implies that the
effect of the income elasticity of demand and price rises for environmental
goods override the effects of discounting.
EIC
EI30
Figure 5.3: Increasing expected impact increments over time
EI20
EI10
0
10
20
This argument
is made on the
basis of economic
properties of30environmental
Year
invasives. There is an additional, biological reason why we might see
impact
taking this shape with environmental invasives. When a non-native species
has a harmful impact on a harvested commodity, this is likely to be felt as
soon as production of that commodity begins to decline. Further the additional
costs of controlling that pest species, e.g. by pesticide application, are often
incurred in response to these early losses. For species invading the
environment, where the impact will probably be on biodiversity or ecosystem
services, it is likely that a proportionately greater amount of spread and
damage must be incurred before a negative effect is perceived and acted on.
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A New Agenda for Biosecurity, August 2004
Further, many environmental invasives, such as weeds, may spread relatively
slowly, which would also contribute to a slow, accelerating pattern of impact at
different time horizons.
Although easily presented conceptually, difficulties remain in quantifying
environmental impact over time, which means that generating and studying
such patterns using models like ours will be difficult. There has certainly been
a marked increase in research related the environmental valuation since the
early 1990s. The catalyst for much of this work came in 1989 when the oil
tanker Exxon Valdez struck Bligh Reef in Prince William Sound, Alaska,
spilling more than 11 million gallons of crude oil. This immense spill
endangered millions of migratory shore birds and waterfowl, as well as many
species of marine mammals. In response to public outcry over an
environmental catastrophe of this magnitude, the number of environmental
valuation studies increased dramatically.28
However, while much of this research has been effective in estimating use
values, including values associated with morbidity and mortality, recreation
values and property value changes, it has been less successful in eliciting
non-use values (Adamowicz, 2004). The heterogeneity of preferences based
on non-use values, and indeed the heterogeneity of specific sites and
marginal changes in their condition present major obstacles for environmental
and ecological economists. The use of behavioural economics and stated
preference techniques may yield some answers in future by modelling
preferences in the context of respondent memory, attitudes and opinions.
This is set to become a particularly exciting area of research involving
attitudes to environmental sustainability issues, ecosystem and species
uniqueness, irreversibilities and irreplaceabilities. But, to date there is no
practical means of eliciting non-use values for environmental goods accurately
and cost-effectively.
5.3 Cross-over effects and variability
Having identified three distinct expected impact curve shapes, it is possible to
see how a policy-maker might begin to make decisions and prioritise invasive
species according to the flow of species impact over time. Conceivably,
impact-minimising/benefit-maximising policy makers may be faced with the
situation where the optimal biosecurity risk management strategy is quite
different for different time horizons. It may be that expected impact curves
related to different species cross over one another at some point in time.
Assume, for instance, a policy-maker is concerned with three hypothetical
invasive species, A, B and C, characterised by the respective expected impact
curves EIA, EIB and EIC shown in Figure 5.4. Further, assume they have
sufficient information to track the flow of invasion impact accurately over time,
including environmental effects. So, variance around the mean is negligible.
An optimal risk management strategy will largely depend on the length of time
policy-makers consider relevant.
28
The number of publications concerning environmental valuations jumped from lass than 10
in 1990 to almost 100 by 1991. By 2003, the number was over 470. Of these, stated
preference techniques demonstrated the highest growth in number (Adamowicz, 2004).
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A New Agenda for Biosecurity, August 2004
If an appropriate time horizon for biosecurity policy-making were deemed to
be 10 years, an impact-minimising strategy would clearly involve the targeting
of pest B. The explosive impact of this pest early in the time horizon gives it
high biosecurity significance in the short term, so making its exclusion a policy
priority the invasive impact reduction is maximised. Pests A and C have
significantly lower significance in the short term, so there appears to be little
strategic merit in targeting biosecurity policies towards them over a 10-year
timeframe29. By year 18, the EIC curve crosses the EIA curve as the presence
of pest C begins to cause economic damage, but by year 20 the benefits of
targeting pest B still outweigh those of other invasives. However, by years 25
to 30 the situation is quite different. The EIC curve has now crossed over the
EIB curve, and pest C now represents the species of greatest biosecurity
significance to the region. Hence the time horizon for policy-making clearly
determines the focus of resource allocation.
Expected Total Invasion Impact
EIC
EIB
EIA
Figure 5.4: Cross-over effects
0
10
18
20
25
30
Year
Information constraints involved in modelling invasive species impacts
invariably mean policy-makers are often faced with a much broader range of
estimated impacts over time on which to base decisions. Hence, the mean is
often a relatively poor representation of the distribution of expected invasive
impacts over time. Modelling the behaviour of non-indigenous species
introduced to new environments typically involves the use of broadly defined
parameter distributions rather than point estimates due to a lack of reference
material. This is demonstrated in the case studies presented in Chapter 4.
The expected impact curves in this section contain plots of the 5 and 95 per
cent confidence intervals in addition to the mean.
29
If pest A can be thought of as a typical disease of crops, pest B as an environmental
invasive and pest C as an animal disease, the optimal risk management strategy begins to
look familiar. In the short term, the explosiveness of animal diseases (particularly so called
“OIE diseases”) creates a great deal of policy interest. However, over a much longer time
horizon, the significance of other pests, particularly those with adverse effects on
environmental amenities, can increase dramatically. A failure to control such pests early
effectively passes the burden of control on to future generations. This explains why many
environmental invasives are virtually ignored by biosecurity policies for considerable periods
of time. Once they begin to feature prominently in biosecurity risk profiling for specific
regions, as in the case of pest C by the 25-year mark of Figure 6.4, they can be extremely
difficult to control, and impossible to eradicate.
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A New Agenda for Biosecurity, August 2004
If the same information is now provided for the hypothetical pests A, B and C,
the decision of which to treat as a biosecurity priority becomes somewhat
more complicated. Figure 5.5 shows why this is the case. If only the mean
expected impact curves are considered the situation has not changed as far
as the decision-maker is concerned. Organism B has the highest invasion
impact until year 25 after which it is surpassed by C, and organism A has the
second highest impact up until year 18 when it is surpassed by C. But, if the
variability of impact estimates are taken into account, indicated by the broken
lines either side of EIA, EIB and EIC, prioritising becomes dependent on the
decision-maker’s attitude to biosecurity risk. For instance, the impact of
organism C may exceed that of B by as early as year 18, so a decision made
over a 20-year time horizon by a risk averse policy maker may lead to C being
more intensely targeted with biosecurity resources than B. Similarly, the
impact of A may be exceeded by that of C as early as year 10.
On the other hand, the cross over between C and B may occur as late as year
32. In light of this information, a policy-maker looking over a 20-year time
horizon may choose to target resources towards the exclusion of B at the
expense of A and C. Note that a decision-maker is indifferent between A and
C over 20 years.
Expected Total Invasion Impact
EIC
EIB
EIA
Figure 5.5: Cross over effects with variance included.
0
10
18
20
25
30
32
Year
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A New Agenda for Biosecurity, August 2004
Chapter 6 – Horizon Scanning and Impact Levels
What likely future changes in the next 20 years will have the greatest effect on
the magnitude of non-native species risks? Our approach to horizon scanning
is to consider what changes may affect the key parameters of our general
model, and in what direction. In Table 6.1, the left hand column shows these
model parameters as they affect the introduction, development and impact of
non-native species invasions. In our case studies (Chapter 4), sensitivity
analysis revealed that parameters relating to introduction, spread and impact
were key for at least one of the six invading species examined.
Table 6.1: The relationship between model processes, their drivers and response
Model Processes
Specific Drivers
INTRODUCTION
Entry
Trade
Transport
DEVELOPMENT
Establishment
Growth and Spread
Climate
Land Use
IMPACT
Direct
Export-related
Other (e.g. non-market)
Market
Values
Responses
Exclusion
Interception
Detection
Containment
Monitoring
Eradication
Control
Acceptance
Adaptation
GENERAL DRIVERS: Technology, Education, Resources, Policy
In the second column of Table 6.1, we suggest specific drivers that will
influence these processes. That is, a change in a particular driver over the net
20 years, such as climate or the markets for commodities, will change the
parameter values for particular processes in our model. Drivers may affect
each other. Hence a change in values, e.g. how we subsidise agriculture, may
affect markets or land use. A change in markets, e.g. for cereals, may also
affect land use (area cultivated) and the level of trade (grain imported). In our
view, changes in all of these drivers are likely to emerge in the next 20 years.
To draw out how future trends might affect these drivers, we first consider
current “futures thinking”. The following likely future trends are drawn from
futures literature (e.g. The Foresight Programme, Prime Minister’s
Performance and Innovation Unit) and from other Defra Horizon Scanning
initiatives (Rural Futures; Sustainable Rural Policy and Land Use
(SURPLUS)):

Climate will change worldwide, with substantial effects on access to
water and on local agricultural production. In UK, climate will either
warm or cool, but the current view is that by 2050 the south of
England will have a climate similar to the south of France, with an
increase in extreme weather events.

Society will become more sensitive to and protective against “shocks”,
e.g. financial crashes, environmental disasters, new diseases,
terrorist actions. A culture of risk management and as a perverse
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consequence, passing risk away from institutions (government and
industry) to the individual, may spread (Power 2004).

World trade and capital flows will continue to increase. Global trade
negotiations will continue to reduce transaction costs and dissemble
trade barriers. International business, and an expanded EU and will
increase the movement of people in and out of UK.

There will be a decline in agriculture as a sector, a move to fewer,
larger farms, and towards growing reliance on imported produce,
partly but not substantially offset by a growth in the “local foods”
market.

A “rights culture” will focus more on the individual than on society, and
there will be less acceptance of government action on behalf of
society, e.g. the proactive eradication of new non-native species.
Political devolution in UK will further reduce centralised decision
making regarding public threats, but this may be countered by a
growing, centralising powers in the EU, or even in a global
environmental body.

Postmodern values in UK will place more emphasis on quality of life.
This will lead to differentiation of the countryside, to create more
recreational opportunities, accompanied by counter-urbanisation.
Individuals will invest more time in voluntary work and advocacy of
specific issues, probably raising interest in animal rights,
conservation, food and food safety. At the same time, people will
holiday more abroad, gaining different perspectives on native and
non-native species and what makes an attractive environment.

There will be a decline in a sense of national identity, with people
feeling more part of an international society, with implications for how
society will regard non-native species, and “alien-ness” in general.

Our approach to human health will move from one of “diagnosis and
cure” to one of “predict and prevent”. We may reflect this trend onto
our approach to conservation and countryside management. In the
face of new disease threats, vaccination may be considered more
humane and sophisticated than destruction of invaders or affected
populations. Technological advances will be key to this shift towards
prediction and prevention, particularly a capacity for remote sensing
and monitoring.
We can now identify links between these possible future trends and the
proposed drivers in Table 6.1. Attitudes towards risk, a “rights culture” and
prevention vs cure relate particularly to prevention and management issues in
the third column, and we will consider these in Chapter 7. Of the others,
climate change may affect both the rate at which species establish in the UK
and their pattern of spread, but how? Growing trade and markets will
influence the introduction of new species, but also the future of local
agriculture, and hence the economic value of protecting crops and livestock
against non-native species. Decline in agriculture, possibly driven by changes
in trade, will affect land use and therefore the process of invasion. Finally,
there are issues of changing interest in the countryside, the environment and
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issues of national identity that raise the social questions of whether nonnative species, as they spread, will be seen as reducing or enhancing the
quality of life.
We will now look in more detail at these three areas: climate change, trade
and markets, and social questions in more detail, to identify the direction in
which these future factors may drive impact of future non-native species
invasions.
6.1 Climate change
The UK Climate Impacts Programme has reported in some detail climate
predictions for the UK over the next century (Hulme et al. 2002). Predictions
of relevance to the biological parameters in our model include changes in
temperature, precipitation, cloud cover, snowfall, soil moisture and the North
Atlantic Oscillation (Table 6.2). These climate predictions are likely to
influence the intrinsic population growth rate of populations through: increases
in the length of the growing season of many invertebrate and plant taxa;
increases in the fecundity of many taxa; increase in the annual number of
generations or broods of animals, and decreases in development time of
invertebrates (see Table 6.3). The other main impact is likely to be on the
probability of establishment of non-native species as winter conditions
become less harsh and the range limit of many taxa moves north (this also
suggests that the range limits of native or naturalised species may change).
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A New Agenda for Biosecurity, August 2004
Table 6.2: Principal predictions of the UKCIP02 (Hulme et al. 2002), and their hypothesised
influence on model parameters. Confidence level: High, medium or low, is a
qualitative assessment of the reliability of these predictions given by UKCIP.
Predictions
Variable
Temperature
Confidence level

Annual warming 0.1 to 0.5 ºC
per decade

Greater summer warming in
SE than NW
H

Greater warming in summer
and autumn than winter and
spring
L

Greater night time than day
time warming in winter

Greater day-time than nighttime warming in summer
H
L
Expected influence on model
parameters
↑ Pest of many invertebrates,
reptiles (increased rates of
overwinter survival) and
some plants (range limits
move north).
↑ r, due to extended
growing season in plants
and increased voltinism in
many
invertebrates,
potentially increased winter
survival of other groups.
L
Precipitation
Cloud cover
Snowfall
Soil moisture
North
Atlantic
Oscillation

Wetter winters for all UK
H

Drier summers for all of UK
M

Reduction in summer and
autumn cloud, especially in S
L

Increase in winter cloud
L

Totals decrease everywhere
H

Long runs
winters
snowless
M

Decreases in summer and
autumn in SE
H

Increases in
spring in NW
and
M

More positive – more wet,
windy, mild winters
L
of
winter
↓r - Dry summers may
reduce r in summer-growing
plants and some plant
pathogens
↑ r - Increased radiation in
summer will increase plant
growth.
↑ Pest - Increased winter
survival of perennial and
winter annual plants and of
most animal taxa.
↓r - Dry summers may
reduce r in summer-growing
plants and some plant
pathogens
↑ Pest - Increased winter
survival of perennial and
winter annual plants and of
most animal taxa.
Many studies have looked at the impact of climate (Table 6.3) on the species
geographical ranges and it is clear that substantial changes are likely to occur
across the UK. A combination of species loss and in invasion of new species
will likely result in a large among of compositional change in communities (e.g.
Bakkenes et al. 2002). Many such studies are based on climate envelope
mapping, which predicts species range changes based on their climatic
tolerance. While this is a possibly a good prediction of future equilibrium
conditions, an important consideration is the action of biological constraints
during the transient dynamics of the immediate future (indeed, if climate
change continues indefinitely then transient dynamics will be the norm for the
foreseeable future). For example, the rate at which northern range limits
change will be determined, in part, by dispersal characteristics (Cain et al.
2000; Whitlock and Millspaugh 2001) and species with obligate interactions
with other species (e.g. many parasitic or specialist consumer species) will be
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constrained by the range of their symbionts (Davis et al. 1998; Baker et al.
2000). As a consequence of these constraints, there is also likely to be a lag
between species loss and the appearance of “replacement species”, which
will make ecosystems more susceptible to invasion by non-native species
(Manchester and Bullock 2000). Moreover, the early invaders will be a biased
subset of the species capable of tolerating the new climatic conditions: those
with faster dispersal without obligate symbionts (Malcolm et al. 2002).
Table 6.3: Survey of the ISI publications database with search terms: climate change and
(species invasions or species range). The entries don’t represent a review of
literature related to the effect of climate change on life-history parameters of
species in general, but in particular those referring to invasive species and
changes in geographical range
Specific predictions related to warming
↑ phenology and voltinism in butterflies, (Bryant et al. 2002)
↑ northern range limits of many ocean taxa (Scavia et al. 2002)
↑ in colonisation and intrinsic growth rate of mosquitoes (Alto and Juliano 2001)
↑ range, ↑overwintering, ↑ voltinism,
↑ growing season of insect pests (Porter et al. 1991)
↑ of oyster disease in Chesapeake Bay (Cook et al. 1998)
↑ density and range of ticks in Sweden (Lindgren et al. 2000)
↑ brood number in tit species (Visser et al. 2003)
↓ larval development time in spittal bug (Whittaker and Tribe 1996)
General predictions of multivariate climate scenarios
Plant ranges shift NE loss of ~ 1/3 of species – more stable in N and W. up to ~1/3 new
species (Bakkenes et al. 2002).
↑ susceptibity of ecosystems to invasion (Manchester and Bullock 2000)
Knotweed and Himalayan Balsam move 5º N for 1.5ºC (Beerling 1993)
↑ Butterfly N range margin (Hill et al. 1999)
↑ N limit of Colorado beetle and mistletoe (Jeffree and Jeffree 1996)
Tidal pools in US – ↑ southern species, ↓ northern species (Sagarin et al. 1999)
Large range changes expected in forest trees in Europe (Sykes and Prentice 1995, 1996;
Sykes et al. 1996)
Evergreen trees become more competitive with ↑ growing season – Switzerland (Walther
2002)
Range changes of Karnal bunt and Colorado beetle (Baker et al. 2000)
Hence, we predict that climate change will see an increase in the probability
that non-native species will establish as species ranges shift north and native
communities become more susceptible to invasion. Non-native species
themselves may be more likely that native species to be adapted to, and
invasive in, locally changing ecosystems characterised by climate change,
because species with such biological properties will have been selected by
the introduction process itself. In general, climatic mapping studies predict that
the pool of potentially invasive non-native species will increase. Finally, It
should also be noted that climate change is also likely to induce changes in
spatial extent and distribution of habitat types, both natural and agricultural.
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The affects of these changes, and consequences for non-native species
spread, may be substantial, but the processes are too difficult to predict in
advance.
6.2 Trade and markets
We explore here the trend towards greater trade liberalisation and its
biological (rate of introduction of non-native species) and economic (value of
agriculture, land use) impact.
6.2.1 A conceptual model for trade and introduction
A biological impact of increased trade could be the introduction of more nonnative species. This is logical but simplistic. A regression of global trade
volume in any commodity over recent decades against generally rates of
introduction of new species is likely to be positive, but not evidence of
causation. We need to know more about pattern – how quickly, from where,
by what means will new species arrive? This can be done for a particular
species and its likely pathway of introduction (e.g. VLA 2003). Such taxon by
taxon analysis is beyond the scope of this project. Instead, we explore here a
general, conceptual model for how changing trade will affect introductions.
For this discussion, “trade” effects include all anthropogenic means by which
new species are brought into a country. Invasion specialists often refer to the
“four T’s”: trade, travel, transport and tourism, to describe introduction
pathways.. New species enter countries on commodities and goods, with
people and their possessions, on transport (e.g. ships, planes) and in
containers used for transport (e.g. wooden crates). Further, many non-native
species risks are commodities themselves, and are intentional introductions,
such as exotic pets, garden plants, and new varieties of game fish and fishing
bait. A number of studies have considered in some detail the diversity of
pathways by which non-native species might be introduced (Defra 2002; ISAC
2002).
In natural, non-anthropogenic processes of species colonisation, species from
a potential “global species pool” enter a new region through a series of “filters”
relating to dispersal factors (ability to arrive in the new habitat), abiotic factors
(tolerance to the new physical environment) and biotic factors (interactions
with other species) (e.g. Belyea and Lancaster 1999). Each of these filters
defines a subset of the global pool, such that the overlap of all subsets defines
the actual species pool capable of introduction. Species pool theory can be
modified to include anthropogenic “trade” impacts on community assembly by
replacing the dispersal constraints with the constraints of trade sources (e.g.
the trading partners of the recipient country) and constraints of the transport
process from old to new regions. This is illustrated in Figure 6.1, where each
filter is favourable to a certain subset of species, and the potential pool is the
overlap of these subsets. While biotic and abiotic constraints govern the
likelihood of establishment following introduction, source and transport
constraints govern the likelihood of species arrival.
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A New Agenda for Biosecurity, August 2004
Figure 6.1: A species pool model in which the pool of potential non-native species is defined
by the action of abiotic, biotic, trade and transport constraints
The rate at which species move from the actual species pool into a country
like UK depends on the volume of trade moving from these sources down
these transport routes over time. In one of the few studies relating trade
volumes to non-native species introductions, Levine and D’Antonio (2003)
postulate that the sampling of the actual species pool over time by continuing
trade will reach an asymptote which is the size of the pool – eventually all of
the species which can be introduced will be, and this will happen more quickly
if trade volume itself is increasing over time. Their model fits well the
relationship between cumulative agricultural imports and cumulative nonnative species for the US, indicating this asymptotic effect.
Such a maximum number of species assumes a fixed pool. If the pool itself is
changing in size, due to a growth in sources or a change in transport which
puts more species in the potential pool, then our model would predict an even
greater rate of new species introduction. Let us now consider each factor
affecting pool size and its likely future trends:
Abiotic – changing climate will change the potential pool. Section 6.1 suggests
that it will actually increase the pool size.
Biotic – invasibility of ecosystems is affected by biotic factors. Agricultural
systems have always been particularly invasible. Natural ecosystems become
more invasible as they become less diverse and more stressed (Manchester
and Bullock 2000) and an increase in land in transition from agriculture may
have a similar effect. Global warming effects (Section 6.1) suggest that
natural systems will become more invasible. Thus, likely biotic trends will
increase pool size.
Transport – poor survival during transport has excluded many species from
introduction. Survival depends particularly on journey time and the containers
used. Growing air travel and the growing use of sealed containers will
therefore increase transport success. Deliberate, commercial movement of
pets and plants represents a form of protected transport which is increasing.
For instance, the introduction of the corn rootworm (Diabrotica virgifera
virgifera) from North America into Europe is likely to have occurred by
transport of fragile, adult beetles via direct military transport flights between
US and the Balkans in the 1990s. Subsequently the pest appears to be
moving around Europe, and into UK, on air flights, appearing near airports.
Until such rapid, long distance air transport was possible, it is unlikely that this
pest, whose larvae feed on the roots of maize, could ever have spread to and
within Europe so quickly.
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Source – As trade diversifies and involves more countries, Britain samples a
greater proportion of the global species pool. Biogeography, the non-uniform
distribution of species around the world, indicates that adding a new region
increases the pool more rapidly than simply adding that geographical area
alone would suggest. Further, every time a species moves to a new region, it
expands a new local species pool, thereby increasing its chance of being
sampled through trade into UK. For instance, our example of corn rootworm
above indicates that both North America and continental Europe are now
sources of introduction of this species into UK.
In addition to the relatively “passive” effect of sampling from a greater number
of sources and, inevitably, introducing new species, there is a more active
process operating today. Consumer interest in new and exotic foods, pets,
plants and new eco-touristic destinations will deliberately seek out and sample
a larger global species pool, as we see today in the focus of the horticultural
trade on delivering new exotic species to British gardeners.
The rate of introduction of new species in future, therefore, will be strongly
influence by whether or not global species pools are being depleted (i.e.
everything is getting everywhere) or opened up (ie new sources of new
species are appearing). Some data on patterns of introduction contains
enough geographical detail to shed light on this.
Figure 6.2 illustrates trends in sources of non-native plant species from some
of the data examined in Section 2.3.5. Remember that establishment follows
about a century after introduction, so this data reflects patterns of introduction
(trade and transport) from an earlier period, tempered by abiotic and biotic
factors affecting establishment from (usually) gardens into natural habitats.
For each source region, the number of establishments per decade are
illustrated. This data covers an enormous time span, and hence may not be
useful for looking at recent developments, but it does reveal significantly
different patterns in the rate of naturalisation from different source continents
(P < 0.001, r2 = 0.33; GAM curves fit by non-parametric smoothing function fit
by cubic B-splines). There is a shift in the relative importance of source
continents over time and the data clearly indicate that the rate of
establishment of species from Europe and the Americas is beginning to level
off or fall. It is possible that an asymptote for introduction is being reached for
these regions, as we might expect from a longer history of trade and transport,
and climatic similarity. On the other hand, establishments from other
temperate regions, such as Asia, are increasing.
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12
10
8
6
30
Asia
0
0
2
10
20
6
4
S. America
0
0
0
2
5
4
6
8
10
12
N. America
S. America
N. America
Europe
15
20
yr
10
Africa
2
0
1500 1600 1700 1800 1900 2000
yr
60
1500 1600 1700 1800 1900 2000
yr
40
Europe
Oceania
1500 1600 1700 1800 1900 2000
20
Species naturalisations per decade
Oceania
40
Asia
4
Africa
8
10
Doubtless other data sets could be analysed for trends in source countries.
Anecdotally, those familiar with non-native species introductions will identify
several strong, recent cross-taxon trends. In particular, the opening of the
former Soviet Union, China and Indochina to trade in the late 1900s has had a
dramatic impact on the movement of forest pests and diseases worldwide.
Timber and other trade from Russia and Asia has led to new introductions of
non-native forest insects into North America, Europe, and New Zealand, while
the rate of introduction of new forest pests into China has risen dramatically.
In another context, patterns of international development assistance have
fostered in the mid to late 1900s the movement of many agricultural pests an
diseases between southern continents, particularly from South American and
Asia into Africa, through introduction of new crop and animal varieties and
provision of food aid, contaminated with pests and weeds.
1500 1600 1700 1800 1900 2000
1500 1600 1700 1800 1900 2000
yr
yr
1500 1600 1700 1800 1900 2000
yr
Figure 6.2: Patterns of first wild record of naturalised non-native plant species in the UK and
Ireland. Curves are fitted non-parametric cubic B-splines (3 d.f.).
In conclusion, all factors influencing the rate of introduction of new species:
trade, transport, source, abiotic and biotic, appear to favour a future increase
in introductions and establishments. We therefore predict that the annual rate
of new species arrivals will increase in the next 20 years across taxa, and that
an increased volume of trade will not be a necessary condition (or necessarily
a good predictor) for this increase, but may accelerate it.
6.2.2 The effect of trade on UK agriculture and land use
The trend towards freer trade with respect to agricultural products within the
EU and the global economy will have a negative effect on domestic prices.
Price support schemes and import restrictions currently transfer market power
to domestic producers. Rather than facing a highly elastic demand curve,
domestic industries receiving significant amounts of protection are able to
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charge a price for their product above the world price. If the restrictions to
trade are removed, imported product will enter the market and force the
domestic price downwards towards the world price.
From a consumer perspective, the general level of prices can be expected to
fall, and the range of products available to rise. The degree to which this
could occur will be partially dependent on price differentials between domestic
and imported products. The prices of some goods may rise in the short term if
the ‘landed’ price of imported commodities is significantly higher than the
prevailing subsidised price in the domestic market, but this will not typify food
products across the board.
Little empirical evidence concerning world and domestic market price
differentials within EU countries is available. Estimates produced by the
Organisation for Economic Development (OECD) Consumer Support
Estimates over recent years suggest EU prices are significantly higher than
world prices, costing consumers between €50.6 billion and €62.8 billion per
annum (OECD, 2001). Dominating this calculation are milk products and beef
and veal industries, while the sugar industry also achieves a high percentage
CSE due to high EU prices (IEEP, 2002).
The effect on some producers will be severe, particularly where there are
large differences between the domestic and world price for food products.
If the restrictions on imported products are relaxed and downward pressure is
place on the domestic market price, then the capacity of domestic industries
to absorb the impact of increased competition and lower market prices will
depend on their production function and relative efficiency.
Around 64 per cent of the CAP budget was spent on plant industries in 2002,
while the livestock industries received around 23 per cent (CEC, 2003). But,
this does not mean that the impact of trade liberalisation policies will be
greatest for crops. While direct payments to farmers growing crops forms the
bulk of CAP support payments, livestock industries receive additional market
support in the form of tariffs, quotas and export refunds. Moreover, in
association with the global trend towards freer trade, on-going reform of the
Common Agricultural Policy (CAP) is likely to lead to the general level of
agricultural subsidisation and price support in Britain being further reduced
over time, with livestock industries being the main target for future reforms
Generally, this will mean that while all domestic agricultural industries will be
placed under increased competitive pressure from international suppliers,
livestock industries will be particularly affected. This is despite the relief
associated with the fact that industries that use plant products as inputs into
their production processes can expect costs to fall somewhat as market
liberalisation takes effect.
The decoupling proposals of the 2003 CAP reform are expected to weaken
incentives for all domestic agricultural producers to remain on the land, with
livestock producers being particularly affected by the new focus on increased
efficiency and competition. It follows that there may be a strong political
imperative both to provide a ‘welfare net’ to catch producers falling out of the
domestic industry, and to deal with the associated implications of land
abandonment. . This has clear implications for non-native species invasions.
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Land taken out of cultivation, and hence in transition, may be more invisible
and may also become a source of invasive species into cultivated land. To the
extent that livestock producers are those primarily affected, then the
implementation of the decoupling reforms could see upland regions as those
mainly affected by land abandonment unless policy steps are taken to limit
this. Upland areas may have particular environmental value, as watersheds,
tourist attractions and areas of unique biodiversity, catchments. Hence the
may have non-market value which could be affected by non-native species.
With the value (price and quantity) of domestic agricultural production
declining relative to imports, the economic impact of agriculture-affected
invasions on the UK economy will decline.
Besides this direct effect of liberalisation on lowering the value of agricultural
production and, hence, protection, there will be an additional effect associated
with lower domestic prices for consumers and its policy implications.
Specifically, Cook and Fraser (2002) have shown that, where the price of the
imported product is less than the locally-produced product, consumers’
interests will be enhanced by a lower level of protection from non-native pests.
This is because protection systems, such as quarantine, will tend to restrict
imports of cheaper goods, thereby raising domestic prices. Hence, where
imported food is cheaper, government will find it more difficult to defend
agricultural protection measures for local production systems, either through
prevention or eradication. Note also that the lack of impact of the FMD and
BSE crises on prices of meat, as a result of rapid import substitution, only
further affirms that consumers may be complacent about, or even opposed
to, protecting domestic producers in the face of a threat to local agricultural
production.
In summary, the combined effect of trade liberalisation and CAP reform is
expected to lower the value of both crop and livestock production in the UK
through reductions in both prices received and quantities produced by local
industries. Therefore, it is unlikely that the expected economic impact of an
invasion on any agricultural activity will be larger in the future than it is now,
and more likely that it will be smaller. This is despite the fact that trade
liberalisation and CAP reform may actually increase the probability of invasive
species entry and establishment. Moreover, bearing in mind the prediction
that livestock industries are likely to be the most affected, particularly in
association with the recent decoupling reform of the CAP, the effects of trade
liberalisation and CAP reform on land use are likely to take the form of some
land abandonment, mainly in marginal agricultural areas (e.g. uplands). And
although this land is where government support for environmental
conservation is likely to be strongest, agricultural land that has been
abandoned can be expected to be particularly vulnerable to invasion by nonnative weeds, and can provide undesirable reservoirs for pests and diseases.
Hence, impact of invasive species affecting environmental goods, including
biodiversity, desirable watershed properties and grazing as a land
management activity, could increase as a result of decoupling.
It follows that although trade liberalisation and CAP reform is likely in the next
20 years to reduce the economic impact of non-native species affecting
agriculture, it could also increase the (non-market) impact on environmental
goods which would be associated with the transfer of agricultural land to
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different uses. In this case, were government to make an investment in
prevention and management of non-native species problems on purely
(market) economic grounds, the likely increase in the (non-market) value of
investing in protecting environmental goods would be neglected.
Note however, that this conclusion is conditional – it will apply only if the
reduced economic impact on domestic agriculture arising from trade
liberalisation dominates any increase in expected damage caused by new
non-native agricultural pest problems. In particular, it is conceivable that some
threats will be so great as to lead actually to an increased, if transient,
commitment to agricultural protection. Export market effects, for example, can
tip the scales in this direction, as we have seen in our case studies, and
graphically in the case of the FMD epidemic. Alternatively, increased
agricultural protection could arise if new species have multiple effects, e.g. on
agriculture and health of humans or native species, as in the case of
introduced zoonotic or non-specific diseases, respectively.
6.3 Social issues
Future trends that affect how we regard our local environment are likely to
have substantial effects on how we view the nature and impact of non-native
species introductions. We have seen already that changes in trade and
markets may change the balance of agricultural and non-agricultural land, and
that consumer preference for cheaper food may indirectly favour the decline in
agricultural areas and greater risk of introduction of non-native pests and
diseases of agriculture. A future societal trend towards preferring local or
regional foods may work in the opposite direction, but we have suggested that
this trend, farmers markets and local labelling notwithstanding, may not be as
strong as that driving increased, cheaper imports.
If invasives in agricultural systems will be more “tolerated” in future, what
about invasives affecting environmental systems and goods, the growth of
which we have predicted in the last section? We have seen already (Section
3.4) that non-market values which we associate with environmental goods like
biodiversity are difficult to measure and are not well reflected in our model.
We have suggested (Section 5.2.3) that including these values into
projections of impact of non-native species harmful to the environment will
create a pattern of future impact very different to that of agricultural invasives.
Its time-lagged, exponential form reflects the slow spread and impact of many
environmental invasives, but also presumes that society may value
environmental goods more in future as incomes rise, and that society will
become more concerned about the loss of environmental goods as they
become more scarce, due to the impact of non-native species.
We will now explore how, in a horizon-scanning context, this presumed
continuing concern about non-native species affecting the environment will
change with possible social change, including changes in our sense of
national identity, in quality of life and the appreciation of the countryside.
To approach this subject, we have contracted the Institute for Environmental
Philosophy and Public Policy (IEPPP) at Lancaster University to execute a
literature review and workshop on how society views non-native species
currently, and how it might do so in future. The results of the study and the
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workshop are presented in Appendix 2. This study goes deeper into social
issues surrounding non-native species, and appropriate Defra policy, than we
do here, and we recommend that it be read separately to this study.
6.3.1 What is a non-native species?
This project has adopted the term “non-native species” as it has progressed,
having started using interchangeably such terms as alien, invasive alien,
naturalised, pest, weed, etc., all in the context of biosecurity. Like others, we
encountered a complexity and confusion of terminology and settled for the
most factual term. However, this complexity has far-reaching implications for
social perceptions. Put simply, society has no simple view of non-native
species and what may be undesirable about them.
The UK has a history of biological colonisation, extinction and re-colonisation,
associated with glacial periods and human movements that makes it
extremely difficult to define a clearly “native” fauna and flora. Existing scientific
approaches to presenting degrees of nativeness (e.g. archaeophytes,
neophytes) or threshold dates for such (e.g. post-glacial, Roman occupation,
1500) mean little to people today. Strong cultural associations with non-native
species (e.g. poppies, rabbits, sycamore) obscures this further, as does a
longstanding culture of deliberate and broadly benign introductions (e.g. of
garden plants, game species). From a society perspective terms like “alien” or
“non-native” have little biological or cultural meaning.
The public may be more responsive to descriptions which specify why a
species is undesirable, independent of its origin. Causing “outbreaks”,
displacing local species, causing diseases or other harm to local species and
human activities are more clear because they are properties of species, but
they are not necessarily associated with being non-native alien species.
Conflicts of interpretation, e.g. between a garden plant and an environmental
weed like Rhododendron, are resolved if these species are represented by
their properties in a particular context.
As an island nation, the UK and Europe have a relatively relaxed view about
biological colonisation, and a long history of accepting and incorporating new
species. A concept of dangerous alien-ness emerges more clearly in “settler
cultures” (e.g. North America, Australia, New Zealand, South Africa), where
there is a greater distinction between human activity, e.g. farming, towns,
cities, and unaffected, natural “wilderness”. The distinction between managed
and “wild” ecosystems is not sharp in Europe, in fact it would be difficult to
identify European “wilderness”, or a desirable European picture of harmonious
nature where human activity is absent. It is noteworthy that the current global
agenda for invasive alien species, e.g. in the Convention for Biological
Diversity, has been largely designed and drive by these “settler cultures”.
Finally, the concept of non-native or alien species as threats is potentially
offensive in our multi-cultural society. A xenophobic or even racialist
interpretation may be placed on words like biosecurity, alien, non-native, nonindigenous, and invasive (i.e. implying intention, as distinct from e.g.
“outbreak” which is purely descriptive), and on terms for non-native control like
“rhodo-bashing”.
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6.3.2 New species and societal change
Our society will likely grow more “cosmopolitan” and less “parochial”. The idea
that people should stay where they belong today is regarded as parochial,
while mobility, change, adaptation and adoption are promoted as desirable,
cosmopolitan properties. Some invasion specialists have argued that people
should be allowed to cosmopolitan but other species should not. However, as
people become more globally aware and comfortable with non-native species,
through cultural mixing, tourism, specialist interests (e.g. gardening, fishing,
exotic pets), non-native biological diversity, like cultural diversity, may be seen
increasingly as a good thing, providing it is benign.
A “preservationist” view of nature is traditional in UK, and underlies for
instance the original concept behind Sites of Special Scientific Interest
(SSSIs). As preservationists, we seek to preserve our natural history against
threat, for future generations. A “conservationist” perspective, more focused
on dynamic balance and sustainability of viable ecosystems, is growing in
popularity. In future, it is possible that a third, “evolutionary” or “adaptive”
perspective will emerge, which will focus on the capacity of species and
ecosystems to adapt to change. Such a perspective may celebrate not so
much the unique biological features of particular species, but their capacity to
adjust to and thrive in new environments, like we do as human beings. In such
a perspective, non-native species may be more acceptable and popular, as
“adapters” or “survivors”. The focus would shift then towards adapted
individuals or populations, which would also be in harmony with a growing
focus on individual rights (e.g. the rights of animals to life independent of
context).
Overall, therefore, the idea that there will be a growing public anxiety about
non-native species problems may be misplaced. Threatening new diseases
may bring many parts of the public together around a prevention or
eradication campaign, but threatening new garden escapes may not. Even
with diseases, society may view approaches which protect animals from
disease, e.g. vaccination, far preferable to eradication of disease by killing
animals. Hence having the diseases as a non-native species may no longer
be a social issue, only a trade issue.
While government is locked in a tradition of preventing and eradicating nonnative species in the name of agricultural protection, environmental
conservation and international commitments, the public perception may be
changing. For economic reasons discussed in the previous section, the small
part of the population advocating measures against non-native species may
decline in future as well.
These perspectives are all highly subjective. The evidence base regarding
public opinion on non-native species and the problems which they cause is
very limited in UK. Tools do exist for building such an evidence base, but they
have rarely been applied to non-native species issues (see Appendix 2).
6.3.3 Which way will the future go?
Integrating this sociological perspective with future trends relating to the
valuation of the environment and the use of the countryside, we come up with
two possible, contrasting answers to this question.
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On the one hand, a growing concern about the environment, conservation and
the welfare of species, domestic and wild, will lead to the public being even
more aware and concerned about the threat which non-native species pose to
the countryside. Growing counter-urbanisation, recreation in the countryside,
voluntary work and campaigning for conservation issues will probably
increase awareness of non-native species threats to the environment.
Regionalisation will further increase awareness and concern by promoting
local conservation issues. These factors would lead to the threat from
environmental invasives becoming an even greater public concern in 20
years. This, in turn would coincide with the visible emergence of
environmental problems caused by long-established non-native species, as
they slowly compound, such as the spread of weeds or the elimination of local
species, and further strengthen this trend. This pattern of societal change is
now emerging in countries like South Africa, USA and Australia, through
government and NGO campaigns and substantial growth in press coverage
(not unlike that emerging in UK, see Fig. 1).
On the other hand, society in 20 years may value the countryside and its
conservation and use equally highly, but not be so concerned about the
native-ness of the species which comprise it. A great proportion of tomorrow’s
society, like today’s, may remain unaware of non-native species, particularly
those invading natural habitats in the countryside. Those who are aware may
be confused by definitions or, in being more cosmopolitan that today, may be
more accepting of new species. People may even celebrate this enhanced
biodiversity. Even the most environmentally conscious citizens may have few
problems with ecosystems that incorporate new species and remain functional
and attractive.
It is difficult to draw a median trend out of these divergent scenarios. Indeed, it
seems very likely that we will see both scenarios played out with respect to
different threats. Overlaying this potential divergence of public opinion are
other factors that would create uncertainty in public reaction:

Society may react differently to different taxa independent of their
economic impact, as we discussed above, e.g. a new, large, noisy
vertebrate may draw a different response than a new weed or insect,
however invasive.

The sequential nature of non-native species problem may generate
broad swings in public perception simply as a result of timing and
clumping of events. UK has seen a recent string of animal disease
events: BSE, FMD, bovine TB and avian flu which may stimulate a
more prevention/eradication position amongst the public than had
these been more spread out.
While the overall trend in public opinion could go either way, we can be pretty
certain that it will be hard to predict future attitudes on the basis of present
information, and multiple scenarios will be required.
6.4 Conclusions
Pulling together these future drivers for non-native species impact, we come
to the following conclusions. These, of course, assume a “do no more than we
are doing now” strategy by government.
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
Likely climate change suggests that the rate of establishment and
spread of non-native species is likely to increase relative to rates which
we are using for current model predictions, with particular impact for
environmental invasives.

Likely trends in trade suggest that the rate of introduction of non-native
species is likely to increase relative to rates currently observed – new
sources will probably counter any effect of saturation, and an
increasing trade volume will only accelerate this.

Likely trade and market trends suggest that public concern about
protecting agriculture from invasion will, if anything, decline, while a
growing value and use of environmental goods may cause the opposite
trend there for environmental invasions.

Society in general does not share a concept or concern about nonnative species, and may become more or less tolerant of increased
environmental impact in future. Intolerance is likely to be focused
locally on campaigns against particular environmental invasives, not on
perceived national threats. Further, public opinion is likely to fluctuate
widely due to this combined uncertainty of invasion and uncertainty of
public response.
Thus, while the future will be characterised by more problems, the national
response to this may not be proportionate, is likely to focus on environmental
invasives, and will have a high degree of uncertainty and fluctuation. Given
that our government, like many at present, presumes that it is addressing
biosecurity threats in a context of growing public concern, and focuses its
efforts on tangible, agricultural threats, this future may be surprising.
One strong message from the social analysis above is that, in the absence of
a clear public concept of non-native species, and with poor current
awareness, government emerges not only as an interpreter of public opinion
but also as a driver. The way in which the government presents the non-native
species issue may be critical to whether and how much society values it in
future. To do this well, government must have tools to estimate real and
relative risk, and must be very careful about terminology and the social
context in which non-native species problems are viewed. In this context
Defra plays many roles – agricultural and environmental. The public may
entrust Defra with decisions on non-natives only if its position and strategy are
clear and well promoted.
6.5 A Quantitative Approach to Horizon Scanning
The ecological-economic model developed for predicting the impact of nonnative species introductions can be used for horizon scanning, by defining a
control case based on current parameter values, and comparing this to a
scenario where those parameters change in line with horizon scanning
predictions. In the model context, we are asking, over a 20 year period, if key
initial parameters were different, what would be the change in predicted
impact? By measuring the difference in the EDcrit between “control” and
scenario, we can measure quantitatively the degree to which biosecurity
significance is likely to change.
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Note that this involves running the model under different parameters over the
same time period, not changing parameter values gradually over time, as may
be more realistic (e.g. climate change may slowly change establishment rate
over 20 years, but we would model this as the impact of different
establishment rates over a full 20 year period). What we are looking for,
therefore, is a broad indication of the direction of change caused by changing
scenarios and its magnitude relative to other effects.
Such horizon scanning could be done with any organisms, e.g. any of the six
case studies we presented in Chapter 4. However, we chose to illustrate the
approach with a new example, Foot and Mouth Disease (FMD). Below, the
same reporting format is used as for the case studies presented earlier.
However, in addition to a control case, two different future scenarios are
presented. These relate to increased meat imports and trade liberalisation.
They are both hypothetical - more refined and specific scenarios are certainly
possible - but our objective here is to demonstrate the approach, rather than
to make specific, realistic predictions.
6.5.1 Description
Foot and Mouth Disease (FMD) is a highly contagious vesicular disease of all
cloven hoofed animals, causing vesicles in the mouth, on the teats, and on the
skin between and above the hoofs. It is most severe in cattle and pigs, and
while rarely fatal to adult animals (i.e. <5% of cases), production losses can
be significant as a result of weight loss. The acute phase of the disease lasts
8-15 days, after which recovery is gradual. Cattle are the worst affected, and
may be left with permanent scarring on their tongues and mouths that hinder
eating. In addition, feet deformities, mastitis and damaged heart muscles can
also lead to permanent weight loss (Wroot, 2001). Mortality rates of young
and weak animals can exceed 50 per cent. The virus is the sole member of
the aphthovirus genus family Picornavitidae, of which there are seven
serotypes (A, O, C, SAT 1, SAT 2, SAT 3 and Asia 1). These can be further
divided into around 60 subtypes on the basis of quantitative serological tests.
FMD is endemic in many countries of Africa, the Middle East, Asia, South
America and parts of Europe.
FMD outbreaks periodically occur in the EU, and have done so for centuries.
The Greek historian Aristotle, cultural and philosophical adviser to Alexander
the Great, wrote of a cattle plague in 350BC, which is now believed to have
been the first recorded outbreak of FMD. The Italian physician Fracastorius
provided the first detailed description of the virus in Venice in 1546. The first
report of FMD in Britain occurred in 1839, and the most recent in 2001. This
latest outbreak involved the Pan-Asian O strain of the virus, which first
appeared in India in 1990 as is now the most widely distributed of the seven
strains worldwide (Wroot, 2001). The disease forms a major constraint to
international trade, being a notifiable disease under the OIE Agreement (AHA,
1998).
6.5.2 Affected Industries in the United Kingdom
Table 6.4: Industries affected by Foot & Mouth Disease
Affected Industries*
Gross Value of
Production (5yr
Gross Value of
No. (5yr Avg.)**
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*
**
Avg.)**
Exports (5yr Avg.)**
Cattle and Calves
(Beef and Veal)
£1,410,800,000
£17,755,800
10,343,000
Pigs and Pig Meat
£971,900,000
£688,162,400
5,588,000
Sheep and Lambs
(Mutton and Lamb)
£691,800,000
£132,124,200
35,832,000
AHA (1998). 30
DEFRA (2002B).
6.6.3 Control Case
Assume that no eradication campaign is to be mounted against an FMD
outbreak in Britain in future. Instead, assume the arrival of the virus triggers a
system of widespread vaccination. As a result, the costs of production will
rise, a significant proportion of export markets for affected meat products will
be permanently lost, and livestock losses (particularly amongst young
animals) will increase. The existence of the EU, OIE and the International
Vaccine Bank is not assumed to hinder a vaccination campaign31.
Induced Changes in Average Total Cost of Production
Production cost increases will result because of the need for vaccinations
between 1 and 2 times per year, depending on the average immune response
of the animals and the level of challenge after vaccination. This is simulated
using a discrete distribution with possibilities of 1 or 2 vaccinations each with
an equal likelihood of occurrence (i.e. Discrete ({1,2}{1,1})). Coverage of at
least 85% is considered the minimal acceptable level (Wroot, 2001). Costings
are based on OIE manufactured cost of £0.30/hd to £0.50/hd per dose
(NOAH, 2003). As a founding member of the International Vaccine Bank,
supply shortages to Britain are not anticipated to persist over time, and costs
are likely to remain relatively constant in real terms. The average costs of
administering vaccinations are assumed to be £2.70/hd per vaccination (i.e.
approximately 2 minutes per animal).
Induced Changes in Average Total Revenue
Yield Loss
The impact of FMD is expected to be most severe on the cattle and pig
industries. Assume up to a 50% mortality of animals less than one year old,
and up to 5% mortality in mature animals. The disease is generally less
severe in sheep, and it is assumed that although no mature animals are lost
as a result of FMD, up to 50% loss in lambs will be experienced (AHA, 1998).
Taking into account total livestock numbers and age distribution of affected
animals, total yield losses can be expressed as Pert(10%,20%,30%).
30
Goats, deer and horses are also listed as hosts for the FMD virus, along with dogs, cats,
hedgehogs, rats and indeed humans (Wroot, 2001). These hosts are not considered in this
assessment.
31 Under ‘normal’ circumstances, these groups require a “stamping out” policy whereby
infected and possibly infected animals are slaughtered and animal movement restricted to
eliminate the virus. See OIE (2003).
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Export Revenue Loss Attributable to Loss of Pest Freedom Status
While the export losses resulting from a loss of Britain’s FMD status are likely
to be substantial, they will not be total due to the distribution of the virus
throughout the world32. To illustrate this point, consider beef exports from
Britain. Table 6.5 shows the principal destinations for these exports in the
1990s.
32
Much will depend on the strain concerned.
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A New Agenda for Biosecurity, August 2004
Table 6.5: British Beef Exports (Tonnes)
Country Receiving
1990
1995
France
67,000
80,000
Italy
4,000
42,000
Netherlands
9,000
17,000
Spain
1,000
7,000
Other EU
16,000
45,000
South Africa
3,000
27,000
Non EU
14,000
28,000
MLC (2000).
Of the destinations for UK beef exports, only South Africa experiences ongoing problems with FMD. Whilst being able to maintain pockets of FMDregional freedom, it is doubtful these would be maintained in the long run. If
so, the imposition of a ban on beef imports from Britain on the basis of FMD
endemicity would not be legal under the terms and conditions of the WTO. It
may therefore be possible to maintain sales to South Africa in the long term, a
market that represents approximately 11% of total beef export sales33.
Although total export bans are unlikely to persist, long term export market loss
is still expected to be large. It is specified as Pert(70%,85%,100%).
Biological Model Parameters
Many of the following parameters are of little significance due to the speed
with which FMD is expected to spread upon entry. Epidemic diseases are
difficult to parameterise using a model of this type, and some parameters
listed in table 2 are inconsequential due to the relatively high cost of export
bans which ensure in the first 12 months of introduction and persist over tie.
That is, regardless of the number of animal infections, the infection rate
amongst herds or the total animal population, the costs of export losses
accrue from the time of an initial report. At best, area quarantine strategies
and animal movement restrictions within Britain are expected to slow the rate
33
The distribution of such a contagious pathogen is likely to increase over time, placing more
trading regions in a similar situation. In terms of animal health, this is not the most desirable
of situations. On the other hand, from an animal welfare point of view, the preference for
vaccinations over large scale culling would certainly be welcome. Moreover, if the spread of
diseases like FMD leads to a decline in live animal transport, the welfare of the animals will
almost certainly rise. These moral and ethical concerns seldom enter into scientific or
economic debate since they are largely negative externalities created by the live animal trade
for which the industry bears no cost, and for which society (as the primary ‘consumers’ of
animal welfare) pays no formal abatement costs. In the modern environment of global trade
and international treaties, this is a most unfortunate situation, and one that needs to be
addressed.
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A New Agenda for Biosecurity, August 2004
of naturalisation by 1 to 2 years.
characterising the control scenario.
Table 6.6 lists the parameter values
Table 6.6: Parameterisation – Control Case
Parameter
Assumed Parameter Value
P(Entry)  P(Establishment)
PertAlt(5%, 0.00091, "m. likely", 0.0077, 95%, 0.024)
(VLA, 2003)
Amin (hd)
Pert(100,200,300)
Amax (hd)
53,982,000 (DEFRA, 2002B)
R
Pert(15.0,17.5,20.0)
Nmin
Pert(1,2,3)
K (Nmax)
Pert(1.0M,5.5M,10.0M)
Smax
Pert(70,85,100)

Pert(0.10,0.15,0.20)
D
Pert(800,1000,1200)
Non-Market Effects of Naturalisation
Human Health
Reports of humans affected by direct FMD infection are extremely rare. Up to
1994, only 40 such cases were known to have occurred worldwide. The
probable mode of transmission in most of these cases is thought to have been
drinking contaminated milk (Wroot, 2001). Given the rarity of human health
complaints attributable to FMD they are ignored in this analysis. This is
consistent with VLA (2003).
Environmental
Susceptible indigenous deer populations may be affected by FMD
naturalisation, including Red Deer (Cervus elaphus) and Roe Deer (Capreolus
capreolus). While it is doubtful these populations are threatened with
extinction by FMD, mortality rates among young animals are expected to rise.
In addition, infected individuals may experience considerable distress due to
legions in the mouth and on the feet, causing weight loss and heart strain.
This is possibly overstating the effects of FMD since the 2001 outbreak
appeared to have a negligible impact on wild populations.
C. capreolus are native in Scotland, but became extinct in England during the
18th century but successfully reintroduced in the 19th century. Their current
population is estimated at around 500,000. This includes approximately
150,000 animals in England, 350,000 in Scotland and around 50 in Wales
(Harris et al., 1995).
Cervus elaphus from native stocks are only confirmed in parts of Scotland and
north-west England (Lowe & Gardiner 1974). The current total population is
estimated at around 360,000, including 12,500 animals in England, 347,000 in
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A New Agenda for Biosecurity, August 2004
Scotland and fewer than 50 in Wales. In addition, there are a further 7500 in
parks and 52,125 on farms (Harris et al., 1995).
No appropriate studies have been found to elicit existence values for these
deer populations. However, Loomis et al. (1989) use a travel cost method to
gauge the willingness to pay of both hunters and non-hunters to see dear on
hunting and/or viewing expeditions in California. The Geographic differences
alone make benefit transfer a dubious business indeed. However, in the
absence of more appropriate studies it will suffice for a broad estimate of deer
values. The values extracted from the study indicate that members of the
public interested only in dear viewing were willing to pay US$0.40 per trip, on
which an average of 6 deer were seen. Hence, an approximate average
value per head may be in the order of US$0.07/hd, or £0.04/hd (EVRI,
2003)34.
Assuming the effects of FMD on the native deer population in Britain are of
the same order of magnitude as farmed animals, numbers may fall by
between 7% and 10% per year. Ignoring the population dynamics of deer
populations, this equates to a £2,840 to £9,470 environmental impact per
year.
Deleterious effects on non-native animal populations offset the negative
impact on native deer populations. Susceptible species include the following,
the details of which are taken from White and Harris (2002):
Sika deer (Cervus Nippon) – introduced to Britain in the 1860s, current
population of approximately 11,500;
Fallow Deer (Dama dama) – introduced during Roman/Norman occupation,
current population of 100,000;
Reeves’ Muntjac (Muntiacus reevesi) – introduced in the early 1900s, current
population approximately 40,000;
Chinese Water Deer (Hydropotes inermis) – introduced in 1915, current
population 40,000;
Pere David’s Deer (Elaphurus davidianus) – introduced in 1963, current
population a mere 30 animals.
Feral Goat (Capra hircus) – introduced in the Neolithic age, current population
estimated at around 3,565
Feral Sheep (Ovis aries) – introduced in the Neolithic age, current population
estimated at approximately 2,100
Feral Pig (Sus scrofa) – introduced during the 1800s, current population only
200 animals.
34
Since this estimate relies on travel cost, it has been derived from those prepared to travel to
see deer in the wild. It therefore excludes those whom are not prepared to travel, but who
gain utility from the fact that the deer population is sufficiently large and in a general state of
health. Moreover, average values are a poor substitute for marginal values since they take
no account of relative scarcity.
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A New Agenda for Biosecurity, August 2004
Socio-Economic
Typically, the socio-economic effects of an FMD outbreak, specifically the
2001 outbreak in Britain, have been estimated using the impact on the tourism
sector. In an eradication campaign, the countryside must effectively be closed
down, including many roads and footpaths. As a result, the tourist dollar is
effectively redirected to other substitute goods and services (e.g. international
travel, urban gymnasiums, etc.). This leaves businesses in the country faced
with greatly diminished revenue, and ‘consumers’ of the countryside with
depleted utility by being forced to undertake less preferable leisure pursuits.
VEERU (2003) estimate the total cost to the tourism industry from the 2001
FMD outbreak to have been around £595 million. This comprises of £170m
lost from domestic tourists who chose to travel overseas rather than in the
British countryside, and £425m value added forfeited from international
visitors forced to go elsewhere due to movement restrictions. The majority of
these costs cannot be said to be attributable to the virus itself, rather to the
response invoked by Government and industry. In a naturalisation (as
opposed to eradication) scenario, lost tourism income is likely to be less,
although to what extent it is not clear. Movement restrictions will still apply in
non-affected regions, but once vaccination becomes standard practice this will
no longer apply. Blanket restrictions are unlikely.
6.5.3.1 Results
X <=£2.44 B illio n
95%
X <=£0.00
1.0 5%
Mean = £1.03 Billion
Probability
0.8
0.6
0.4
0.2
Figure 6.3: Cumulative distribution of the critical level of Expected Damage (ED crit) over 20
years for the control case – Foot & Mouth Disease
0.0
0.0 is consistent
0.5
1.0 that reported
1.5
3.0
This result
with
in2.0Harvey2.5
(2001), where
the3.5annual
benefits of exclusion are estimated
at £1.2
billion.
Values
in £ Billions
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A New Agenda for Biosecurity, August 2004
No. Animals Affected ('000)
60,000
95%
50,000
40,000
30,000
Mean
20,000
10,000
5%
0
10
20
30
Figure 6.4: Incidence/Time and variability – Foot & Mouth Disease
Year
£3,000,000
£2,500,000
95%
£'000
£2,000,000
£1,500,000
Mean
£1,000,000
£500,000
5%
£0
10
20
Figure 6.5: Expected Invasion Impact (EI)/Time – Year
Foot & Mouth Disease
30
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A New Agenda for Biosecurity, August 2004
Sensitivity Analysis
Table 6.7: Sensitivity Analysis – Foot & Mouth Disease
Parameter
P(Entry)  P(Establishment)
Average Total Cost –
Cost per Vaccination
Average Total Revenue Loss
– Yield Loss
Average Total Revenue Loss
– Export Losses
Animals Affected Upon
Introduction (Amin)
Maximum Number of
Affected Animals (Amax)
Intrinsic Rate of Spread (r)
Pest Density Immediately
Upon Introduction (Nmin)
Maximum Attainable Pest
Density (K)
Maximum Number of Satellite
Infestations (Smax)
Intrinsic Rate of Satellite
Generation ()
Infection Diffusion Coefficient
(D)
*
Change in Parameter Value
(%)
Resultant Change in
Expected Damage (%)
- 50.0
- 48.8
+ 50.0
+ 38.7
- 50.0
- 4.3
+ 50.0
+ 3.8
- 50.0
- 20.7
+ 50.0
+ 17.9
- 50.0
- 30.2
+ 50.0*
+ 15.0
- 50.0
- 4.3
+ 50.0
+ 0.8
- 50.0
- 5.1
+ 50.0*
+ 4.3
- 50.0
- 5.6
+ 50.0
+ 1.9
- 50.0
- 2.1
+ 50.0
+ 0.2
- 50.0
- 0.8
+ 50.0
+ 0.9
- 50.0
- 2.2
+ 50.0
+ 1.1
- 50.0
- 1.5
+ 50.0
+ 1.8
- 50.0
- 2.1
+ 50.0
+ 2.6
Sensitivity test value beyond a maximum attainable value, and is therefore purely for illustration.
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A New Agenda for Biosecurity, August 2004
6.5.4 A Trade Change Scenario
Consider a situation where imports of meat with a risk of FMD contamination
were to increase, due perhaps to a new source of meat products from an FMD
affected country, say in Asia or Africa, or the spread of the disease into an
existing source area. Then, in the absence of any UK government response to
this new threat, the probability of introduction and establishment would
increase. Below we model an approximately 5-10 fold increase in probability.
What effect would this have on predicted losses?
Model Parameters
Table 6.8 lists the parameter values characterising the control scenario.
Table 6.8: Parameterisation – Climate Change Scenario (Foot & Mouth Disease). The arrow
indicates parameters which are increased in the scenario.
Parameter
Assumed Parameter Value
Change From
Control Scenario
Vaccine
£0.30-£0.50/hd/dose
-
No. Doses
Discrete(1,2)
-
Yield Loss
Pert(10%,20%,30%)
-
Export Revenue
Pert(70%,85%,100%)
-
P(Entry)  P(Establishment)
Uniform(0.001,0.05)

Amin (hd)
Pert(100,200,300)
-
Amax (hd)
53,982,000 (DEFRA, 2002)
-
r
Pert(15.0,17.5,20.0)
-
Nmin
Pert(1,2,3)
-
K
Pert(1.0M,5.5M,10.0M)
-
Smax
Pert(70,85,100)
-

Pert(0.10,0.15,0.20)
-
D
Pert(800,1000,1200)
-
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A New Agenda for Biosecurity, August 2004
Results
1.0
X <=£0.00
5%
X <=£2.57 B illio n
95%
Mean = £1.13 Billion
Probability
0.8
0.6
0.4
0.2
0.0
0.0
0.5
1.0
1.5
2.0
2.5
3.0
3.5
Values in £ Billions
Figure 6.6: Cumulative distribution of the critical level of Expected Damage (EDcrit) over 20
years in the trade change scenario – Foot & Mouth Disease
X <=-£1.91B illio n
5%
1.0
X <=£2.04 B illio n
95%
Mean = £0.11 Billion
Probability
0.8
0.6
0.4
0.2
0.0
-4.0
-3.0
-2.0
-1.0
0.0
1.0
2.0
3.0
4.0
Values in £ Billions
Figure 6.7: Cumulative distribution of the critical level of Expected Damage (ED crit) differential
between the control case and the trade change scenario over 20 years – Foot &
Mouth Disease
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A New Agenda for Biosecurity, August 2004
Not surprisingly, a trade factor which simply increases the probability of entry
and establishment will, over 20 years, make establishment more likely earlier
and hence increase impact, in this case by about 10%.
6.5.5 A CAP Reform Scenario
Imagine, as is likely for European agricultural policy, the general level of
agricultural subsidisation for the sheep and cattle industries declines at some
future date. Since this reduction is indexed to the rate of subsidy reduction
across the EU, the number of domestic producers leaving the market is lower
than many anticipated. Nevertheless, the collective industries susceptible to
FMD infection are two thirds of its size in 2003, as reflected in Amax in Table
6.9 below.
Increased imports have increased the likelihood of FMD
introduction and establishment to a level similar to the last scenario. This rise
is attributable to contaminated meat products being fed to susceptible
domestic livestock, specifically pigs.
Model Parameters
Table 6.9 lists the parameter values characterising the control scenario.
Those with arrows have been changed in the direction indicated, relative to
control.
Table 6.9: Parameterisation – Trade Liberalisation Scenario (FMD)
Parameter
Assumed Parameter Value
Change From
Control Scenario
Vaccine
£0.30-£0.50/hd/dose
-
No. Doses
Discrete ({1,2}{1,1})
-
Yield Loss
Pert(10%,20%,30%)
-
Export Revenue
Pert(70%,85%,100%)
-
P(Entry)  P(Establishment)
Uniform(0.001,0.05)

Amin (hd)
Pert(100,200,300)
-
Amax (hd)
35,988,000

R
Pert(15.0,17.5,20.0)
-
Nmin
Pert(1,2,3)
-
K
Pert(1.0M,5.5M,10.0M)
-
Smax
Pert(70,85,100)
-

Pert(0.10,0.15,0.20)
-
D
Pert(800,1000,1200)
-
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A New Agenda for Biosecurity, August 2004
Results
X <=£0
X <=£2,16 B illio n
95%
1.0 5%
Mean = £0.98 Billion
Probability
0.8
0.6
0.4
0.2
0.0
0.0
0.5
1.0
1.5
2.0
2.5
3.0
Values in £ Billions
Figure 6.8: Cumulative distribution of the critical level of Expected Damage (ED crit) over 20
years in the CAP reform scenario – Foot & Mouth Disease
X <=-£1.92 B illio n
5%
1.0
X <=£1.79 B illio n
95%
Mean = -£0.05 Billion
Probability
0.8
0.6
0.4
0.2
0.0
-4.0
-3.0
-2.0
-1.0
0.0
1.0
2.0
3.0
Values in £ Billions
Figure 6.9: Cumulative distribution of the critical level of Expected Damage (ED crit) differential
between the control case and the CAP reform scenario over 20 years – Foot &
Mouth Disease
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A New Agenda for Biosecurity, August 2004
The overall effect of this policy change is multi-factored and more complex
than simply a change in imports, and leads to a 5% decrease in impact
relevant to the control scenario.
6.5.6 Conclusion
The exclusion of FMD from Britain saves the economy around £1 billion per
year, which makes it a virus of considerable significance. While the virus is
excluded, livestock owners are spared the expense of vaccinating animals
against the disease and the losses associated with being unable to access
major export markets for meat and dairy produce. If future changes in trade
increase the probability of establishment, the predicted impact will increase
relative to the control case. Hence the predicted future cost to the UK
economy of FMD will be higher.
A change in CAP subsidy may also increase trade, by making imported meat
more competitive, creating the same problem. But now, this change is
countered by the falling area of production due to reduced UK
competitiveness. The net effect could be in fact a decrease in the UK impact
of FMD, relative to the scenario where production continues as at present.
Hence, to consider UK trade liberalisation as increasing risk and impact due to
increased imports is simplistic. Further, farmers facing more intense
competition and poorer margins will be doubly disadvantaged by increased
risks of FMD, but the importance to the national economy will be lower,
precisely because there is less production to be affected. In the absence of
government intervention, it is easy to see how a downward spiral in production
could be induced by the combined effects of falling prices and increasing
biosecurity risks.
For policy makers thinking about prevention or eradication, this difference may
be important. In the case of FMD, losses are still very substantial and will
certainly justify intervention. But in a commodity of less value, a government
may be challenged under trade liberalisation to justify the cost of
prevention/eradication of biosecurity risks
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A New Agenda for Biosecurity, August 2004
Chapter 7 – Prevention and Eradication of Non-native
Species Threats
As well as being aware of the level of risk presented by potentially invasive
organisms, biosecurity policy-makers need to plan how to react to potential
invasions before the event. The pursuit of a ‘zero risk’ biosecurity policy is not
appropriate. The sheer abundance and diversity of invasive species makes
zero risk a technical impossibility. With this in mind, it is imperative that
biosecurity authorities maintain a reactive capacity to respond rapidly and
decisively to new species invasions, and that any response follows a
predetermined plan to maximise the net benefits from scarce resources 35.
Response efforts must be subjected to economic evaluation to ensure
resources are being put to efficient and desirable use. If not, situations may
arise where great expense is incurred in managing an outbreak of an invasive
species of relatively minor significance whilst more serious outbreaks do not
receive the attention they deserve.
A prevention strategy reduces the probability of entry and/or establishment of
a non-native species.
A management strategy is a generic term
encompassing any form of action taken by public institutions to minimise,
reduce or eliminate an invasive species and/or the economic damage inflicted
by that species once it has invaded36. Hence, eradication, partial control, and
even a “do nothing” approach all represent different management strategies
that may be adopted in response to an outbreak.
In this chapter, we contrast particularly eradication and prevention strategies.
Most government policy today focuses on one of these two alternatives. Most
studies of non-native species problems assert that “prevention is better than
cure” (SoA, 2004). Where prevention is not achieved, most governments see
their role as removing or eradicating introduced threats before the combined
costs of impact and management grow to damage the economy. However,
eradication is not always possible, and problems may become chronic, often
requiring continuous management. At this point, the responsibility usually
shifts from government to the private sector. Hence, in the models in Chapter
3, which explore the “government does nothing” approach, the cost of pest
control or animal vaccination against a non-native species falls on the
producer. Only rarely would a government commit itself to paying the cost of
continued management of an established non-native species problem, for
instance where that species invades national parks or other governmentmanaged lands. This chapter demonstrates how economic analyses can play
an important role in identifying the benefits and costs of these different
responses, and can distinguish between desirable and non-desirable policy
options.
The politics of ‘managed risk’, as opposed to ‘zero risk’, are somewhat delicate. After all, a
political representative will be more comfortable with a policy statement that risk of invasive
species incursions is to be minimised than with a statement that a certain amount of annual
economic loss is expected by the government, or is regarded as acceptable (Gascoine,
2001).
36 Note that economic analyses should take place before an actual invasion takes place so a
decision on what management strategy to adopt has already been reached. This way
reaction time is minimised.
35
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A New Agenda for Biosecurity, August 2004
7.1 Prevention and eradication strategies – an overview
In all cases the higher the net benefit generated over time, the more desirable
the overall responsive strategy will be. Wherever a management option
involves the removal of an invasive from affected regions, it will generate a
Total Benefit (TB) stream over time. The benefits produced will depend on
the extent and success of the removal process, which in turn will depend on
the extent and location of the outbreak upon commencement of the response.
The provision of management effort (E) for a given strategy will generate a
Total Cost (TC) stream over time. Hence, by examining the flow of net
benefits over time economists can identify the most appropriate form of
response.
The analytical framework presented below is static. It only deals with one
time period, which would be the life of a specific prevention or management
strategy. By comparing the costs and benefits of that strategy over its lifetime
a net benefit stream can be used as a measure of desirability. For instance,
policy-makers may be interested in prevention technologies to lower the
probability of an invasive species entering a region. These technologies will
have set-up and running costs over time, but they will produce notional
benefits over the time horizon being considered (be it one, five, ten or twenty
years, or more). An alternative to investing in prevention may be to simply
eradicate outbreaks as they occur over that same time period. The costs of
doing so largely depend on the time until detection, and the benefits may be
high but relatively short lived. The inclusive period model, in which the actual
time period may run over many years, can be used to help make a decision
on which of these alternative strategies is optimal in terms of the delivery of
net benefits to society.
In the following sections a framework is presented that can be used in net
benefit estimation for alternative prevention and eradication strategies. While
a dynamic modelling approach may give more detail and power to our
understanding of the interplay of a management measure and the population
growth and spread of an introduced species, its output is highly specific to
particular species problems. Our static modelling approach, by contrast, takes
a broad, comparative view of the outcome of two principle strategies for
government, projected into the future, in order to help decision makers
conceptualise the general problem of invasive species management.
7.2 Eradication
Before discussing the relative merits of an eradication strategy, a brief
introduction to the Total Cost and Total Benefit functions associated with any
management activity is needed. These two functions allow us to estimate the
flow of net benefits over time, and therefore the economic merit of different
management options.
Let us assume that the life of an eradication strategy (i.e. the period over
which all incursions will be eradicated and post-eradication measures may be
adopted) is known with certainty, as are the costs and benefits it will involve.
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A New Agenda for Biosecurity, August 2004
In this conceptual discussion, the Total Benefits generated by removing a pest
are a function, , of management effort per year, E37. i.e.
TB   ( E )
(16).
Management effort, like any other input into a production function, is subject to
diminishing returns. That is, there is a limit to the overall benefit that may be
achieved and as this is approached successive increases in management
effort will yield successively smaller increments in total benefits from reducing
prevalence. This is true of any invasive species up to the point where the
population is eradicated. If enough management effort is expended, the
complete removal of the introduced organism may be achieved, at which point
marginal benefits should be at their lowest point. However, there is one
exception. The complete removal of an invasive species that has reduced
export market access, like Newcastle disease or FMD, produces a sharp
increase in total benefits at the point of eradication. Up until eradication is
fully achieved and demonstrated, management effort produces declining
marginal benefits from successive increases in management effort. Once
eradication is confirmed we would expect a spike in Total Benefit due to the
sudden export revenue generated by achieving area-freedom from the
invasive. Beyond the point of eradication there are no additional benefits to
be had by further management effort38 other than new prevention activity.
The benefits accruing to the economy as a result of annual management
effort, as described above, are shown in Figure 7.1. Here total benefits are
depicted as a function of management effort. Beginning at zero and moving
left to right along the horizontal axis, the returns to management effort
increase at a decreasing rate until the point Eerad is reached, and eradication
achieved.
As there are fewer and fewer affected areas/livestock as
management effort increases, costs associated with chemical treatments,
vaccination, yield losses and the like gradually decline39. However, export
losses that result from the initial loss of pest area freedom remain until
eradication is achieved. This requires a very high level of management effort,
Eerad, but once reached the total (and marginal) benefits suddenly increase to
TBmax1. However, this increase does not persist. At levels of management
effort beyond Eerad the returns to investment will be zero since the invasive is
no longer present.
37
The complexities of management over multiple time periods are dealt with in section 3.5.
Conceivably, management effort expended above and beyond the amount necessary for
the eradication of one outbreak may actually prevent subsequent outbreaks.
39 These can be thought of as variable benefits of management effort.
38
136
Total Benefit of Management Effort (TB)
A New Agenda for Biosecurity, August 2004
TBmax1
TB
TBmax2
0
E erad
Management Effort
Figure 7.1: Benefits of management effort – e.g. an export-limiting disease
In contrast, the total benefit curve for a crop pest like the Colorado Potato
Beetle (Leptinotarsa decemlineata), where there is no export benefit to be
gained from eradication, is maximised at TBmax2. TBmax1 and TBmax2 represent
extreme situations. Some invasive pests may cause some industries limited
export losses due to the need for costly pre-export requirements for certain
markets or a complete prohibition from only some markets. These will usually
be localised where plant industries are concerned, although exceptions could
certainly occur. For example, the stigma surrounding the extensively
researched wheat disease Karnal Bunt may cause the market response to an
invasion to be severe, resembling an animal disease40.
In terms of cost, the response or management strategy that takes place
following an invasion (no matter what it involves) requires an investment of
effort, E. Assume that the costs associated with each unit of E include an
opportunity cost of labour (or the wage rate, w)41. There are also likely to be
variable capital costs (c) involved (such as tools, equipment, chemicals,
vehicles, etc.), which will largely be determined by the infested area, and the
population abundance and density at each point in time. But, assume here
that the costs are directly proportional to the effort expended on management
activities. The Total Cost (TC) of a particular management strategy is given
by
i.e.
TC  (c  w).t
(17).
Where;
c = variable capital costs of management activities;
w= opportunity cost of time spent on management effort.
40
See for example Thorne et al. (2004), Brennan. (1992), Stansbury and Pretorius (2001).
The term opportunity cost here refers to the income an economic agent must forgo in order
to participate in invasive species management activities. The wage rate, w, is used as a
simple proxy for this opportunity cost.
41
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A New Agenda for Biosecurity, August 2004
There may also be a fixed cost component to consider in the TC function. In
practice, the TC function for specific management options can be specified in
detail. Management effort will almost certainly exhibit decreasing returns, as
described earlier. But, here assume a simple linear function can be used as a
generalised representation of management costs.
7.2.1 Eradication, net benefit maximisation and EDcrit
It may be that the benefits accruing to affected parties in the economy from
removing the entire population/area of an invasive species are more than
sufficient to offset the costs of an eradication strategy. Figure 3.3 depicts
such a situation. Here, the TB curve lies everywhere above the TC curve in
the diagram. No matter what the extent of management effort required to
remove the invasive organism, the benefits of doing so will always outweigh
costs.
The point at which net benefits are maximised corresponds to the level of
effort that produces the greatest total benefit relative to total cost. In the
situation depicted in Figure 7.2, this involves investing Eerad, and subsequently
removing the entire population of the introduced species.
TC,TB
TB
EDcrit
TC
0
E erad
E crit
Management Effort
Figure 7.2: Eradication of an incursion
The diagram contains a ceiling for total costs, labelled EDcrit. This line
indicates the expected total damage the invasive has been estimated to inflict
over the specified time period using the techniques applied to the case studies
in Chapter 4. Recall from this chapter that EDcrit is effectively a representation
of one point along the EI vs time function related to the organism in question.
In Chapter 4 cumulative distributions of EDcrit over 20 years were used for
each case study. Here, we use the mean EDcrit value for that 20 year period.
In other words, we are asking “will the cost of eradication achieve net benefits
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and remain below the expected total cost of doing nothing over that same
period?” As Chapter 5 postulates, expected impact of harmful non-native
species varies over time in one of three ways. It either increases at a
constant rate, a decreasing rate or an increasing rate over time, depending on
the nature of the invasive species. Since EDcrit is a point estimate of invasion
impact over the life of a management strategy, it too will vary according to the
nature of the non-native species problem and the number of years it
represents.
A total cost ceiling like EDcrit can be used as a decision rule in the static model
to indicate where the inter-temporal benefits and costs of management are
expected to break even. As long as the costs involved in management effort
invested in controlling an outbreak are below this critical level, these costs will
be below the expected benefits of exclusion. EDcrit is a measure based on an
expected value within a potentially wide probability distribution, and should
therefore only be used as a guide to the break-even level of management
cost. In some cases, we may wish to explore the cost and benefits of an
eradication strategy relative to a value other than the average risk of a “do
nothing” approach (for instance, some percentile that covers a greater portion
of the risk distribution, even up to the maximum loss), thereby using other
features of the impact distributions in the case studies of Chapter 2. As
Figure 7.2 has been constructed, the optimal response involves an investment
in management effort of Eerad, which lies to the left of the critical level of effort,
Ecrit. Levels of effort to the right of Ecrit will lead to a net social loss regardless
of how effectively the outbreak is controlled since the expected benefits
produced over the life of the management strategy are insufficient to offset
costs. A net gain can be expected to accrue to the economy from eradication
as long as Eerad lies to the left of Ecrit42.
The inclusion of the value Ecrit provides an important link between the pest
prioritisation exercise (i.e. Chapter 5) and the static framework used here to
examine the net benefits of response activity. As mentioned above, EDcrit can
be constructed as needed, with respect to different time horizons or different
criteria of risk minimisation and levels of precaution desired.
7.2.2 Multiple net benefit maximisation options
The choice of whether to eradicate or control an outbreak will not always be
as straightforward as in Figure 7.2. Consider once more an invasive species
with large export market access implications, as depicted in Figure 3.1. If we
now introduce a TC curve to this diagram, the situation becomes more
complicated. Policy-makers faced with the scenario shown in Figure 3.4 can
maximise the returns to management effort in one of two ways. They can
either control the outbreak with an investment of E* effort which is insufficient
to remove the species, or they can embark on the more costly option of
expending Eerad. The problem for policy-makers is that Eerad represents a
much larger investment than E*.
42
When analysing the eradication option, it must be recognised that the probability of success
will almost certainly be less than one. There is a need to incorporate risk into the TB and TC
curves. Stochastic models are therefore needed in the analysis of actual eradication
proposals.
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TC,TB
A New Agenda for Biosecurity, August 2004
NBerad
TB
EDcrit
TC
NBcontrol
0
E*
E erad E crit
Management Effort
Figure 7.3: Eradication or strategic management?
So, although efficiency arguments are important, there is also an issue of
scale to be considered when making resource allocation decisions43. This is
where the critical level of expected damage becomes, as the name suggests,
critical. Although a greater amount of net benefit can be generated by
pursuing the eradication option as opposed to the control option, Eerad lies
very close to Ecrit.. Despite this, a higher net benefit is still expected to be
generated over time, and eradication is likely to be the preferred option of
policy-makers. It should be noted that Ecrit incorporates the risk of further
outbreaks over the specified time horizon, and Eerad the cost of eradication of
all outbreaks over that period. By contrast, the control option is complex, and
not easily interpreted in a static model – a partially controlled population will
increase again, in a manner dependent on the level of control at one time, i.e.
the size of the population left to recover. This would require a dynamic
modelling approach to represent accurately.
7.3 Prevention
In the context of our modelling approach, prevention of non-native species
introduction involves reducing the risk of entry and establishment of new
species. The sensitivity analyses in the case studies of Chapter 4 revealed
that the probabilities of entry and establishment for many invasive species are
highly sensitive variables in determining biosecurity significance. As such,
any measures that can be used to affect these parameters may be very
successful risk management techniques. Effective biosecurity strategies for
many non-native species may entail inspection and interception at points of
43
This is particularly true where the opportunity costs of scare biosecurity resources are
large.
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A New Agenda for Biosecurity, August 2004
entry, such as docks and airports. Or it may involve pushing the reach of
domestic biosecurity services well beyond the regional boundary being
protected. In the case of the UK, this may involve the use of sanitary and
phyto-sanitary market entry requirement on known pest pathways, quality
assurance and livestock/produce certification schemes. The expected effect
these will have on the biosecurity significance of the pests they target is to
reduce the probability of entry over a time period, and hence the probabilistic
expected impact over that period.
Expected Invasion Impact
Clearly, any reduction in the probabilities of entry and/or establishment will
shift the EI curve for the pest concerned downwards. This is illustrated in
Figure 7.4, which depicts an invasive with a high initial impact (characteristic
of an OIE disease). The shaded area below the EI 0 curve and above the EI1
curve therefore represents the expected net benefits from a pre-invasion risk
management strategy.
EI0
EI1
0
10
20
30
Year
Figure 7.4: Pre-invasion biosecurity measures
When viewed in this form, the role of benefit and cost analysis in pre-invasion
risk management strategies becomes clearer. Before investing in new
technologies to reduce the probability of importing a pest with a product, or
detecting contaminated produce entering the UK, biosecurity authorities must
weigh up the expected gains from utilising the technology and the costs
involved in developing and installing it.
If these development and
implementation costs outweigh the anticipated benefits (the area between the
EI0 and EI1 curves in Figure 7.4), a benefit cost analysis will reveal a negative
benefit to cost ratio. On the other hand, if costs are much less than expected
benefits, then the new technology would repay investment.
While the benefits of prevention strategies need to be compared to their costs,
it is also important for policy makers to compare prevention vs. eradication. In
stark contrast to the situation depicted in Figure 7.2, an invasive species
outbreak may have no economic solution through eradication, as shown in
Figure 7.5. Here, the TB curve can be seen to lie everywhere below the TC
curve. No matter what the investment in management effort required to
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A New Agenda for Biosecurity, August 2004
TC,TB
contain or eradicate outbreaks over the management period, the costs to the
economy will always outweigh the benefits. Even investing Ecrit produces a
net loss of NL.
TC
TB
EDcrit
NL
0
E crit
E erad
Management Effort
Figure 7.5: No solution through eradication
This re-enforces the need for ex ante analyses to estimate the TB and TC
functions related to prevention and eradication for individual species. These
analyses will need to consider a variety of points along the expected invasion
impact curve (or EI curve, recalling Chapter 4) for each species to allow for
time lags between arrival and detection.
7.4 Evaluating prevention vs. eradication policy options
The framework described above can provide policy-makers with information to
help identify optimal biosecurity allocation for particular kinds of non-native
species. To do this, we need to know the timeframe over which the resource
allocation decision is to be made. Then, expressing each option in terms of
the net benefits it is expected to produce over time allows a direct comparison
between alternative strategies.
Consider the following example. Assume a policy-maker wants to know which
option, prevention or eradication, is a better strategy in the long term for a
pest with the properties we have ascribed to Colorado Potato Beetle in
Chapter 4.1. Say that a prevention technology is available to them that would
have a fixed cost of £175,000 (i.e. incurred in the first year), and a further
£10,000 annually thereafter over a 30 year period 44. If it was known that such
a technology could reduce the probability of entry from Very Low (i.e. a
44
Assume no capital depreciation to keep things simple.
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A New Agenda for Biosecurity, August 2004
probability of between 0.001 and 0.05 – see Table 2.1) to Extremely Low
(between 0.000001 - 0.001), the shift in expected invasive impact over time
brought about by the reduction in the probability of entry would resemble
Figure 7.6. This is exactly the same type of effect described in the previous
section (i.e. Figure 7.4). The total expected benefit generated by the new
technology is represented as the vertical distance between the two EI curves.
£350,000
£300,000
EI Without Prevention
Technology
£250,000
£200,000
£
£150,000
EI With Prevention
Technology
£100,000
£50,000
£Figure 7.6:
Total Expected Benefits of a hypothetical prevention technology for Colorado
0Potato Beetle
10
20
30
Year
If we now deduct the annual costs of the new technology, we derive a flow of
expected net benefits, illustrated in Figure 7.7. This benefit flow begins with
the £175,000 investment in the technology. Although benefits are produced
as soon as this technology is in place, they are initially insufficient to offset
costs. Hence, expected net benefits are negative over short time horizons,
but steadily grow for longer time horizons as the effects of the technology
begin to reduce expected invasion impact over time. Eventually the expected
benefits produced by avoiding invasive species outbreaks are sufficient to
offset the initial cost outlays. In other words, the total expected benefits up to
any point in time are represented by the vertical distance between the
horizontal (or time) axis and the Expected Net Benefit curve.
£300,000
£250,000
£200,000
£150,000
Net Benefit
£100,000
£
£50,000
£0
-£50,000
10
20
30
-£100,000
-£150,000
-£200,000
Year
Figure 7.7: Expected Net benefit stream for prevention technology
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A New Agenda for Biosecurity, August 2004
The present value of this expected net benefit stream can be determined by
adding the discounted expected net benefits in each year over 30-years.
Effectively, this provides the decision-maker with a single, comparable
measure of the desirability of the prevention option, i.e.
TBi  TCi
i
i  0 (1  d )
n
PV (TB0 ,..., TBn )  
(18).
where;
d = discount rate;
n = number of years;
TBi = total expected benefits occurring in the ith year;
TB0 = total expected benefits received immediately.
Using this criterion in conjunction with the stochastic model of Chapter 3, a
distribution of net benefits can be calculated in exactly the same fashion, and
is shown in Figure 7.8. In this case the present value of net benefits from
pursuing the prevention option has an expected value of around £50,000 over
30 years.
X <=£0
5%
X <=£890,000
95%
Mean = £50,000
Figure 7.8: Distribution of the present value of net benefits for the prevention option.
-3
-2
-1
0
1
2
3
4
Values in £ Millions
The option of eradicating outbreaks as they occur can be examined in a very
similar way. Assume that each outbreak has a fixed size and eradication cost
of £200,000. That is, each time an outbreak appears over a 30-year time
period it is immediately eradicated at a cost of £200,000. The total benefits of
eradicating an outbreak can be determined using the EI curve by setting the
probability of entry and establishment equal to one (i.e. an outbreak has
occurred). Note from Chapter 4.1 that at its highest point, the expected
impact curve of Colorado Potato Beetle reaches £325,000. This is equivalent
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A New Agenda for Biosecurity, August 2004
to the situation for a probability of entry and establishment equal to one,
because if we take a long enough time horizon the probability of entry and
establishment for this invasive species at some time over this time horizon will
in effect be one.
The number of incursions expected over a given time frame (like 30 years)
can be estimated from the model using Monte Carlo simulation. For simplicity
in this example, the probability of entry and establishment of the insect
reduces to one significant outbreak every 40 years (0.025) 45. Assume that
this incursion takes place in the middle year of a 40-year time horizon in order
to give an “average” impact. Using equation (18) and an assumed discount
rate of 7 per cent, it is possible to calculate the expected net benefits of the
eradication strategy in the same way we did for the prevention strategy.
TB20  TC 20
(1  d ) 20
£325,000  £200,000

 £32,300
1.07 20
PV (TB20 ) 
(19).
In this particular example, the net benefits expected from pursuing a
prevention strategy outweigh those anticipated from an eradication strategy
(i.e. compare £32,300 with the mean of the distribution depicted in Figure 3.9,
£50,000). Note that a probability distribution of expected net benefits from
eradication could also be produced by simulating an outbreak at any time over
the 40-year time horizon, but that the mean value of this probability
distribution will be approximately £30,000.
This is a purely hypothetical example. If instead the cost of eradicating an
outbreak was £70,000, then the expected net benefits from eradication (i.e.
£65,900) would exceed those of prevention.
In summary, this sort of economic decision rule, if used in a biosecurity
resource allocation context, will enable a reasoned choice to be made
between prevention and cure policy options for the net benefit of society.
Specific cases will require specific analyses. The precise tools to conduct
such analyses require further development, and comparison to more complex
but potentially precise dynamic modelling alternatives. This approach,
however, has the value of allowing broad comparisons between groups of
organisms, patterns of impact and approaches to control.
7.5 Multiple technological options
It may not be the case that just a single technology is available to use against
an invasive species outbreak. Often, there will be a range of techniques for
removing an invader from affected areas, each of which has its own unique
TC schedule. It follows that some alternatives may be better than others in
terms of efficiency and the level of net benefit delivered to society.
Figure 7.9 presents an invasion scenario where there are two alternative
management technologies available, A and B. The first, characterised by TCA
45
Here we use Markov chains to determine the equilibrium or fixed probability of entry and
establishment vector. See appendix 1,
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A New Agenda for Biosecurity, August 2004
TC,TB
in the top frame, is relatively costly when compared to the second,
characterised by TCB. There are greater cost increments associated with
using technology A to produce each additional unit of TB than the second.
Eradicating an outbreak with technology B is the cheaper alternative. Using A
would still earn a net benefit (of NBA), but this is small in comparison to the net
benefit generated by eradicating with technology B, NBB.
EDcrit
TB
NBA
TCA
NBB
TCB
0
E erad
Management Effort
Figure 7.9: Alternative management technologies
Note that the EDcrit cost ceiling does not intersect the TCA or TCB schedules.
Thus, a net economic gain is expected from using either technology to
eradicate the invasive. Also note that Eerad does not represent a net benefitmaximising level of management effort with either technology.
7.6 Technical change and how to value it
Associated with this Horizon Scanning project, Defra asked its Agencies to
consider present procedures and future trends in biosecurity, with an
emphasis on new approaches to prevention, detection, containment,
eradication and management. The resulting State of the Art Review of
Biosecurity Risk Management has contributions from




CEFAS on aquatic environments
CSL on plant health, bee health, mammals, birds and related diseases
IAH on exotic viral pathogens
VLA on animal diseases and veterinary public health
All submissions identified non-native species introductions as increasing in the
future in their particular sectors, mostly due to trade liberalisation and travel. A
few counter-trends were identified, e.g. the reduced movement of live
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livestock would reduce introduction by that pathway, but the overall
impression was one of a growing number of pathways of introduction and
growing traffic along them.
All submissions identified technological advances that will broaden the scope
and increase the speed of detection, identification, response and control.
Much of this relied on advances in information technology and biology,
particularly molecular biology.
The key technological advances associated with improved prevention,
including interception at entry, were:






Improved communication on new threats and movement of species –
joined up databases at a national and international level will facilitate
and speed the exchange of diagnostic tools, biological information and
information on distribution and movement. Formal, governmental
information sharing may be superceded by informal specialist
networks.
Improved legislation, with the addition of an “EU layer”, and a trend
towards risk management and self-imposed industrial codes of conduct
which will move the cost of prevention from the public to the private
sector.
Widespread adoption and harmonisation of risk assessment methods
across sectors, including stochastic modelling approaches such as that
adopted in this project
Methods to label and trace organisms or their containers, so as to
identify origin, where they have been in transit and what has been done
to them
Rapid detection of new organisms on imported material or arriving
people/animals involving sophisticated scanning methods. This may
include remote sensing of animal or plant populations to detect
infection or attack by non-native species (e.g. transmitters implanted in
domestic animals which report animal health status)
Rapid identification of new species or strains using molecular methods
(PCR, micro-array)
Key technological advances associated with improved eradication were



New technologies for control, including improved chemical, biological
and genetic methods and improved delivery systems. Marine systems
show the poorest prospect for development of eradication systems, but
even here new research is underway on acceptable and effective
chemical and biological measures (GISP 2004).
Development of portable devices for local detection and identification of
relevant organisms, connected to real time GIS systems for
management of a control campaign.
Improved economic models that allow “running cost benefit analysis” to
determine the marginal value of further control effort
In light of the models considered in this Chapter, there appear therefore to be
substantial prospects for improving prevention and eradication. With respect
to prevention, the challenge is to develop methods with a broad coverage
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(e.g. a range of potential invasives), so that the development and operational
cost can be spread across more than one benefit stream. Otherwise, the
target species will have to have large potential Ecrit to support development
and implementation of the new technology. Methods which allow traceability
and rapid identification of a range of species may therefore be most
economical.
For eradication, we have modelled a static situation, presuming that each
eradication has a known cost. In reality, this will depend critically on degree of
growth and spread of the introduced species, with cost of eradication possibly
increasing exponentially over time since establishment. Hence those
eradication methods which are most rapid may be most cost effective. Local
diagnosis, logging and transmission of data and GIS systems will be key to
controlling an outbreak before it spreads. Stochastic approaches derived from
the concepts in the static models can allow the inclusion of risk in later
development of these methods.
Putting these priorities together, the set of new technologies which would
seem to have particular benefit for both prevention and eradication are those
which allow the rapid detection and identification of a range of target species,
connected to a logging and transmission system that permit targeted action
(e.g. eradication). The development of these generic, rapid detection and
diagnostic tools may constitute the best government investment in research to
make prevention and eradication more cost effective.
A specific point must be made about vaccination, and the general issue of
making species or ecosystems resistant or resilient to invasion, so as to
supplant entirely the need for prevention or eradication. Our models have
shown the substantial effect which export constraints have on the economic
impact of non-native species. This compels governments to eradicate these
“listed” organisms, usually animal, fish or plant diseases, and those costs may
be high, with many indirect elements, as seen in the recent FMD control
programme. Eradication methods like culling are both expensive and
unpopular.
The development of vaccination as an alternative to export bans, eradication
and culling has enormous potential economic advantage. Not only may it be
far cheaper than an extensive eradication campaign, but the cost can be
moved from the government to the private sector, even to the extent that
vaccination was a legal requirement, paid for by the producer. While the
tradition of the OIE and related international agreements is strong, it is not
likely to stand against such a strong economic argument, once effective
vaccines can be developed. This concept can be extended to similar forms of
“resistance”, e.g. the genetic engineering of crops and animals for resistance
to pests and diseases. Once again, this will enable government to move the
cost of protection on to the private sector, namely the producer who
purchases the resistant strains.
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Chapter 8 – Conclusions
This project was predicated on the hypothesis that national biosecurity could
be approached on a cross-sectoral basis through a general ecologicaleconomic model that would allow quantitative comparison of risk and impact.
Policy makers who presently face demands from different agencies for
biosecurity support for different kinds of problems have a difficult time
comparing their relative importance, despite their common characteristics of
introduction, spread and impact. Our method has provided a tool by which
policy makers may undertake such quantitative comparisons. In this project,
we have presented this tool, and some examples of how it might be
developed, but we have not applied it. It is likely to require much testing and
refinement before it has operational value in this way.
However, we have been able to use this tool to extract some broad features of
biosecurity systems which will benefit policy, independent of application to
specific situations. Below we summarise these findings, all of which come
from the integration of ecological and economic thinking and a commitment to
looking “across the silos” in which biosecurity has traditionally been
developed.
8.1 Are biosecurity risks increasing?
Concerns about “growing biosecurity risks” are often based solely on a
perceived increase in international movement of commodities and peoples.
There is a poor evidence base that risk is increasing, and yet many institutions
are seeking new funds to anticipate and address this change. We have sought
to establish an evidence base to test this hypothesis, using available
databases to examine the trend in establishments over time. We can conclude
that, over recent decades, with two exceptions (vertebrates and aquatic
species), establishments of new species have been increasing. The trajectory
of this trend shows no indication of levelling off, although we have argued that
this is theoretically possible as species pools are exhausted. There is also
evidence of substantial accumulation of future threats, e.g. with plants and
their long lag period in establishment and spread. Finally, an ecological
analysis of introduction rates, based on species pool theory, suggests that for
every element contributing to such rates, there is an increasing trend: greater
volume of trade, increasing sources of species, improved transport pathways
and improved biotic/abiotic conditions for establishment.
We conclude that the hypothesis that risk is increasing as a result of
increasing introduction and establishment is probably valid. However, we are
concerned that a better evidence base be established, so that this risk can be
quantified across different kinds of species and sectors, lest policy and
investment in prevention be misdirected in future. Our studies have revealed
that there is considerable historical evidence of interception and introduction
in Defra agencies and other organisations that could be compiled and
analysed to build this evidence base. CSL is already doing this with insect
pests and diseases (R. Baker, pers. comm.). The benefits of this go beyond
the prediction of future “base rates of introduction” for models such as we
have developed. The pattern of introductions in the past – origin, taxa,
affected ecosystems, rates at different periods, etc., provides a valuable
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database to test hypotheses and understand processes governing
introduction. For instance, we might ask of such an analysis “to what extent
have certain regions been exhausted as a source of new problems?”, or “what
is the relationship between specific trade volume in relevant commodities and
rate of establishment?”
8.2 Can we take a general approach to predicting the economic impact
of future introductions?
There are a wide range of possible modelling approaches to predicting the
economic impact of non-native species. In our discussions with Defra
agencies, it was clear that several, different modelling approaches were being
used by different groups, usually for the evaluation of particular species under
particular assumptions. The cost effectiveness of prevention or control
measures was a particular interest in such modelling.
Our model was developed for generality, in contrast to most modelling efforts
to date. We have chosen a stochastic approach, using @Risk. Defra agencies
are also attracted to this approach. We chose a stochastic modelling process.
This is ideal for simulating the full range of uncertainty that may arise from a
potential invasion, but the downside is that the costs of obtaining reliable
uncertainty data are high. We have also chosen a static modelling approach,
i.e. looking at the likely impact of a new species or its control over a fixed
period. An alternative would be a deterministic, dynamic model,
parameterised with median values. This has some advantages, using a static
model we have had to assume that a non-native species is either excluded or
eradicated, a static model cannot handle as easily an option whereby a
species is managed at a certain level – here a dynamic model is more useful
to predict the balance between control of the pest and its growth and spread.
However, for our broad approach, comparing species and methods over long
time or “programme” periods of 20 years, we believe that a static, stochastic
model is more useful to policy making.
While our case studies can be much improved on in terms of
parameterisation, we believe that they show the capacity to generate useful
and comparable impact predictions over significant time periods for a wide
range of non-native species problems.
However, we encountered one very significant problem in the development of
such models, namely the need to quantify impact of non-natives on goods
without market value (e.g. biodiversity and some other environmental goods).
Our survey of non-native taxa revealed the level to which environmental rather
than agricultural effects are now contributing to or even dominating the likely
overall impact on non-native species. However, we simply do not have the
evidence to properly evaluate the non-market impact of non-native species
problems, and as a consequence we do not have the opportunity to generate
quantitative measures that combine non-market with market impacts to give
an overall value. Development of such evidence is critical.
A semi-quantitative approach is, however, possible. For instance, we can
develop a system of categorisation that takes into account market impacts,
non-market impacts and other “social effects”, and the confidence which we
can place on our impact estimates. We may present these elements as:
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A New Agenda for Biosecurity, August 2004
1. Expected Impact (EI) – the expected quantifiable (market) economic
damage caused by a species becoming established, as per our model
2. Confidence – the extent of variability in the EI estimate, and therefore the
level of confidence that a policy-maker can have in this as a measure of
biosecurity significance, as per our model
3. Non-Market and Social Effects – all other effects of a species becoming
established, including non-market environmental effects (both use and
non-use) and indirect effects on e.g. human health and rural economies.
Each of these categories can be “scored” as High, Medium or Low (i.e. 1 – 3)
for each species. Further assume there are three pests of concern, A, B and
C, and that scenarios expected to characterise future invasions are X, Y and
Z. There is also a control case where the circumstances surrounding an
invasion represent those of the present. The available information concerning
economic, probabilistic and non-market implications of pest introductions can
be combined in an Impact Table, an example of which is provided by Table
8.1.
Table 8.1: Impact Table
Scenario
Pest
Expected
Damage
Confidence
Social Effects
Score
A
High - 3
Low – 1
Medium - 2
6
B
High - 3
Medium – 2
Low - 1
6
C
Medium - 2
Medium – 2
Medium - 2
6
A
High - 3
Medium – 2
Medium - 2
7
B
Low - 1
High – 3
Low - 1
5
C
Medium - 2
Medium – 2
Medium - 2
6
A
Medium - 2
Medium – 2
Medium - 2
6
B
Medium - 2
Medium – 2
Low - 1
5
C
Low - 1
Low – 1
Medium - 2
4
A
High - 3
High – 3
Low - 1
7
B
High - 3
Low – 1
Medium - 2
6
C
Medium - 2
Low – 1
Medium - 2
5
Control
Case
X
Y
Z
Simply adding the scores attributed to each pest enables them to be ranked in
order of severity (or strategic significance) for each scenario.
i.e.
Control Case
A=B=C
X
B<C<A
Y
C<B<A
Z
C < B < A.
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A New Agenda for Biosecurity, August 2004
In this way we can identify how to achieve the highest expected social gains
from the investment of limited biosecurity resources under a range of possible
circumstances. Here, for instance, although the threat posed by each pest is
equal in the control case, an examination of anticipated pest impacts under
the three scenarios X, Y and Z clearly identifies A as the pest with the highest
strategic significance.
8.3 Are some kinds of risk consistently more important than others?
Available data and case studies suggest that introduction of non-native
species that restrict export of commodities will have a much more dramatic
effect than those which do not. Non-market effects can greatly intensify
impacts of non-native species, complicating the issue, but a species with
strong export-restricting effects, strong environmental (e.g. biodiversity
effects) and indirect effects on human health will always have impacts several
orders of magnitude greater than “average” harmful non-natives. A number of
animal diseases presently threatening the UK have these properties.
Further, our study suggests that the predicted impact we attribute to a new
species will depend critically on the time horizon chosen. This means that
priorities over a ten year period may “cross over” and reverse on a 20 year
timescale.
Most importantly, we postulate, but do not prove, that non-native species
whose effects are principally environmental (and possibly largely non-market),
will show delayed, acceleration impacts over increasing time horizons. Hence,
they may have impacts which cross over with those of agricultural pests on
long time horizons. This raises a critical policy question – should we invest
limited resources in preventing dramatic, near-future invasions which affect
agriculture in favour of preventing long-term invasions which will degrade the
environment for future generations?
8.4 Are future societal trends going to change risk substantially?
Likely changes in the environment (climate change), the economy (trade
liberalisation) and society are all more likely to increase than to decrease the
risk of non-native species invasions in the future. However, the model also
shows that current arguments for increased investment in prevention and
eradication of non-native species, in order to anticipate these future trends,
are overly simplistic. Scientists, including those from Defra agencies, argue
that an increase in the rate of introduction and establishment of new species
will increase economic impact. Hence, the government should be putting more
resources into biosecurity in future, as trade liberalisation increases this rate.
However such an argument does not recognise the other economic effects of
trade liberalisation, namely the fall in income from domestically-produced
commodities that must compete with (cheaper) imports. This element of trade
liberalisation will reduce local production, and hence the damage to the
economy, countering the effects of increased trade. Tomorrow’s consumers
may prefer reduced protection of UK agriculture, if this will open markets for
even cheaper imported agricultural goods. Trade liberalisation effects on the
rate of introduction of non-native species do not in themselves justify
increased investment in biosecurity. Traditionally strong responses to nonnative species threats in countries like Australia, New Zealand and the USA
152
A New Agenda for Biosecurity, August 2004
reflect in part a strong commitment to the protection of the domestic
agricultural sector, which may not be a feature of UK government policy in the
future.
The other important conclusion of our horizon scanning is that the most
distinctive trend in all non-native species issues, namely their impact on the
natural environment, may change in dramatically different ways in future. On
the one hand, a populace whose increasing wealth leads to greater valuation
of the rural environment and the preservation of species and habitats
threatened by non-native invaders may drive a strong demand for non-native
species prevention and management, as has occurred in the USA and
Australasia. On the other hand, UK lacks the clear distinction found in these
countries between “wilderness” and “settled landscapes”, and its tradition of
assimilating species, re-inforced by its growing cosmopolitan nature and
changes brought by global warming, may cause a populace to be less
concerned about changes in species, as long as an ecosystem is healthy,
attractive and sustainable. These divergent future scenarios require further,
sociological research to resolve – at the heart of them lies fundamental issues
about the future of conservation of great relevance to Defra.
8.5 Can we prioritise investment in control methods?
In this project, we have chosen not to use our model to analyse different
specific control methods for specific pests. Rather, we have developed a
theoretical approach to evaluating two general measures – prevention and
eradication. We argue that these are the substantive government options –
long term management of a chronic non-native problem should not be a
government activity. Such continuing management costs should be born by
producers or owners of the threatened commodity.
We have established an approach to compare the relative value of prevention
and eradication over a particular time horizon. In doing so, we challenge the
simplistic notion that prevention is better than cure (= eradication), and
provide a quantitative approach to making such decisions. Our analysis
suggests that there are no a priori reasons to favour one or the other
approach.
Likely future technology for prevention and eradication, as identified by the
State of the Art Review of Biosecurity undertaken by Defra institutes, will
probably benefit prevention and eradication in equal measure. Rapid
detection and diagnosis will reduce both introduction rates and allow
accelerated eradication programmes. While no clear bias is revealed by future
technology, there is the clear message that government should invest in these
technologies so as to make both prevention and eradication more effective.
8.6 How can policy makers use this study?
Currently, in making decisions about non-native species problems, policy
makers will use scientific advice and public opinion to create good
assessments of the relative importance of different current risks. However,
there are a number of factors which may distort the ability of policy makers to
direct the greatest resources to the problems of greatest national economic
importance:
153
A New Agenda for Biosecurity, August 2004





immediacy – the current crisis gets priority (“the squeaky wheel gets
the oil”)
institutional bias – certain taxa are prioritised because their mandated
agencies have historically greater power, while other taxa are “without
portfolio”
distributional issues – the disparity between winners and losers in
biosecurity problems
international commitment – the UK has made EU and international
commitments to prioritise risks and their management independent of
their national importance
intra-institutional ranking – specific agencies will rank risks amongst
mandate taxa, according to internal analysis of potential impact, and in
different ways.
In this study, we have come to five general conclusions which we feel will help
to address some of these biases and improve policy making on non-native
species problems. We summarise them below as bullet points:





the problem will get worse, more harmful species will appear,
biosecurity will need more resources, but our evidence base for more
precise prediction is poor.
it should be possible to establish and use economic models to compare
economic importance and impact of non-native species problems
across taxa, sectors and agencies; broad, cross-sector taxa patterns
emerge quickly from such analysis.
the environmental impact of non-native species is a “dark horse” on the
horizon – we do not have the evidence to evaluate it and therefore to
incorporate it into predictive modelling The nature of its development
and impact could make it far more important in future than we would
presently expect, and it is not at all clear whether the public will come
to view this as a big or a small issue as their environmental awareness
increases.
the growing demand across all agencies and sectors for increased
investment in prevention and technology should be assessed in the
context of the value of protecting the industry in question - trade
liberalisation, a principle cause of increased non-native species risk,
will also change the relative value of industries – consumers and
producers may have very different agendas regarding biosecurity.
government should focus on both prevention and eradication: there is
no a priori reason to favour one over the other; an economic approach
to evaluating these alternatives is preferable; and the current
opportunities for technological advance may improve both options.
These conclusions provide a “rough guide” to investment in non-native
species problems. There are some clear indications of opportunities for
research which will improve this process. The modelling approach which we
have developed here is demonstrated, but not tested or validated, and this
may be a useful next step.
154
A New Agenda for Biosecurity, August 2004
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Appendix 1: Finite Markov Chains
A dynamic mathematical model in which the probabilities of events in a time
period are determined by the occurrence of events in previous time periods is
known as a Markov chain model. A finite Markov chain is said to define a
system where an agent faces the prospect of one of a finite number of events,
X1, X2, …, Xn, occurring in any one time period, t. The probability of an event
Xi occurring in a time period, t + 1, conditional on event Xj having occurred in
period t, is pij. The probabilities pij (i = 1, 2,…, n; j = 1, 2,…, n) are positive
values, and sum to unity. These may be arranged in a transition matrix, P,
where i defines the row and j the column:
(A1)
P = (pij)
The elements in the matrix are conditional probabilities indicating the
probability of being in the “state of the world” defined by the row given that the
system was in the state indicated by the column in the previous time period.
This model can be applied in the current context, where the events we are
concerned with can be defined as a “with pest” state and “without pest” state.
If the initial probabilities of being in either state are specified, it can be
determined what the likelihood of being in a certain state in any future time
period. If we denote the probabilities of the events X1, X2, …, Xn occurring at
any time t by p1(t), p2(t), …, pn(t), we have:
pi (t  1)   pij p j (t )
(A2)
j
The set of all these equations can be expressed in matrix form:
pi (t  1)  Pp(t )
(A3)
where p(t) is a column vector with elements p1(t), p2(t), …, pn(t). By applying
this previous equation repeatedly, we obtain the following:
p(t )  P t p(0) .
(A4)
It can be demonstrated (Moran, 1984) that under a variety of conditions the
vector p(t) will converge to a unique vector p as t increases.
The initial probabilities attached to the with and without pest states of the
world will be dependent on the effectiveness of quarantine policies in place at
the outset of the analysis. So, changes to these policies will alter these
probabilities, and so different policies can be specified in this fashion (Hinchy
and Fisher, 1991).
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Appendix 2: Non-indigenous species in the UK:
exploring their meanings in human and social terms
Niall Scott and Claire Waterton, Institute for Environment Philosophy and
Public Policy (IEPPP), Lancaster University
July 2004
Acknowledgements
This social science research forms a contribution to the DEFRA Horizon
Scanning research project ‘A New Agenda For Biosecurity’ carried out by
Professor Jeffrey Waage and colleagues at Imperial College, London. We
would like to thank Professor Waage, Jo Pearson and colleagues at Imperial
for their support in carrying out this research. Part of the research involved
hosting a workshop at Imperial on 21st May 2004. Our thanks are due to all
who participated in the workshop, in particular Nigel Clarke of the Open
University and Judy Ling-Wong from the Black Environment Network for their
stimulating presentations.
Executive summary








Policy makers are doing the right thing in exploring and trying to
understand the historical, social and cultural contexts within which the
Non-Indigenous Species (NIS) debates are being played out.
Particular attention might be paid in policy circles to questioning the
underlying assumptions about what nature is in the UK context, as well
as the further question as to what NIS do in relation to that assumed
nature.
This implies an avoidance of thinking in essentialist terms about nature
as native/non-native, alien etc. As well as being ecologically
problematic, such terms are culturally insensitive.
There are various different ‘publics’ in the UK. These publics are likely
to judge the problem of NIS not solely in ecological terms but in the
context of their own empirical knowledge, as well as their knowledge
about institutions who are supposed to control environmental threats.
Attempts to enrol publics as informants or stewards of NIS may be a
useful way forward. Crucial to the success of such initiatives is to base
any proposed activities within existing value systems and to find ways
of reciprocating effort, so that belonging within such a vigilant
community reaps the right kinds of rewards for its members.
It would appear that the management of environment threats in the
future might well be based on approaches which avoid a static
reification of what nature and nature management is, instead turning
towards more flexible understandings of human-nature partnerships.
In the NIS debates, this would translate into a more pragmatic, perhaps
more human-centred (less ecologically centred) definition of NIS and a
correspondingly more flexible approach to their management.
The challenge for policy, should such a scenario unfold, would be to
ensure that certain agreed upon goals – e.g. maintenance of
biodiversity, ecosystem health, human health etc. – were able
simultaneously to be upheld.
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1.0 Background to the Research
This report presents the results of a research project undertaken by Niall Scott
and Claire Waterton at the Institute for Environment, Philosophy and Public
Policy between January and June 2004 in collaboration with Professor Jeffrey
Waage at Imperial College London. The research was commissioned as part
of a wider DEFRA Horizon Scanning project ‘A New Agenda For Biosecurity’
which aimed to predict likely changes in Britain’s biosecurity risk profile in the
future.
The problems of non-indigenous and invasive species have been the subject
of attention in international sphere for several years (e.g. IUCN 1985, 1987,
FAO 1995, OTA 1993, Wittenburg and Cock 2002). Recent reports suggest
an intensification of the perception of aliens as a significant policy/societal
problem (e.g. Mooney and Hobbs 2000). At a national UK level, nonindigenous species have presented relatively few radically harmful or
economic problems to date. However, DEFRA have recently sought to think
ahead towards potential future problems regarding non-indigenous and/or
invasive species. Whilst the DEFRA Horizon Scanning project (of which this
small sociological component forms a part) aims to characterise ecological
and economic models and scenarios that would assist DEFRA in projecting
policy planning forward, this research aims, in a complementary fashion:




To explore the issue of non-indigenous species in human and social terms;
to think about how non-indigenous species are characterised, and by
whom (which social groups use which terms, and how?);
to reflect on issues of definition, classification and naming, and the use of
terms by policy and regulatory bodies;
to mark out, on the basis of previous research on contemporary
environmental threats, the salient issues to explore in thinking about the
public perception of non-indigenous species.
The report is broken up into 5 sections:
2.
3.
4.
5.
6.
Historical and social context of non-indigenous species
Characterisations of non-indigenousness and of threat
Issues of definition and clarity
Public perceptions and the future
Conclusions and recommendations
2.0 The historical and social context
The benefits and the problems of non-indigenous species have been the
subject of international attention, correspondence and debate since at least
the early 19th century, the beginning of a period spanning just over one
hundred years in which over 50 million Europeans migrated to the NeoEuropean lands overseas (Crosby 1986:5). The colonisers brought with them
the flora and fauna they knew, much of which would thrive and become the
basis for new settler societies and for a vastly more globalized economy than
had ever been seen hitherto.
As Crosby notes of that period, however, ‘ the exchange of animals, tame,
feral or wild, between the Old World and New World has been as one-sided
as the exchange of weeds’ (ibid: 193). The Old World – Europe and the UK as
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part of it – was the recipient during that time of remarkably few non-native
species in exchange for the weeds, crops, stock and wildlife that it exported to
the new lands. Thus the British colonial experience of these early movements
of plants and animals was as a net-exporter of her own biota, as well as a
harvester and trader at-a-distance of the produce that they afforded from the
New World. This relationship was not only part of the success of European
Imperialism but an important basis for the present era of global free trade in
which we witness unprecedented human and non-human mobility and traffic
across the globe, as well as a kind of global scaling up of both economic,
social and ecological problems and issues. From the beginning of this period
of intensification of the traffic of humans, plants and animals there were
already differential benefits to different groups within society.
Arguably, one of the consequences of a more completely globalized world is a
tendency towards increasing homogeneity (Bright 1999, Mooney and Hobbs
2000). Today’s ‘economic performance’ of the world (measured, for example,
in terms of total global imports) outstrips any previous period in history. This
performance, however, has been built on ‘an increasingly homogenised
foundation of information, finance, culture and ecosystems’ (McNeely 2001). It
is within this context of the globalisation of trade and an unprecedented global
mobility and homogeneity of goods, services, humans, non-humans that a
current concern about non-indigenous species and the threats they may bring
appears to have arisen more prominently within the UK. As Clarke notes
(Clarke 2003: 165) many of these trends are consistent with Ulrich Beck’s
Risk Society thesis (Beck 1992) – a thesis which explores the phenomenon
whereby certain feedbacks within our social and economic system begin to
threaten the very system itself. The destructive feedback loop of the ‘Risk
Society’ is most graphically represented through economic assessment of the
lost revenue from selected harmful non-indigenous species originally
introduced in order to reap economic profits (e.g. OTA 1993, Wilgen et al.
1996, McNeely 2001), but it is equally relevant in contemplating a more
ecological viewpoint concerned more about the destruction of nature or
nature’s diversity (e.g. Elton 1958).
Those in the social sciences have been gently criticised for taking this idea too
far: ‘Such is the enthralment with the idea of a pervasive social undermining of
biophysical forces and processes that environmentalists and social scientists
alike are speaking of the ‘end of nature’ (McKibben, 1990; Strathern, 1992;
Giddens 1994)’ (Clarke 2003:165). In the UK, as we shall see in more detail
below, the concern about the risks of non-indigenous species has registered
particularly strongly in terms of threats to biodiversity rather than economy.
What is being recognised in addition to the idea that non-indigenousness can
be taken to mean a substantial threat, however, is the unpredictability and the
complexity, as well as the ‘naturalness’ and vitality (Clarke 2003) of
interactions now taking place within the contemporary human and biotic world.
So whilst the UK may have been an intentional exporter in past colonial times,
what appears to be of current concern is the possibility of accidental or
unintentional introductions occurring, perhaps leading to the spread of
introduced species which are difficult to control, unpredictable in their effects
and difficult to trace in terms of responsibility for causing such effects.
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Concepts of uncertainty, accident, unforeseen events and unpredictability
have undoubtedly taken on a greater significance in the national and global
psyche since September 11, 2001. A new orientation to the unpredictability of
life may also be translating into systems of governance such as regulatory
and planning bodies within government. This is part of the context in which the
interest in non-indigenous, invasive species has arisen.
Summary of Section 2:

The UK has historically been an intentional net-exporter of biota to the rest
of the globe.

As a corollary of processes of colonialism, increased global trade and
increased mobility around the world, the world is recognised to be
becoming both more complex and more homogeneous.

Part of the complexity of the contemporary world lies in its
interconnectedness and in the recognised capacity for unpredictable
events to occur.

Despite the recognition that different social groups within society will bear
more costs than others from the problems associated with non-invasive
species, the unintentional and relatively unpredictable nature of
introductions make the pinning of responsibility for ecological/social and
other potential (e.g. health) costs difficult to establish.

A concern about non-indigenous species in the UK has arisen in a context
where non-homogeneity is becoming increasingly valued (expressed in
environmental terms as biological diversity) and where the recognition of
complexity, uncertainty and unpredictability within systems of governance
is becoming more sensitive.
3.0 Conceptions of Non indigenous species amongst different actors
3.1 Mapping out different discourses
Part of the research aimed to map out where indigenous species were
entering into common discourse or debates in the UK, who was talking or
writing about them, and in what terms. This section documents some of the
varying ways in which NIS are currently being characterised, as far as a
literature search on NIS relating to environmental issues was able to
ascertain.
A large proportion of the literature on non-indigenous species, especially in
the context of the environment, concerns non-indigenous species as a threat
or potential threat. It is well known that much of this literature relates to parts
of the world where the historical context has been quite unlike that of the UK –
for example North America and the Antipodes. Much of the literature also
relates to ‘hotspots’ of concern – for example areas of high endemism found
in ‘small islands’. Much less research has been carried out from a UK
perspective. In surveying the literature regarding non-indigenous species in a
UK context, it was apparent that interest in the subject stems from a range of
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actors/social groups. Groups identified as contributing to literature about noninvasive species are:
1) Statutory environmental and conservation bodies and government
funded research councils, e.g. DEFRA, the Joint Nature Conservation
Committee, English Nature, Scottish Natural Heritage, Countryside
Council for Wales, the Natural Environment Research Council.
2) Environmental and conservation non-governmental organisations
(NGOs), e.g. the International Union for the Conservation of Nature,
Friends of the Earth, Greenpeace, Plantlife, The Black Environment
Network, British Trust for Conservation Volunteers, The Countryside
Alliance, The Soil Association, The Japanese Knotweed Society.
3) Academic writing within conservation publications and the ecological
and environmental sciences.
4) Articles within the sociological literatures including sociology of science,
philosophy and philosophy of science.
The majority of the literature dealt with NIS in the context of ecology,
economic impact, and conservation. There were fewer reflections upon the
issue in terms of social and cultural dimensions of NIS, with the exception of
one NGO - the Black Environment Network - where this featured strongly.
Threats were characterised by all the above actors under two main headings:
Biological: Threats to biodiversity, biomass, plants and animals themselves,
ecosystem integrity, diversity, extinction from both native and non-indigenous
sides.
Social: Threats to economy, people, communities, security, ideas and beliefs,
culture, history, change, human integrity.
The most commonly expressed threat concerns a general threat to
biodiversity, with a range of different specific emphases which we outline
below. At a general level, the debate in the UK is consistent with internationallevel debates located within bodies such as the International Union for the
Conservation of Nature (IUCN) and in debates being carried out under the
auspices of the UN Convention on Biological Diversity. The premise is
straightforward (as the statement below from IUCN indicates): biological
diversity globally is held to be at risk. ‘Biological diversity faces many threats
throughout the world. One of the major threats to native biological diversity is
now acknowledged by scientists and governments to be biological invasions
caused by alien invasive species.’ (IUCN 2000)
In the UK context, this threat to biodiversity is expressed by a range of bodies
each stressing different emphases. Thus the DEFRA Review of Non-Native
Species (Fasham and Trumper 2001) recognises the threat to biodiversity, but
also the benefits that can be brought to humans: ‘The impacts of non-native
species can be serious; they can transform ecosystems, damage crops, alter
habitats and threaten native biodiversity. Non-native species can also bring
considerable benefits in terms of both economic gains and quality of life.’
(Fasham and Trumper, 2001: 7)
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A negative human involvement is often recognised in the literature, implying a
sense of responsibility, for example by NERC : ‘Biological Invasions by nonnative or ‘alien’ species are widely recognised as a significant component of
human caused global environmental change often resulting in a significant
loss of the economic value, biological diversity and function of invaded
ecosystems’ (Birnie et al, 2004)
‘NIS’ as a concept is seen to be open to interpretation and the literatures
reviewed reflect this. As McNeely argues, the ‘‘noxious invasive’ of one
cultural group is the ‘desirable addition’ of another group’ (McNeeley 2001).
As we begin to look at the different discourses through which NIS are talked
about and represented, we can see that the potential threats of such
organisms are not solely be grounded in ecological criteria, but are strongly
bonded onto concepts used to identify origin, authenticity and responsibility
(Hattingh 2001). Groups can use the NIS issue opportunistically, perhaps
using NIS rhetorics to support prior goals. The Countryside Alliance (CA), for
example, as an organisation that claims to represent and promote the
interests of rural people, use the notion of non-nativeness in their targeting of
specific animals, such as the mink as a priority for control and eradication:
‘The Countryside Alliance along with many land use organisations is
of the view that the eradication of the American mink is a desirable
objective given the mink’s non-native status and devastating effect
on other wild mammal, bird, fish populations’. (Countryside Alliance,
2002)
For the Royal Society for the Protection of Birds (RSPB) the salient issue is
‘Wild Bird Crime’. Non-native species arriving in the UK through illegal trade
are highlighted as a significant threat to biodiversity: ‘This trade is one of the
most significant factors, after habitat destruction, driving species to extinction.’i
Upon considering the issue of hybridisation of the ruddy and white headed
duck, non-indigenous species are accorded a status of a primary threat:
“Globally, non-native species are considered the most important threat to
biological diversity after habitat loss.”ii
3.2 What do the different discourses tell us?
What we can begin to see by looking at the discourses of NIS employed by
different UK actors is first, that the NIS species issue is not one thing: it is
refracted through the different concerns of British policy and NGO institutions.
Second, in some cases NIS is referred to in a very general way. In others it is
taken up in the context of a specific example. Often it may be used as a tool to
strengthen an existing agenda or policy.
Third, as well as highlighting specific threats, NIS narratives within NGOs and
other groups give insight into what a desirable nature is held to be for such
groups. If we take a ‘constructivist approach to understanding environmental
issues (e.g. Macnaghten and Urry 1998) these threats to a desired nature are
not static and ‘given’. They are actively constructed through processes of
information exchange, issue definition, campaigning etc. and require social
endorsement to be robust. The debates and references to NIS therefore need
to be seen these terms – they are part of a flow of rhetorical as well as idealist
and ‘realist’ claims about an existing or desired state of nature.
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Fourthly, many of the claims will arise not just as rhetorical positions but
through deeply embodied relationships developed within certain social groups
(e.g. the hunting groups associated with Countryside Alliance). As such they
will be important to group members’ identities and allegiances.
The policy implication of these four points lies in the importance of recognising
that conservation bodies and NGOs are connected to ‘publics’ through their
memberships, their campaigning and their publicity. Although this study has
not empirically gauged the nature of public opinion about NIS, it may be likely
that some strands of policy and NGO campaigning discourses will be
recognised, embodied and practiced, supported and/or disagreed with by
different sections of the public. The way that NIS are (legitimately)
represented and understood, in other words, already derives as much from
the social and value-laden context as it does from reality in nature (usually
represented through numbers of non-indigenous species and a quantification
of their threat).
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Summary of Section 3

Many different policy, social, environmental, and campaigning groups take
up the vocabularies associated with Non-Indigenous Species in various
ways. We have only represented a few of such perspectives in this report.
But rather than understanding the claims and positions taken up by such
groups as ‘matters of fact’ (Latour 2004), a sociological perspective can
begin to show how such knowledges and claims are in fact highly
embodied, multiply produced (sometimes multiple claims may be made by
the same institutional actor), and interpreted within specific social, cultural
institutional, value-rich and political contexts. They might better be
described as ‘matters of concern’ – statements about nature that are
produced through a rich human, non-human, institutional and cultural
milieu (Latour 2004).

The implications of this line of argument is that these narratives do not
exist in a social or cultural vacuum but arise from society, springing forth
from existing values, practices and positions. They will therefore be (or
have the potential to be) recognised, interpreted and contested by ‘the
public’ in its various forms. These narratives therefore need to be
recognised, especially by those responsible (e.g. policy and decision
makers), as properly belonging to society, rather than deriving from the
facts of nature which may later be communicated to society. This may be
an important point to note when we consider some of the
recommendations that exist to educate and inform the public about the
issue of non-indigenous species in Section 5.
4.0 Issues of Definition: ‘Non-indigenous’, ‘Non-Native’, ‘Alien’, and
‘Invasive’ Species
The debate surrounding the definition of NIS may be important for the
communication and development of policy regarding NIS in a public context.
Of initial note is the observation that in the examples of literature quoted
above, (reports by DEFRA, NERC, Countryside Alliance and the RSPB), there
is little or no consideration of the potentially positive contributions that NonIndigenous Species can make to biodiversity, for example the possibility that
NIS trees can provide habitats that are effective at supporting a wide range of
species, both NIS and indigenous. This is the first point regarding definition: in
a broad sense NIS are seen in terms of their negative impacts. This finding
within the literature is mapped also within newspaper reports of NIS (see
Briefing Report for the Horizon Scanning Project ‘A New Agenda for
Biosecurity’ 20.02.2004: p. 19).
Definitions are achieved through three main sets of discourses:



discourses of native and non-native distinctions,
discourses of invasion/invasiveness
discourses of alienness.
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We examine each of these three sets of terms and the ways in which they are
used below.
4.1 Native/non-native definitions
The recent DEFRA Review of Non-Native Species Legislation and Guidance
(Fasham and Trumper 2001) adopts the definitional criteria outlined by the
IUCN. These include:
“Native Species (indigenous): A species, subspecies or lower taxon,
occurring within its natural range (past or present) and dispersal potential
(i.e. within the range it occupies naturally or could occupy without direct or
indirect introduction or care by humans).”
and
“Non-native species (alien, non-indigenous, foreign, exotic): a species,
subspecies or lower taxon occurring outside of its natural range (past or
present) and dispersal potential (i.e. outside the range it occupies naturally
or could not occupy without direct or indirect introduction or care by
humans). This includes any part, gametes or propagule of such species
that might survive and subsequently reproduce.” (Fasham and Trumper
2001: 7)
It has become clear in the process of carrying out the research, and perhaps
especially from the Workshop held o the 21st May 2004, that the native/nonnative distinction is neither historically nor socially meaningful in the UK
context. This is because the UK has a history of biological colonisation,
extinction and re-colonisation associated with glacial periods and human
movements, making it extremely difficult to define a clearly native fauna and
flora.
Existing scientific approaches vary in presenting degrees of nativeness by
presenting threshold dates, e.g. post-glacial, Roman occupation, 1500 a.d. A
definitional distinction made by the New Atlas of the British and Irish Flora
(Preston et al., 2002) is of ‘archeophyte’ (referring to natives) and ‘neophyte’
(for non-native species) for which an archaeophyte is defined as one that
became naturalised before 1500, and a neophyte as one that was introduced
into the British Isles after 1500, (or was causally present, but naturalised after
subsequent reintroduction). Using this definition, whilst archaeophyte plants
are to be considered as part of the UK biodiversity and cultural heritage and
should be given a conservation status equivalent to native species, neophytes
are generally not given this status. In some exceptional circumstances
neophyte plants also warrant conservation attention.
Using a particular date to provide a definitional criterion to give to the
indigenous/ non-indigenous distinction is supported by D.A. Webb (1985),
and is taken up in the DEFRA report (2001) as well as in ecological literature
(see for example Manchester and Bullock, 2000). Webb defines a native plant
as:
”one which has evolved in these Islands or which has arrived there
since that date by one means or another before the beginning of the
Neolithic period…an alien on the other hand is one which reached
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the British Isles as a consequence of the activities of Neolithic or
post Neolithic man or of his domestic animals” (Webb, D.A. 1985).
This gives an approximate date at 6000 years BPE for the cut-off point for
something being designated as non-indigenous. Such a definition is arrived at
through ecological/scientific criteria which include fossil evidence, historical
evidence, habitat, ease of known introduction, geographical distribution
genetic diversity, reproductive pattern and supposed means of introduction.
Regarding public response to a definition that uses a particular date to
establish non-indigenous status, it is worth considering whether a date such
as 6000BPE is relevant in terms of its significance outside scientific/ecological
discourse. Such a cut off point may well rule out species that are
characterised as ‘neophytes’, and therefore non-native, but that are now, in
cultural terms, considered to be quintessentially British, such as the Horse
Chestnut.
Since there exists no single scientific basis for native or non-native status of
an organism, and given the acknowledged contingency of scientific definitions
in this area, a flexible approach to the definition of native/non-native is
desirable. Indeed the utility of the terms native and non-native in a public
policy context deserve questioning. Scientific approaches often fall short when
there is a need to accept the variability and shifting nature of the boundaries
of what is native and what is non-native.
What perhaps deserves greater recognition is that terms such as non-native,
‘alien’ etc., even when used within a scientific context, have strong human
political and cultural dimensions, where, for example territorial associations of
national boundaries can be (falsely) associated with habitat boundaries.
Episodes within UK history need to be recognised as having an impact in the
construction of what does/does not belong in the biological/natural realm.
This may mean a switch to understanding species more within the context
(including the human, cultural, political and institutional context) within which
they are found.
4.2 Invasiveness
‘Invasive alien species’ is used as a scientific short hand for a highly complex
ecological phenomenon. The properties of outbreak, displacement of other
species, invasiveness, or the introduction of new diseases harmful to resident
species are however not necessarily associated with being alien species. To
this end the suggestion by Kirsten Schrader- Frechette (2001) is useful in that
it highlights both the context in which a species occurs as well as the
behaviour of species. Shrader-Frechette suggests a definition classification
based largely on what the organism in question does - how it behaves in
relation to its environment. Thus species can be regarded in terms of:



Short dispersal distance/ Long dispersal distance
Novel/Common (to the area being colonised)
Minimal Impact/Great impact
where ‘long dispersal, novel, great impact species’ are seen to present he
most problems. This is a useful classification as it assesses the current status
of an organism in its habitat with a view to weighing up whether it is likely to
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cause a problem or become invasive. In the CBD, an invasive alien species is
defined as one that threatens biological diversity.
In general the ‘invasiveness’ of a species, broken down into an understanding
of ways in which the qualities of invasiveness arise (as in the definitions of
behaviour bulleted above) can be a useful denominator within the policy
context. An accumulated sense of invasiveness implies an understanding of
the behaviour of a species as well as of the surrounding context of a species:
especially that which is given to be under threat/likely to be overcome. The
negative and militaristic connotations of the idea of ‘invasion’ are, however,
worth considering as potentially antagonistic and resonant with political
scenarios that may be unhelpful in the public realm (see following section).
4.3 Alienness
The problem of connecting racist language and xenophobic attitudes to issues
of NIS has been the subject of commentary and debate within scientific and
popular scientific literatures (for example Daniel Simberloff (2003), Michael
Pollan (1994) and Mark Sagoff (2000)). Many ecologists may be wary of the
connections that might be made. Baskin, for example, talks of ‘lurid charges of
xenophobia or ‘ethnic cleansing’’ and criticises the use of ‘sloppy use of terms
which can provide grist for critics’ (Baskin 2002: 298).
But whilst Baskin criticises the use of ‘sloppy’ language, she also
demonstrates how powerful discourses can be, as well as the difficulty of
controlling them. The term alien is found in many definitions and descriptions
of Non-Indigenous Species. Territorial and politically conservative resonances
of the term ‘alien’ frequently accompany it, raising some questions as to the
utility of the term. The work of the NGO, the Black Environment Network
(BEN), holds that the term alien and discourses surrounding it generate
unhelpful racial and militaristic analogies. The xenophobic language of alien
species will cause discomfort to many, especially to ethnic minorities who see
the application of alienness as derogatory in the context of immigration and
asylum seeking and even as life threatening due to recent acts of racial
hatred. BEN’s argument is that policy terminology needs to be sensitive to
human cultural and ethnic identities recognising multiculturalism in Britain.
Thus threats presented about the NIS issue in the social context need to
include a consideration of language in relation to identity, community, ideas,
beliefs and cultural history.
As BEN rightly point out, immigration terminology has the potential of being
echoed in the environmental and ecological arena, where the term ‘alien’ can
be equated with bad and native can be equated with good. Other terms that
are considered offensive include: biosecurity, non-native, non-indigenous,
invasive (implying intention as distinctive from ‘outbreak’ which is more
descriptive). Phrases such as ‘Rhodo –bashing’ and policing the borders have
strong negative links to racial discrimination.
But there are further problems with the concept of alienness. The concept of
alienness is commonly related to issues of territory and space and bears the
assumption that alienness poses a threat to a background scenario of stasis
and ‘integrity’. Thus nature becomes both temporally and spatially charged
with an idea of stopping still (or having stopped still at a certain point) that can
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be seen as parochial, non-cosmopolitan, even paranoid. Such ideas can be
seen to be in tension with feelings of cosmopolitanism, change, or simply an
openness and acceptance of ‘the new’. Feelings of stasis are also in tension
with the active creation of a sense of belongingiii. Conservationists (e.g.
Barker (n.d.), Rodwell, see Footnote 3) now argue that it is at least
questionable, and perhaps undesirable, to freeze species and associated
habitats, and the values attached to them, in time.
Alternative discourses to that of alienness might be more site and context
specific using descriptive means, such as ‘Species Established in the Wild’;
‘Species Recently Established’; ‘Species Supporting Attractive or Desirable
Others’. This approach can be considered to reflect Schrader-Frechette’s
definitional criterion given above, giving more of a sense of what particular
species do, or what their function is, rather than a sense of what they are
deemed to be (alien, non-native etc.) terms that are forged in relation to a
much more abstract scale of time and space.
4.4 Definitions: a need for clarity?
Kirsten Schrader-Frechette (2001) in a philosophical approach to the definition
of Non-Indigenous Species bemoans the lack of consensus regarding the
terminology in the debate and the confusing ambiguities that remain, even in
the field of ecology. She criticises Webb’s (1985) definition as arbitrary and
stipulative and puts forward the view that Non-Indigenous Species should be
referred to in a more context specific way which gives a sense of the activity
and behaviour of species in question.
Whilst such an idea is difficult to contest, the mutability and performativity of
languages as well as organisms deserves some attention (Clarke 2003). A
disciplining of language use is unlikely to erase the culturally rich (as well as
sometimes offensive) uses of descriptive terms for the kinds of life that are
being referred to in the NIS domain. Historians (e.g. Thomas 1985, Foucault
1970) have shown the intimate connections that humans are accustomed to
make between ideas of social and natural order and such social/natural
border crossings in the realm of language, analogy, norms etc. are unlikely to
cease with respect to this specific issue. The implication of this is that, no
matter what clarity may be required in terms of definition, it may still be
important to understand the cultural terminologies and resonances that certain
‘natural’ ideas may engender.
Summary of Section 4

The contingency of the scientific definitions of Non-Indigenous Species
underlines the need to question the different vocabularies used to describe
them.

Vocabularies of alienness and non-nativeness in particular arouse strong
negative and racialist connotations which could be avoided through
steering away from the use of these terms.

Definitions alluding to the behaviour of a species, which may add up to a
picture of ‘potential for outbreak’, or ‘invasiveness’, are useful definitional
tools.
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
Reference to the context in which species are found is another means of
avoiding essentialists and territorially provocative claims about ‘native’ or
’alien’ species.

Although the need for definitional clarity cannot be denied, a sociological
approach would simultaneously seek to understand all the different
terminologies and rhetorical connections in terms of the way in which they
are used in everyday discourse. Calls to ‘discipline’ language will only be
successful in certain spheres and policy makers will still need to
understand the wider public resonances of the terms they adopt and
promote.
5.0 Public Perceptions
The question of public perceptions of Non-Indigenous Species initially
appears to be a difficult and evasive issue, since it is likely, as McNeely notes,
that ‘few people in any part of the world consciously perceive that they have
been negatively affected by IAS [invasive alien species], either directly or
indirectly’ (McNeely 2001).
However this does not mean, as discussed in Section 3, that publics are
somehow out-of-touch. Since this study has relied only on desk material we
draw largely in this section on other studies that have tried to gauge the way
that people perceive environmental threats. By making analogies with others
studies of environmental risks we may get some insight into the ways in which
people in the UK might engage with the issue of non-indigenous species.
But first, we need to explore what it is that people might be responding to,
should they be asked to think about non-indigenous species. What, in other
words, would the object of their perceptions be? This is partly a theoretical
point, but it has clear implications for the way in which policy anticipate and
understand public reactions.
The following table adapted from Marris et al. (2001) suggests that the object
of public responses to risks, far from being a straightforward characterisation
of the risk in question, is much more complex:
‘Previous research on public perceptions of risk has revealed that the object of
public responses can be any or all of the following:




Risk magnitudes, as described by scientific authorities, for example as
death frequencies;
Risk qualities, e.g. psychometric attributes described by Slovic et al. such
as voluntariness, risk/benefit distribution, catastrophic potential, risk-trend,
familiarity, visibility etc. (Slovic, 2000);
Institutional mismanagement of those risks;
Dominant institutional definitions of the issue as imposed in official
approaches ( e.g. neglect of dimensions and variables which are salient to
the public);
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

Dominant definitions of the public as undifferentiated and stereotyped (e.g.
as ignorant, prone to hysteria, instrumental only, individualistic) implicit –
and often explicit – in expert discourses of the issues (and also in some
research approaches);
The technology as a whole social experience and projection.’
Adapted from Marris et al. 2001: 19
Marris et al. (2001) argue that risk perceptions research needs to reflect the
breadth of issue through and within which publics will form perceptions.
Perceptions of risk therefore need to be understood not just as focussed on
an objectified risk (the likelihood of X event occurring with Y consequences)
but also in relation to lay-expert dynamics, including how expert institutions
understand public responses.
5.1 The supposed problem of ‘lay ignorance’
Much of the literature on Non-Indigenous Species has stressed the need for
education and information campaigns to be informed by ecologists and other
experts and delivered to the public (e.g. IUCN 2000, Baskin 2002). Underlying
these calls is an assumption that people do not know enough about alien
species to have any judgement on them, or do not know enough to be able to
contribute in any way to controlling them.
This assumption tallies with evidence from other studies which have
demonstrated a strong mythology amongst policy and decision makers
concerning the limited knowledge of the public regarding technical regulatory
and environmental risk issues. In their study of the policy and public views of
Genetically Modified Organisms (Marris et al. 2001), a dominant assumption
amongst policy players in five European countries was that people were
ignorant about scientific facts and that this was the cause of a problem
concerning public responses to GMOs.
In the part of the study that looked at lay perceptions of GMOs (as opposed to
policymakers’ impression of lay perceptions), the study found that, although
citizens were largely ignorant of the scientific technicalities of genetic
manipulation, ‘this lack of knowledge did not explain their response to
agricultural biotechnologies’ (Marris et al. 2001: 9). Public perceptions
regarding GMOs were not, in the main, based on false beliefs about GMOs.
Instead, people raised issues based on:



their own lay empirical knowledge of the GM debate, including nonspecialist knowledge (for example gained through gardening and bee
keeping);
knowledge about human fallibility, based on their own daily experience;
knowledge about the past behaviour of institutions responsible for the
development and regulation of environmental risks.
Research carried out on public perceptions in other areas concerning
environmental threats have indicated that people’s expressed ‘attitudes’
towards a particular risk (e.g. nuclear risks, or hypothetically, the risk of an
invasive species clogging up a waterway) arise from their own lay knowledges
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and experiences. As such, they are often as much to do with public perception
of, and trust in, the institutions supposed to be controlling the risks, as they
are about any seemingly discreet or quantified risk itself (Grove-White et al.,
1997, Wynne, Waterton and Grove-White, 1993).
Environmental risks and threats in other words are normally judged by people
as risks in context. Thus public perception of non-indigenous species will
partly encompass the judged competency of authorities (such as DEFRA) in
the context of a particular issue. Using the right language, engendering trust
within different communities, building on existing networks and respecting
those who have reliable lay experience of invasive species will all be
important components of lay experience and perceptions of particular species.
5.2 Perceptions of non-indigenous species in the UK
Literature from NGOs and academic sources suggests that there is a
comparative lack of research that has been done on the public perception of
non indigenous species in the UK, except for relatively small surveys that
have been carried out in the context of a particular problem species or
concerning the possible re-introduction of a species (e.g. Panaman 2002;
Prowse 1997, Green 2002). Panaman’s study on public perceptions included,
for example, changes in attitudes towards the possibility of the re-introduction
of Wolves in the Scottish Highlands through education and increased
knowledge and understanding of the species.
Some work on perceptions has focussed on the perceptions of conservation
actors and workers. For example, work by Prowse (1997) surveyed attitudes
of people within organisations involved in conservation and environmental
management in the Northwest of England specifically regarding the
Himalayan Balsam. This survey produced results that showed that 89% of
respondents considered the plant to be a problem.
Paul Green (2002) suggests that current public attitudes in British
conservation towards alien plants has been shaped by an understanding of
negative impacts of alien species overseas, mainly through the reporting of
extinction. Examples he suggests as often quoted are the mass extinctions in
Hawaii and the Galapagos Islands - usually involving animal impacts on
unique ecosystems. He compares this to a much more relaxed attitude about
this issue in Britain in the late 1960’s, quoting Dudley Stamp’s review of
nature conservation in Britain: “despite the immense number of aliens
introduced it is difficult to point to any that form a menace to existing
vegetation”. Green’s opinion of the current state of public opinion is that views
are much more polarised, where, for example, alien equates with ‘bad’ and
native with ‘good’.
Green’s characterisation of attitudes towards non-indigenous species by the
UK public today is contested, however by certain other voices within ecology
and conservation debates. Rodwell (footnote 3) and Adams (1996) seem to
be championing the ability of people to celebrate not only native diversity but a
historically and spatially more complex version of diversity in which human
attachment, emotion and connection explicitly play a part. So, for example,
Adams cheers for the commonly found buddleia species which he sees
growing out of a crack in a multi-story car-park and Rodwell champions the
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sycamore as a culturally and ecologically important feature in certain farm
building settings in the North of England.
The importance of human connections within specific contexts would be a
useful way of exploring publics’ perception of NIS in the future. This would
mean, for example, that a survey (or focus group discussions) would not only
try to ascertain what people felt about a particular species, but about people’s
own connections (in broad terms) with plant and animal species including the
species in question. Such an approach would mean that the ‘context’
surrounding a particular species may include many different aspects. Issues
of framing and definition of ‘the problem’, as well as issues of control,
individual and collective agency would all be important contributions towards
an understanding of how people perceive the issue of Non-Indigenous
Species.
5.3 Public Engagement
Recently attention within both international and national circles has turned to
the question of engaging the public in the issue of non-indigenous species
(McNeely 2001, IUCN 2000, Baskin 2002, Wittenburg and Cock 2002).
Sometimes this engagement is envisaged in a way in which the public
becomes a kind of ‘task force’ . The London-based NGO ‘Plantlife’, for
example, have directed people towards vigilance and reporting of nonindigenous plant species and have requested public sensitivity and
responsibility concerning gardening practices, to be wary and aware of things
such as garden centre purchases. Other institutions such as the Royal
Horticultural Society have made similar requests of their members. Such
approaches are of great interest as they potentially connect what may seem to
be isolated and quantified environment threats with networks of people ‘on the
ground’. They therefore give scope for local practices, adaptations and ways
of dealing with potentially troublesome species to translate upwards into
policy.
At stake here is the forging of new forms of community who are able and
willing to
share the ‘policy problem’ and take some ownership and
responsibility for it. These may be knowledge – based or ‘epistemic’
communities who simply engage in reporting and monitoring (Ellis and
Waterton 2004). Or they may combine this epistemic input with some form of
preventative, stewardship or control practices.
Crucial to the success of such initiatives is to base any proposed measures
within existing value systems and to find ways of reciprocating effort so that
belonging within such a vigilant community reaps the right kinds of rewards
for its members. Of potentially great benefit within such a system is the fact
that those who may be most directly affected by the non-indigenous species
(e.g. farmers, anglers, boat owners, gardeners) might theoretically play a role
in monitoring and decisions about how to manage the problem. The nonindigenous species problem could learn from critiques of similar systems of
enrolment – for example that of farmers’ involvement in Countryside
Stewardship schemes (Morris and Wragg 2003) or anglers’ involvement in
river water monitoring (Waterton 2003, Ellis and Waterton 2004).
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5.4 Thinking about the future
Society today is dealing with the extremes of cosmopolitanism as well as the
tendency towards isolation manifest in fundamentalism or perhaps a kind of
parochialism. There are conflicting concepts at work here where
cosmopolitanism reflects mobility, change and adaptation, parochialism
reflects stasis and remaining where one belongs. It has been argued in this
context that human beings should be cosmopolitan, but other species ought
not to be. ‘We want a world in which people are as free as possible to travel
and to exchange goods and ideas…but at the same time we need a world in
which most other living things stay put’ (Bright 1999: 200, quoted in Clarke
2003: 176).
In the UK context, Bright’s argument resonates strongly with a preservationist
approach to nature – focussing on the continuity of a nature, separate from
humans, that already exists.
Much could be gained in exploring more evolutionary approaches understanding how species adapt to new environments and vice-versa; or on
the other hand, in exploring the potential of more dynamic conservation based
approaches which highlight how ecosystems function and can be restored
subsequent to disruption. These two latter approaches, each with their
different understanding of the relationship between humans and nature, could
lead to a more positive approach to NIS. The policy and management
implications of these different models are only just beginning to be tested in
the management of specific parts of the natural environment in the UK. An
example might be seen where natural processes of silting and flooding are
being allowed to ‘manage’ sea-level rise. One important implication of a move
in this direction, is a greater acknowledgement of a human-nature partnership
– a sense in which there are not two separate domains of nature and culture,
but rather an active nature/culture which is being allowed to reproduce itself.
Further positive and viable approaches to the question of Non-Indigenous
Species may be found in existing nature/culture partnerships – in the realms
of agriculture, horticulture and viticulture – these realms can be seen as
communities that harbour a complex appreciation of non-indigenous species
in the landscape and that might have lessons to impart to other domains.
Changes in the future that will affect NIS involve both physical and human
dimensions. One of the most important physical factors will be the added
effect of climate change on the human contribution to the movement of
species creating NIS problems. Whether NIS do become a problem will not
only be about considering the behaviour of the organism, but will depend
equally on changes in human activity and changes in human values.
At a national level, there is likely to be an increase in cosmopolitanism and
pluralism. As a social and political phenomenon it is anticipated that this will
have an impact on how NIS are perceived and dealt with in a UK context.
It is anticipated that in the light of a growth in cosmopolitanism, the majority of
the population in 20 years time will have less of an interest in NIS than today.
If diversity is increasingly embraced at a cultural level and the language of NIS
is steered towards non-offensive terminology, a more accepting approach
may well develop regarding NIS in UK society.
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However, the possibility of future dramatic events may result in the opposite
trend occurring- towards an embracing of the local, and a more parochial state
of affairs may be encountered. This would be reflected in, for example,
increased interest in local foods, promotion of what is perceived to be uniquely
local in terms of habitats and natural environments depending on the need to
preserve and maintain local character, nature and environment. A driver for
this scenario could come from an increased political emphasis on
regionalisation, where we might even see the identification of a Welsh,
Scottish and English flora and fauna.
It is difficult to predict scenarios which might prevail in the UK in the future.
The signs, even within the conservation and environmental NGO
communities, might be read as suggesting that a more evolutionary
perspective is likely to develop in the general context of the management of
nature, but alongside certain elements of the more ‘fixed’ preservationist and
conservationist approaches outlined above. This would mean embracing a
deeper understanding of change and process which includes humans as part
of, and subject to, change in nature. Conservationists and environmental
managers are only beginning to experiment and find out more about what this
means in practice, however. Many ecologists and conservationists ( e.g.
Baskin 2002) are concerned that society might lose sight of some of its
important goals – for example, the protection of biodiversity - if such an
approach was to be adopted too readily.
Summary of section 5







There exists a lack of research concerning public perceptions of alien
species in the UK.
Drawing on other studies of public perceptions of environmental
threats, however, we can see that environmental risks and threats are
normally judged by people as risks in context.
Thus public perception of non-indigenous species will partly
encompass the judged competency of authorities (such as DEFRA) in
the context of a particular issue.
Using the right language, engendering trust within different
communities, building on existing networks and respecting those who
have reliable lay experience of invasive species will all be important
components of lay experience and perceptions of particular species.
Of the small number of studies that have been carried out, some
studies (e.g. Green 2002) suggest a polarisation of views amongst UK
publics, such that alien is seen to be ‘bad’ and native is seen to be ‘
good’.
Other studies (e.g. Rodwell’s current research, Footnote 3) tend to
champion a more human definition of nature which implies and
encourages the ability of publics to embrace the ‘new’ and to create a
sense of ‘belonging’ for species that do not necessarily fulfil
native/indigenous criteria.
Recent projects which have encouraged lay engagement and
enrolment in policy problems may provide useful models for NIS
management in the future.
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


Future trends in society seem to suggest that more mobile, shifting and
cultural definitions of nature may increase in prominence and
acceptability as a basis for policy decisions. This would mean
relinquishing some long-held notions of preservation within, some
spheres of environmental management - for example, that of
conservation management.
Whilst some experiments exist to understand what a more flexible
(evolutionary- or process- based) understanding of society’s
relationship with nature might be, there are useful warnings from critics
who remind us not to lose sight of some established and valued
practices – such as the maintenance of biodiversity.
It will be important for policymakers to keep publics ‘on board’ with
future policies regarding NIS, especially if new approaches are taken
up. This will mean absorbing the very contextual and varying ways in
which different publics understand NIS, making sure that these are
adequately considered in the formation of policy decisions, and where
possible, ensuring that those that are directly affected by policy
decisions are involved in the decision making and management
processes.
6.0 Conclusions and Recommendations

Policy makers are doing the right thing in exploring and trying to
understand the historical, social and cultural contexts within which the
Non-Indigenous Species (NIS) debates are being played out.

Particular attention might be paid in policy circles to questioning the
underlying assumptions about what nature is in the UK context, as well
as the further question as to what NIS do in relation to that assumed
nature.

This implies an avoidance of thinking in essentialist terms about nature
as native/non-native, alien etc. as well as being ecologically
problematic, such terms are culturally insensitive.

There are various different ‘publics’ in the UK. These publics are likely
to judge the problem of NIS not solely in ecological terms but in the
context of their own empirical knowledge, as well as their knowledge
about institutions who are supposed to control environmental threats.

Attempts to enrol publics as informants or stewards of NIS may be a
useful way forward. Crucial to the success of such initiatives is to base
any proposed activities within existing value systems and to find ways
of reciprocating effort, so that belonging within such a vigilant
community reaps the right kinds of rewards for its members.

It would appear that the management of environment threats in the
future might well be based on approaches which avoid a static
reification of what nature and nature management is, instead turning
towards more flexible understandings of human-nature partnerships.
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
In the NIS debates, this would translate into a more pragmatic, perhaps
more human-centred (less ecologically centred) definition of NIS and a
correspondingly more flexible approach to their management.

The challenge for policy, should such a scenario unfold, would be to
ensure that certain agreed upon goals – e.g. maintenance of
biodiversity, ecosystem health, human health etc. – were able
simultaneously to be upheld.
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i
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ii
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iii
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(Sycamore – a Case Study in Belonging)
185