Aquaculture in the coastal zone: pressures, interactions and externalities
David Whitmarsh and Gianna Palmieri
CEMARE, University of Portsmouth
1.
Introduction
Aquaculture worldwide has grown rapidly, but there is concern over its environmental
impact and sustainability. Within Europe there are quite marked regional variations in
the intensity of marine aquaculture (Table 1), and the fact that this occurs in areas
where there is already pressure on coastal resources from competing activities
suggests that the scope for conflict is considerable. In this paper, the DPSIR (Driving
forces – Pressure – State – Impact - Response) framework is used to explore the issues
and to suggest policy solutions. In formulating policy it needs to be understood that
marine aquaculture not only creates externalities but is also the ‘victim’ of external
costs generated elsewhere by other activities in the CZ. The next section reviews the
evidence relating to externalities caused by marine environmental disturbance, and
provides the context for the more specific focus on aquaculture which follows. The
contextual discussion is important for two reasons. Firstly, it highlights the fact that
externalities associated with the exploitation of marine and coastal resources typically
arise from multiple activities, amongst which aquaculture is but one. To study
aquaculture in isolation is therefore misleading. Secondly, in order to gauge whether
aquaculture causes environmental damage to the marine environment it is helpful to
have knowledge of what is at risk; in other words, we need to know how valuable
marine resources are and to what extent that value is impacted by fish farming.
Several published studies have attempted to place an economic value the marine
environment, and these will be reviewed.
2.
Externalities caused by marine environmental disturbance: an overview
The marine environment provides goods and services, which support economic
activities and the welfare of individuals directly. These resources include commodities
such as fish and raw materials, and services ranging from nutrient cycling, disturbance
regulation, and biological control, to recreational and cultural services. Many of these
resources are unmarketed, which means that no property rights are assigned over their
use and that there are no markets reflecting their scarcity. This means that users can
make resources scarce for others but this is not reflected in any change in their cost of
access to the resources. This market failure is called an externality. In economic
theory a negative externality is said to occur when the production or consumption
decisions of an economic agent have an impact on the utility or profit of a third party
in an unintended way and the generator of the impact offers no compensation to the
affected party. What happens is that some of the costs of private production or
consumption decisions are ‘external’ to the economic agents making those decisions
and, consequently, are not taken into account in their decision process. Externalities
may affect a production activity by modifying the efficiency of the production process
and consequently its profitability, or affect the satisfaction (utility) of a consumer. In
1
the case of the marine environment, for example, an oil spill may reduce the yield for
fishermen and fish farmers and the enjoyment of the marine landscape by visitors.
In order to take into account the impact on human welfare of changes in the quantity
or quality of environmental assets, it is necessary to assign a monetary value to them.
Many criticisms have arisen among those who consider such an approach unethical,
but this is to miss the point. Monetary valuation is essentially a means of measuring
the social importance of the marine environment in a way that enables choices to be
made. To take an aquaculture example, deciding whether or not to impose (say) an
effluent charge on salmon growers as a way of ‘internalising’ the externalities of
nutrient release logically requires the decision-maker to know what economic costs
such emissions impose on society. If it were shown that nitrate and phosphate release
from fish farms represented a significant external cost to society, the legitimacy of
imposing such a measure on the polluters would be upheld. Conversely, if the external
cost were estimated to be trivial, the legitimacy of imposing an effluent charge would
be called into question. Monetary valuation is thus a way of bringing the environment
into the reckoning of cost-benefit analysis, which may then be used by decisionmakers to evaluate policy options in a rational and consistent manner. This raises the
question of how easy it is in practice to estimate such values, and what progress has
been made in valuing the environment. To date, empirical studies related to the
valuation of marine resources have focused primarily on wetlands or on particular
resources (such as mangrove and coral) adjacent to the coast, and only rarely on ocean
resources. A study by Costanza et al. (1997) on the value of the world’s ecosystem
services and natural capital estimates the value of marine systems to be US$20.9
trillion annually (63% of total ecosystem services and natural capital) within which
coastal systems contribute some 50%. Although these are quite crude estimates and
the authors themselves point out the limitations of their study, what is clear is that
marine resources have an economic value, which can be affected by environmental
disturbance of anthropogenic origin. This is demonstrated more clearly in the case
studies discussed in the Appendix, which look in detail at the estimates of economic
value associated with three types of environmental impact: oil spills, nutrient
enrichment, and ecosystem degradation arising from mangrove and coral reef
destruction.
3. Interactions between aquaculture and the marine environment
Aquaculture is affected by the externalities created by other activities, and is itself a
contributor to such externalities. In this section we look more closely at these
interactions and their socio-economic significance. The key linkages are illustrated in
Figure 1, which is based on the DPSIR framework. We may start by considering the
operating performance of commercial aquaculture in terms of a number of standard
indicators (e.g. productivity, costs, prices and profits), which for a given production
system are determined by the state of technology and the prevailing economic
conditions. If these conditions alter, production from the aquaculture sector will
change as firms respond to market opportunities and pressures.
As Figure 1 suggests, operating performance may be impacted by various external
events. Some of these can be classed as either totally or partially anthropogenic (e.g.
acute or chronic pollution, ‘red tides’ linked with eutrophication), while others are
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caused by natural hazards (e.g. storms). Our concern is with the former, where
experience suggests that such events can be financially ruinous to individual fish
farmers and to the sector as a whole. The Amoco Cadiz oil tanker disaster in 1978
caused serious losses to the Brittany oyster farming industry, estimated at 107 million
FF (= US $26 million) at 1978 equivalent values (Grigalunas, 1986), while the more
recent Prestige incident in 2002/3 is reckoned to have led to a drop in the annual
production value of Galician aquaculture (mussels and turbot) of approximately 9
million euros (Garza-Gil, 2005, in press). In another high-profile case, the Exxon
Valdez oil spill off the coast of Alaska in 1989, published statistics strongly suggest
that aquaculture was adversely affected by the incident. Figure 2 shows that the
annual value of farmed oyster production by the US in the Pacific NW region fell
sharply in 1989, due mainly to very low prices. While there were no doubt other
factors at work, it would be remarkable if the events of that year (e.g. concern over the
seafood edibility) had not had some impact on oyster markets. As well as acute
pollution incidents such as oil spills, chronic pollution has also been shown to be
damaging to aquaculture. This is well illustrated by the effects of organotins on
shellfish, a particularly noteworthy case being the culture of oyster (Crassostrea
gigas) in Arcachon Bay (France) where reproductive failure and shell deformation
caused by the presence of TBT resulted in very heavy financial losses by the shellfish
industry in the early 1980s (Santillo et al., 2001). A final example of how external
events may adversely impact on aquaculture concerns harmful algal blooms (HABs),
which in the US are estimated to have caused losses to commercial fisheries (capture
and culture) of between $13 m. and $18 m. annually (Anderson et al. 2000).
The financial burden of these external events arises in different ways. To start with,
and perhaps most obviously, there may be biological injury (e.g. in the form of
increased mortality or reduced growth) which in the extreme may cause a fish farm to
cease production completely. A more common situation, however, is where a
regulatory body responds to a pollution incident by closing a fishery and banning the
sale of the product. In such cases there may no biological injury per se, but the cost to
the aquaculture producer in the form of lost revenue is nonetheless real. The financial
impact of pollution incidents may also be felt where concern over the edibility of the
product translates into a fall in consumer demand, as happened to Shetland salmon
growers following the Braer oil tanker disaster in 1993. Water quality differences
can impose indirect costs on producers, the best example being the need for shellfish
growers in areas not meeting the Class A standard (< 300 faecal coliforms or 230
E.Coli per 100g) to undertake compensatory investment in order to make their product
saleable. In the UK this includes a requirement that shellfish undergo purification or
relaying, which for Class C areas (< 60,000 faecal coliforms or 46,000 E.Coli per 100
g) must be a period of at least 2 months (Younger and Kershaw, 2004). The final way
in which external events may impact financially on aquaculture is through risk
management costs, most obviously via insurance or else through physical precautions
to reduce the probability of harmful incidents occurring (e.g. siting farms in low risk
areas, vaccination of stock, etc.). Indeed, given that aquaculture is perceived by
underwriters as a very high risk activity (Secretan, 2003, p. 11) , such precautions are
generally a condition for obtaining insurance cover. Indicative figures for the
insurance premia charged for specific risks (‘named perils’) that may cause stock
mortality in aquaculture are given in Table 2.
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Aquaculture may itself create externalities via its impact on the marine environment
(Figure 1), and though it is not always possible to quantify the importance of these
effects in monetary terms there is little doubt that they can often be significant for
human welfare. One of the most clearly demonstrated negative effects of aquaculture
(i.e. external costs) is the degradation or loss of critical marine habitat, which may in
turn lead to a reduction in biodiversity. The best documented example of this is the
conversion of mangrove swamps to shrimp farming and the consequent removal of an
important nursery area for juvenile fish. Barbier and Strand (1998) estimate the effects
of changes in mangrove area on the shrimp fisheries of Campeche State (Mexico)
based on a production function methodology, and using data for the decade 1980-90
find that a 1 km2 decline in mangrove area leads to a loss of 14.4 tonnes of
commercial shrimp harvest worth some US$ 144,000. Though mangrove
deforestation over that period was comparatively small, the authors point out that in
the future this is likely to worsen due to urban expansion and mariculture
development (p. 163). In other words, shrimp aquaculture has the potential to impose
significant external costs on society by indirectly causing a decline on the harvest of
capture fisheries. Habitat degradation caused by mariculture may also result from
farming species other than shrimp, however, and we would point to the evidence
which reveals the growing threat that cage aquaculture poses for seagrass meadows in
the Mediterranean (Holmer et al., 2003). Sedimentation of waste products from fish
farms has been shown to adversely impact the growth of seagrass (Posidonia
oceanica) in the vicinity of sea cages, and given that seagrass is an important source
of food and habitat for several varieties of marine organisms it seems likely that fish
farms located inappropriately close to P. oceanica meadows may have indirect
negative effects on the biodiversity and productive capacity of the marine
environment. This is likely to translate into an external cost, most obviously a
reduction in the harvest of capture fisheries, though the magnitude of this loss is as yet
unknown.
Aquaculture may engender a number of other negative externalities, some of which
may impact on commercial fisheries whilst others may be felt by other users of the
coastal zone (e.g. tourists) or by society at large. There are at least two mechanisms
by which fisheries are affected. To start with, it is argued that growth in demand for
fishmeal used in intensive aquaculture increases pressure on stocks such as anchovy,
sandeel, etc., with potentially adverse effects on the production of edible supplies
from marine capture fisheries (Hannesson, 2003). The problem, however, is at least as
much to do with the way capture fisheries are managed as with the growth of
aquaculture per se. Where fisheries are effectively open-access, the over-exploitation
of stocks caused by increased demand for fishmeal is likely to be more severe. A
second way in which aquaculture may impose external costs on capture fisheries is
through the spread of disease and infestations, a particularly controversial example
being the transmission of sea lice from salmon cages to the wild salmon stocks. While
the evidence for this is still a matter of dispute, if the link were to be established it
would be potentially serious in socio-economic as well as ecological terms. Apart
from its impact on fish stocks, cage aquaculture may affect the coastal environment in
ways which reduce its amenity value, possibly with implications for tourism. One way
is through reductions in water quality as a result of nutrient enrichment and
eutrophication, another is through visual intrusion due to the inappropriate siting of
4
sea cages. The external cost of eutrophication from coastal cage salmon farming in
the early 1990s has been estimated using Swedish data to be between 50 and 100 SEK
(= US$ 6.4 and US$ 12.8) per kg of nitrogen (Folke et al 1994), figures which it is
claimed would make salmon farming unsustainable if added to the cost of production
(i.e. internalised). Though this particular study has been severely criticised for its
methodology (Black et al. 1997), it is worth noting that an economic assessment of
eutrophication in the Baltic based on a Willingness to Pay approach (Gren et al. 2000)
has estimated the external cost of nutrient release (N) from all sources to be 62 SEK
per kg N and hence of a similar order of magnitude to that found by Folke et al. This
finding might appear to support the claim that salmon farming is unsustainable, but it
should be acknowledged that the external costs of nutrient release in a water body
such as the Baltic are likely to be higher than in less eutrophic areas. Indeed, evidence
showing that nutrients released by fish farms situated in oligotrophic waters may
increase productivity in local capture fisheries (Machias et al. 2004) strongly implies
that in such circumstances the externalities of cage aquaculture may be positive rather
than negative.
4. Institutional responses and policy implications
The DPSIR framework illustrated in Figure 1 suggests that the response to the
externalities of aquaculture will come from a variety of institutions, most obviously
from government departments or agencies in the form of controls on fish farms
intended to mitigate or prevent potential environmental damage. The regulations that
currently apply to European aquaculture, which for some countries imposes quite
severe restraints on fish farms, is clear evidence that national governments are not
indifferent to such risks (Fernandes et al. 2000, Read and Fernandes 2003). The
institutional response may also come from other bodies and stakeholder groups, and it
is worth noting that consumers are increasingly being called upon to make ‘informed
choices’ when they purchase seafood according to a set of environmental criteria. As
well as established ecolabelling and certification schemes such as that operated by the
Marine Stewardship Council, information is now available to consumers to enable
them evaluate fisheries products using an environmental ranking system. This
approach has lately been pioneered by the Blue Ocean Institute and applied to farmed
fish sold in the US, with species awarded a score based on their supposed
environmental impact (Table 3). A further type of response may come from
commercial firms linked with the supply side of aquaculture, for example by the
provision of pollution-abatement technologies and inputs that are environmentally less
damaging (e.g. altering the nutritional composition of pelleted feed to include less
fishmeal). These responses may not, however, be sufficient to address the problems
suffered by aquaculture or caused by it. Within the DPSIR framework, the ways
institutions respond is a function of the information that they receive, and if that
information is inadequate or unbalanced the response will misdirected. The messages
emanating from some of the environmental groups concerning the supposed
environmental harm caused by aquaculture testifies to the fact that information in this
area is far from objective, and if such messages are listened to and acted upon they
carry the risk that decisions made by government agencies or consumers may be
inappropriate. These considerations underscore the important role of performance
indicators, chosen and constructed so that unbiased and essential information can be
communicated to policy makers.
5
Table 1: Regional variations in the intensity of coastal aquaculture
Country
Denmark
Germany (inc. ex-GDR)
Greece
Spain
France
Ireland
Italy
Cyprus
Malta
Netherlands
Portugal
Slovenia
Sweden
United Kingdom
Croatia
Turkey
Iceland
Norway
Faroe Islands (DK)
Total value of marine
aquaculture production
(Thous €)
Value of production per km
of coastline (Thous. €)
All fishery
products
All fishery
products
16,166
20,179
304,663
253,148
413,537
95,485
181,946
9,766
3,504
78,042
36,319
533
4,151
418,256
13,971
159,571
9,256
1,183,527
165,997
Finfish
16,147
(No data)
292,205
157,745
40,460
53,862
117,514
8,593
3,504
(No data)
14,534
399
3,292
391,868
12,852
158,606
9,175
1,182,473
165,997
2.2
8.4
22.3
51.0
120.7
65.9
23.9
15.1
17.8
173.0
20.3
11.4
1.3
33.7
2.4
22.2
1.9
47.1
148.6
Finfish
2.2
n/k
21.4
31.8
11.8
37.2
15.5
13.3
17.8
n/k
8.1
8.6
1.0
31.5
2.2
22.0
1.8
47.0
148.6
Note: Figures refer to 2003.
Source: Calculated from data published in Eurostat New Cronos Database and CIA The
World Factbook
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Table 2: Aquaculture insurance rates for offshore production systems
Named peril
Bass / Bream
Salmon / Char
Pollution (as defined below)
0.40
0.50
Malicious acts and theft
0.25
0.20
Predation or physical damage by predators
0.30
0.40
Storm, lightning, structural failure, etc.
0.80
0.90
Freezing, supercooling and ice damage
0.10
0.50
De-oxygenation due to competing biological activity
0.60
0.80
Change in water chemistry (inc. pH and salinity)
0.50
0.50
Disease
0.70
0.75
Notes:
(i)
(ii)
Rates apply to the value of the fish stock, the standard method of assessment
being the costs incurred in growing fish to market size.
Pollution is typically defined as ‘the presence of any foreign substance or
material of a toxic nature that causes mortality or results in a total loss of
market value’. (Secretan, 2003, p. 25)
Source: adapted from Secretan (2003)
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Table 3: Environmental ranking of seafood: the Blue Ocean Institute
Scoring system
Fishery product
Farmed fish sold in the US (selected species)
Species
Environmental score
Catfish
Ictalurus punctatus
2.35
Hard clam
Mercenaria mercenaria
3.05
Blue mussel
Mytilus edulis
3.30
Pacific oyster
Crassostrea gigas
2.80
Rainbow trout
Oncorhynchus mykiss
2.15
Atlantic salmon
Salmo salar
0.95
Bay scallop
Argopecten irradians
2.75
Shrimp (US)
Litopenaeus vannamei
2.30
Shrimp (imported)
Penaeus spp.
0.55
Tilapia
Cichlidae family
2.30
Note: The overall score is based on 5 separate environmental criteria, covering
(i) Inherent operational risks (ii) Feed (iii) Pollution (iv) Risks to other species (v)
Ecological effects. The maximum value of the score on any of these criteria and
overall is 4.0. (See the Blue Ocean Institute website for full details of the scoring
methodology).
Source: http://www2.blueocean.org/
Retrieved September 2005
8
Figure 1: Economy – environment interactions associated with aquaculture
Operating performance
of aquaculture
 Productivity
 Costs
 Prices
 Profits
Controls on fish farms
Consumer demand
Technical innovation
Responding institutions
 Ministries and agencies
 Research organisations
 Environmental bodies
 Consumers
 Feed manufacturers
 Equipment suppliers
Externalaties
External events
 ‘Accidents’ (e.g. oil spills)
 Chemical pollution
 Algal blooms
 Natural hazards (e.g. storms)
State of the marine environment
 Critical marine habitat
 Biodiversity
 Wild fish stocks
 Water quality
Socio-economic impacts
 Competition for ocean space
 Incomes (e.g. fishing, tourism)
 Amenity values
 Edible quality of farmed fish
 Human health
 Supplies from wild fisheries
 Public attitudes
9
Farmed oyster production by the US from
the Pacific NW region - total value
40000
1000
35000
900
800
Price (€ / tonne)
30000
Value (€ x 1000)
Farmed oyster production by the US from the
Pacific NW region - prices
25000
20000
15000
10000
700
600
500
400
300
200
100
5000
0
0
1984
1989
1994
1999
1984
1989
1994
1999
Figure 2: Evidence for the effects of the 1989 Exxon Valdez oil spill incident on aquaculture
10
Appendix: Effects on social welfare of marine environmental
disturbance of anthropogenic origin
A1. Economic values and valuation methodologies
Economics attempts to apply a common measurement unit to all the resources and
services that contribute to social welfare. While many of these goods are traded in
markets that express their scarcity, others are not. A sizeable proportion of the goods
and services provided by the environment are not priced and this leads to their
overexploitation. The absence of market prices also means that there is no simple
metric by which to measure economic value, and it is to overcome this problem that a
number of valuation techniques have been developed. An important starting point is
the concept of total economic value (TEV), which recognizes the fact that the
environment provides not simply direct use values but also a range of indirect and
‘passive’ use values to society. Direct use values arise from the direct consumption of
a resource, for example the coast for recreation, or fish as food. Indirect use values are
benefits that are derived from the environment without human intervention, as in the
case of life support services (e.g. gas regulation function of the open ocean) or
ecological services that are inputs into a process of production (e.g. mangroves as
breeding grounds for fisheries). Passive use values are assigned to a resource for its
mere existence (existence value) or for its availability to future generations (bequest
value), as in the case of endangered animal species such as whales. In order to take
into account the impact on social welfare of changes in the quantity or quality of nonmarketed resources and services, economists make use of a range of techniques,
which have different contexts of application. Several categories of technique may be
identified, differentiated according to the type of market they are based on and
whether they make use of actual or potential behaviour of individuals.
The first category includes techniques that are directly based on market prices or
measures of productivity. These can be used when a change in an environmental
resource or service affects actual production or productive capability. When, for
instance, non-marketed goods are used in production as inputs and such values are
reflected in market prices, information about a marketed good can be used to infer the
value of a non-marketed good. This is termed the production function approach. The
second category of techniques uses data on potential expenditure valued in
conventional markets. They are derived from the costs of providing a proxy for the
resource or service to be evaluated. An example of these is the replacement cost
method. Valuation techniques using implicit markets infer values for an
environmental resource or service from a related or surrogate market. An example is
the travel cost method, which uses travel costs as a proxy for the value of an openaccess recreational site. Finally, when there is no market and not even an implicit
market from which to infer the value of an environmental good, a hypothetical market
can be constructed. This is the approach used by the contingent valuation method.
Through the use of surveys, this technique infers demand for an environmental
resource or service by eliciting individuals’ willingness to pay for an environmental
improvement or willingness to accept compensation for a loss in environmental
quality. Currently, the contingent valuation method is the sole valuation method
capable of capturing passive use values.
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In the following sections a number of case studies will be presented which analyse the
impact on economic activities and other aspects of human welfare of three different
sources of marine environmental disturbance of anthropogenic origin, namely oil
spills, nutrient enrichment and eutrophication, and ecosystem destruction.
A2. Oil spills
Tanker accidents have been the cause of most of the very largest oil spills: of the 66
spills in which at least 10 million gallons (34,000 tonnes) of oil were lost, 48 were
from tankers (NOAA). Oil spills happen all around the world. Spills in the size range
of at least 34,000 tonnes have occurred in the waters of 112 nations since 1960 with
particularly high frequencies in the Gulf of Mexico, the North Atlantic, and the
Mediterranean Sea (NOAA). The effects of oil pollution upon the marine environment
can be seen as a decrease in the flow of services provided by a marine resource, which
can extend over a number of years. Impacts on the marine environment of oil spills
include losses for economic activities, most notably fishing, fish farming and tourism
(EEA 1999), and reductions in other aspects of social welfare, such as non-marketed
recreation losses. In order to shed light on the type and magnitude of these impacts,
the cases of two well documented major oil spills from tankers that took place in the
last decades are discussed below.
A2.1. The Amoco Cadiz incident
The Amoco Cadiz ran aground on 16 March 1978 off the coast of Northern Brittany
in France spilling its cargo of 220,000 tons of oil. Over 350 kilometres of coastline
were contaminated. Given the unavailability of catch and effort data by boat,
Grigalunas et al. (1986) estimated losses to off-shore fisheries using a trendextrapolation model based on monthly estimates of the loss or gain in the months
following the oil spill up to the end of 1979. The model was used to capture the
decline of physical outputs, which were then evaluated at the levels of real prices
predicted to have occurred in the absence of the oil spill. Losses to fisheries were
evaluated at 4.7 1978 million dollars. Much worse was the situation for the Brittany
oyster farming industry which accounted for 15% of French oyster production. Oyster
and shellfish losses included the value of stocks destroyed, the value of lost
production during the period of stock recovery, avoidance and cleanup costs, and the
costs of additional promotional activities aimed at preventing any erosion of France’
‘product image’ in this market. Losses were conservatively estimated at 26 1978
million dollars. As for the Brittany tourism industry, an estimated 245,000 visitors
chose not go to Brittany summer resorts following the oil spill and losses were
estimated at 27.5 1978 million dollars. Using interviews with tour operators and a
range of stated preference techniques (travel cost and contingent valuation methods)
non-marketed recreation losses to tourists and to residents were estimated in the range
of 10-80 1978 million dollars depending on the assumptions at the base of the
estimates.
A2.2 The Exxon Valdez incident
On the night of 24 March 1989 the supertanker Exxon Valdez ran aground on Bligh
Reef in Prince Williams Sound, Alaska, spilling 38,800 metric tons of crude oil.
Although the quantity of oil spilt was relatively small compared to other incidents, the
12
Exxon Valdez case is considered to be the worst ever in terms of environmental
damages as it occurred in an ecologically pristine area. In fact, ecological systems
have not fully recovered to date (Exxon Valdez Oil Spill Trustee Council). The case
of the Exxon Valdez spurred a heated debate about liability for the loss of passive use
values and their measurement (see Portney 1994). Lost passive use values were
evaluated through a contingent valuation survey which described the damages caused
by the oil spill and established a referendum market for eliciting the value respondents
placed on preventing a future accident that would cause an equivalent amount of
damage in the Prince William Sound area. Lost passive use were initially estimated at
2.8 1990 billion dollars but in 2003 a new valuation by the same authors estimated
them at 4.87 billion dollars (Carson et al. 2003). Losses to recreational fishing for the
period 1989-1990 were estimated by Carson and Hanemann (1992) using the travel
cost method. These were found to be in the range of 3.6-50.5 million dollars
depending on the assumptions at the base of the estimates. Losses to salmon fisheries
were estimated by Cohen (1995) using a method similar to that used in the Amoco
Cadiz case (see Section A2.1.). These amounted to as much as 155.1 million dollars
for the years 1989 and 1990. However, these estimates should be taken with care
since Alaska experienced several other biological and economic perturbations in those
years that might have contributed to those losses.
A3. Nutrient enrichment and eutrophication
Many human activities use marine ecosystems as free of charge waste sinks for their
by products. Runoff from agriculture, discharges from septic systems and sewers, and
other human-related activities increase the flux of substances such as nutrients
(especially nitrogen and phosphorus) into marine ecosystems. Nutrient enrichment
can favour eutrophic conditions which create negative externalities and impact on
social welfare decreasing the value of the marine environment such that recreation,
fishing, and aesthetic enjoyment may be hindered. However, nutrient enrichment may
lead to positive externalities under certain conditions, as in the case of nutrient-poor
marine ecosystems. For instance, this has been the case in the Mediterrenean Sea,
where nutrient enrichment has led to an increase in fishery productivity (Caddy et al.
1995, Machias et al. 2004). Moreover, even events such as algal blooms, which are
favoured by eutrophic conditions and have caused considerable losses in human
welfare (see, for instance, Wessels et al. 1995), have not a purely anthropogenic
origin. In order to shed light on some of these issue, two case studies are discussed,
namely the case of the Baltic Sea and the case of the Black Sea. These are water
bodies that, due to their peculiar characteristics, are especially susceptible to pollution
from nutrients.
A3.1. The Baltic Sea case
The Baltic Sea environment and resources are particularly vulnerable due to a
combination of biophysical and socio-economic characteristics. The Baltic Sea is the
largest brackish body of water in the world and it is characterised by a very slow
water exchange. The catchment area covers around 1,670,000 square kilometres and
contains a population of about 85 million people. A significant proportion of the
world’s industrial production comes from the countries surrounding it. Current loads
of nitrogen and phosphorous are high. The agriculture sector and sewage treatment
plants are the major sources of nutrient discharges with the former accounting for
13
50% and 33% of nitrogen and phosphorus loads respectively, and the latter for 33%
and 66% respectively (Gren 2000). The Baltic Sea coasts are popular tourist areas.
High nutrients loads cause problems such as water turbidity and algal blooms, which
decrease the recreational value of beaches and archipelagos. The Baltic Sea waters
also serve as spawning, nursery, and feeding grounds for several species of marine
and freshwater fish. Turner et al. (1999) estimated the economic value of
eutrophication damage due to changes in the stream of a range of services provided by
the Baltic Sea. Using the contingent valuation method they estimated the total
economic value of reducing the effects of eutrophication, as well as sub-components
of this total value, such as use (e.g. beach recreation) and passive use (e.g. existence
of preserved species) values. Although the estimate provided (SEK 37,892-69,310
million per year) is highly uncertain due to difficulties in the estimation process, it is
still clear that the value of the services provided by the Baltic Sea is high. The effect
of eutrophication on fisheries appears mixed. While total fish catch has increased
tenfold since the 1940s and eutrophication has been ascribed as one of the causes of
this increase, some species such as cod have been negatively affected by
eutrophication (Gren et al. 1997).
A3.2. The Black Sea case
Like the Baltic Sea, the Black Sea also is very vulnerable to pressures from human
activities. This water body is almost completely enclosed and has shallow waters. It
receives the drainage from a 2 million square kilometres basin covering about a third
of continental Europe (17 countries with a population of around 160 million people).
Discharges of nitrogen and phosphorous are high, largely as a result of agricultural,
domestic, and industrial activities (BSEP 1994). In the case of the Black Sea it
appears that nutrient enrichment may yield some positive welfare effects under certain
assumptions. Knowler et al. (2001, 2002) attempted to model and to empirically
estimate the welfare effects of changes in the level of phosphates on the Black Sea
anchovy fishery in order to assess the desirability of nutrient enrichment programs.
Knowler and Barbier (2000) estimated marginal values of phosphates as an
environmental input in fish production from $45,000 to $713,000 (in US$ 1989/90)
per μM (phosphate unit). The latter value refers to an undisturbed ecological state,
while the former refers to a disturbed ecological state due to the introduction in the
1980’s (likely through ballast dumping, another source of negative externalities in
marine environments) of an invasive species of comb-jelly (Mnemiopsis leidyi), which
both preys and competes with anchovy. In order to evaluate the overall impact on
social welfare of the Black Sea nutrient enrichment, the above figures should be
compared to other values of social welfare that are likely to be negatively affected by
nutrient enrichment (e.g. other fish species, water quality, recreation and tourism).
The authors also tried to estimate the effects of nutrient enrichment on the anchovy
fishery under the assumption that Mnemiopsis leidyi outbreaks are favoured by
nutrient enrichment. In this case results were ambiguous.
A4. Ecosystem destruction
Important ecological systems such as mangroves and coral reefs are under threat in
many countries worldwide, especially in developing countries where these systems
are threatened by the expansion of activities such as agriculture, aquaculture, tourism,
and urban and infrastructural development. The root of this problem is the failure to
14
take into consideration in development decisions the considerable value of the
ecological functions served by these ecosystems. Economic causes and magnitude of
losses of mangrove conversion and coral reefs destruction are discussed below.
A4.1 Mangrove conversion
Mangroves line the tropical and sub-tropical coastlines of approximately 117
countries worldwide (WCMC 1994). They serve important ecological functions such
as the protection of coastlines from sea erosion and hurricanes, and the provision of
nutrients and shelter to many aquatic organisms. They also provide local communities
with commercial and subsistence commodities. Mangroves are disappearing at
alarming rates in many countries. The main factors contributing to this process are
increasing population pressures in coastal areas, overharvesting of timber, and in more
recent years the demand for land by primary sector activities such as mining,
conversion to salt ponds, and agriculture and aquaculture expansion. The disruption of
the ecological and economic functions performed by mangroves has resulted in many
countries in fishery decline, degradation of water supplies, salinisation of coastal
soils, erosion, land subsidence, and the release of carbon dioxide in the atmosphere.
There are few studies on the economic causes of mangrove deforestation. Barbier and
Cox (2003) developed a cross-country analysis (89 countries were included) of the
extent to which economic development influences mangrove loss worldwide. They
concluded that increases in agricultural GDP per person employed have the largest
impact on mangrove loss: a 10% rise resulted in a 10.5%-10.9% loss in mangrove
area. Shrimp aquaculture was also linked with mangrove loss but less severely: a 10%
increase in shrimp aquaculture production was associated with a 0.14%-0.18% loss in
mangrove area. However, losses due to shrimp aquaculture were severe in some major
shrimp-exporting countries. For instance, the same authors (Barbier and Cox 2004)
analysed the factors determining mangrove conversion in Thailand’s coastal provinces
during the period 1979-1996. They found shrimp aquaculture to be an important cause
of mangrove deforestation but also found evidence of other activities such as
agriculture, tourism, industrialisation, and urbanisation contributing to mangrove
conversion. As for the valuation of mangrove services and resources, a few studies
have attempted this analysis, in particular the impact on fisheries of the loss of
breeding and nursery grounds provided by mangroves. Barbier and Strand (1998)
adopted a production function approach in order to evaluate the effect of mangrove
loss on the production and value of the wild shrimp fishery in Campeche State,
Mexico. Results indicated that mangroves were an important input into shrimp fishery
but that the low levels of deforestation during the period under consideration (19801990) implied relatively small losses (US$ 278,704, equivalent to a reduction of 0.4%
in the revenue of the wild shrimp fishery). More substantial losses are to be expected
when mangrove deforestation is high. In Thailand, for instance, losses to the artisanal
demersal and shellfish fisheries due to mangrove losses were estimated to range
between $1.3 to $1.6 million annually during the period 1983-1996 (Barbier 2003).
Values for several mangrove resources and services were estimated by Sathirathai and
Barbier (2001) for a small local community in Southern Thailand (652 residents).
Using field surveys the value of direct use of wood as well as other resources
collected from the mangroves was estimated to be $88 per ha. Benefits in terms of
off-shore fishery linkages were estimated using a production function approach and
were in the range of $21-$69 per ha. The value of coastline protection was estimated
15
using a replacement cost method (based on the cost of replacing mangroves with
breakwaters to prevent erosion) and amounted to $3,679 per ha. Over a 20 years
period, the total present value of the mangrove system to the local community was
estimated to be as much as $27,264-$35,921.
A4.2 Coral reef destruction
Coral reefs are highly productive and biologically diverse ecosystems covering only
0.2% of the ocean floor, yet supporting an estimated 25% of all marine life (Spalding
2001). They provide a range of resources such as seafood, pharmaceutical substances,
and goods traded in the jewellery and souvenir markets, as well as providing
recreational and cultural benefits to a vast number of individuals. They also serve
important ecological services such as the protection of coastlines from sea erosion and
the provision of spawning, breeding and feeding grounds for a multitude of
organisms. Many coral reefs are threatened worldwide. Several causes contribute to
the decline of coral reefs including urbanisation, deforestation, intensive agriculture,
overharvesting of reef resources, destructive fishing methods, and uncontrolled
tourism (Moberg and Folke 1999). Only some of the values associated with coral reef
resources and services have been estimated (Moberg and Folke 1999, Spurgeon
1992). Valuation studies have predominantly focused on the impact of coral reef
degradation on economic activities such as tourism and fisheries. More rarely, other
values were estimated. An example is the study conducted by Cesar (1996) who used
market-based techniques to estimate the welfare losses caused by a range of activities
damaging to coral reefs in Indonesia. For instance, on a 25-year time horizon and
applying a 10% discount rate, blast fishing was found to cause losses to fisheries and
to tourism for US$86.3 6 and US$2.9-US$481.9 per km2 respectively, and to impair
the coastal protection function of coral reefs for a value of US$8.9-US$193 per km2.
Losses due to coral mining activities were found to be US$93.6 and US$2.9US$481.9 per km2 for fisheries and for tourism respectively, and the coastal
protection function of coral reefs was impaired by this activity by US$12-US$260 per
km2. Sedimentation from logging activities was found to cause losses to fisheries and
to tourism for US$81 and US$192 per km2 respectively.
16
References
Anderson DM, Hoagland P, Kaoru Y, and White AW (2000). Estimated annual
economic impacts from harmful algal blooms (HABs) in the United States. Woods
Hole Technical Report WHOI-2000-11
Barbier EB and Cox M (2004). An economic analysis of shrimp farm expansion and
mangrove conversion in Thailand. Land Economics, 80: 389-407
Barbier EB (2003). Habitat-fishery linkages and mangrove loss in Thailand.
Contemporary Economic Policy, 21: 59-77
Barbier EB and Cox M (2003). Does economic development lead to mangrove loss?
A cross-country analysis. Contemporary Economic Policy, 21: 418-432
Barbier EB and Strand I (1998). Valuing mangrove-fishery linkages. A case study of
Campeche, Mexico. Environmental and Resource Economics, 12: 151-166
Black E, Gowen R., Rosenthal H, Roth E, Stechy D and Taylor F. (1997). The costs
of eutrophication from salmon farming: implications for policy - a comment.
Journal of Environmental Management 50: 105-109
Black Sea Environmental Program (BSEP) (1994). Saving the Black Sea. Official
Newsletter of the GEF, Issue No.1
Caddy JF, Refk R, and Do-Chi T (1995). Productivity estimates for the
Mediterranean: Evidence of accelerating ecological change. Ocean and Coastal
Management, 26: 1-18
Carson RT, Mitchell RC, Hanemann M, Kopp RJ, Presser S, and Ruud PA (2003).
Contingent Valuation and lost passive use: Damages from the Exxon Valdez oil spill.
Environmental and Resource Economics, 25: 257-286
Carson RT and Hanemann WM (1992). A preliminary economic analysis of
recreational fishing losses related to the Exxon Valdez oil spill. A report to the
Attorney General of the State of Alaska. http://www.evostc.state.ak.us/pdf/econ1.pdf
Cesar H (1996). Economic analysis of Indonesian coral reefs. Environment
Department Working Papers, World Bank, Washington DC
Cohen MJ (1995). Technological disasters and natural resource damage assessment:
An evaluation of the Exxon Valdez oil spill. Land Economics, 71: 65-82
Costanza R, D’Arge R, de Groot R, Farber S, Grasso M, Hannon B, Limburg K,
Naem S, O’Neill V, Paruelo J, Raskin RG, Sutton P andVan Belt M. (1997). The
value of the World’s ecosystem services and natural capital. Nature 387: 253-260
European Environment Agency (EEA) (1999). State and pressures of the marine and
coastal Mediterranean environment. Environmental Issues Series, No. 5, Copenhagen
17
Exxon
Valdez
Oil
Spill
http://www.evostc.state.ak.us
Trustee
Council
(accessed
10/10/2005)
Fernandes TF, Miller KL, Read PA. 2000. Monitoring and regulation of marine
aquaculture in Europe. Journal of Applied Icthyology 16: 138 – 143
Folke C, Kautsky N. and Troell M. (1994). The costs of eutrophication from salmon
farming: implications for policy. Journal of Environmental Management 40: 173182
Garza-Gil MD, Suris-Regueiro JC and Varela-Lafuente MM. (2005, in press).
Assessment of economic damages from the Prestige oil spill. Marine Policy, in press
Gren I-M (2000). Land use, population and nutrient loads. In: Gren I-G, Turner K,
and Wulff F (eds.) Managing a sea. The ecological economics of the Baltic. Earthscan
Publications, London, UK
Gren I-M, Söderqvist T, and Wulff F (1997). Nutrient reductions to the Baltic Sea:
Ecology, costs and benefits. Journal of Environmental Management, 51: 123-143
Grigalunas TA, Anderson RC, Brown GM Jr, Congar R, Meade NF, and Sorensen PE
(1986). Estimating the cost of oil spills: lessons from the Amoco Cadiz incident.
Marine Resource Economics, 2: 239-262
Hannesson R. (2003). Aquaculture and fisheries. Marine Policy 27: 169-178
Holmer M, Perez M and Duarte CM. (2003). Benthic primary producers – a
neglected environmental problem in Mediterranean maricultures ? Marine Pollution
Bulletin 46: 1372 – 1376
Knowler D, Barbier EB, Strand I (2002). An open-access model of fisheries and
nutrient enrichment in the Black Sea: Errata. Marine Resource Economics, 17: 347349
Knowler D, Barbier EB, Strand I (2001). An open-access model of fisheries and
nutrient enrichment in the Black Sea. Marine Resource Economics, 16: 195-217
Knowler D and Barbier EB (2000). Nutrient enrichment and marine ecosystem
disturbance: A deterministic and stochastic analysis. In: Proceedings of the 10th
Biennial Conference of the International Institute of Fisheries Economics and Trade
(IIFET), Corvallis, Oregon, 10-14 July 2000. Available on-line at
http://osu.orst.edu/Dept/IIFET2000/papers/knowler.pdf
Machias A, Karakassis I, Labropoulou M, Somarakis S, Papadopoulou KN, and
Papaconstantinou C (2004). Change in wild fish assemblages after the establishment
of a fish farming zone in an oligotrophic marine ecosystem. Estuarine, Coastal and
Shelf Science, 60: 771-779
Moberg F and Folke K (1999). Ecological goods and services and coral reef
ecosystems. Ecological Economics, 29: 215-233
18
NOAA (accessed 26/10/2006) http://response.restoration.noaa.gov
Portney (1994). The contingent valuation debate: Why economists should care.
Journal of Economic Perspectives, 8: 3-17
Read P, Fernandes T. (2003). Management of environmental impacts of marine
aquaculture in Europe. Aquaculture 226: 139-163
Santillo D, Johnston P and Langston WJ. (2001) Tributyltin (TBT) antifoulants: a
tale of ships, snails and imposex,. Chapter 13 in P Harremoës, D Gee, M MacGarvin,
A Stirling, J Keys, B Wynne and SG Vaz, eds. Late Lessons from Early Warnings:
the Precautionary Principle 1896-2000, European Environment Agency
Environmental Issue Report No 22, Office for Official Publications of the European
Communities, Luxembourg, ISBN 92-9167-323-4, 2001: 135-148.
Sathirathai S and Barbier EB (2001). Valuing mangrove conservation in Southern
Thailand. Contemporary Economic Policy, 19: 109-122
Secretan PAD (2003). The availability of aquaculture crop (stock mortality)
insurance. Document published by Aquaculture Underwriting and Management
Services.
Spalding MD, Ravilious C, and Green EP (2001). World Atlas of Coral Reefs.
Prepared by the UNEP-World Conservation Monitoring Centre. University of
California Press, Berkeley, USA
Spurgeon J (1992). The economic valuation of coral reefs. Marine Pollution Bulletin,
24: 529-536
Turner RK, Georgiou S, Gren I-M, Wulff F, Barrett S, Söderqvist, Bateman I, Folke
C, Langaas S, Zylicz T, Mäler K-G, and Markowska A (1999). Managing nutrient
fluxes and pollution in the Baltic: An interdisciplinary study. Ecological Economics,
30: 333-352
Wessells CR, Miller CJ, and Brooks PM (1995). Toxic algae contamination and
demand for shellfish: A case study of demand for Mussels in Montreal. Marine
Resource Economics, 10: 143-159
World Conservation Monitoring Centre (WCMC) (1994). Biodiversity data
Sourcebook. Cambridge: World Conservation Monitoring Centre
Younger A and Kershaw S. (2004). Sewage discharge improvements in the Solent
and their effect on the shellfishery. CEFAS Shellfish News No. 17 (May)
19