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Title: Overview of integrative tools and methods in assessing ecological integrity in
estuarine and coastal systems worldwide
Authors: Angel Borja1*, Suzanne B. Bricker2, Daniel M. Dauer3, Nicolette T.
Demetriades4, João G. Ferreira5, Anthony T. Forbes4, Pat Hutchings6, Richard
Kenchington7, João Carlos Marques8, Changbo Zhu9
Affiliations:
1
AZTI-Tecnalia; Marine Research Division; Herrera Kaia, Portualdea s/n; 20110
Pasaia; Spain, aborja@pas.azti.es
* Corresponding author
2
NOAA—National Ocean Service, National Centers for Coastal Ocean Science,
1305
East
West
Highway,
Silver
Spring,
MD
20910,
USA;
suzanne.bricker@noaa.gov
3
Department of Biological Sciences; Old Dominion University; Norfolk, VA
23529; USA, ddauer@odu.edu
4
Marine & Estuarine Research; P.O. Box 417, Hyper by the Sea, 4053, South
Africa; ticky@mer.co.za
5
IMAR - Institute of Marine Research, Centre for Ocean and Environment, FCTUNL, Portugal; joao@hoomi.com
6
Marine Invertebrates, Australian Museum, Sydney NSW 2010 Australia,
Pat.Hutchings@austmus.gov.au
7
Australian National Centre for Ocean Resources and Security, University of
Wollongong, NSW 2522, Australia, richard .kenchington@netspeed.com.au
8
IMAR - Institute of Marine Research, University of Coimbra, Portugal,
jcmimar@ci.uc.pt
9
South China Sea Fisheries Research Institute, CAFS, P. R. China.
changbo@ecowin.org
Abstract
In recent years, several legislations worldwide (Oceans Act in USA, Australia or
Canada; Water Framework Directive or Marine Strategy in Europe, National Water
Act in South Africa, etc.) are intending to assess the ecological quality or integrity,
within estuarine and coastal systems. Most of these legislations try to define quality
in an integrative way, by using several biological elements, together with physicochemical and pollution elements. This approach will allow assessing the ecological
status at the ecosystem level (‘ecosystem approach’ or ‘holistic approach’
methodologies), better than at species level (i.e. the mussel biomonitoring or Mussel
Watch) or just at chemical level (i.e. quality objectives) alone.
Increasing attention has been paid to the development of tools for different physicochemical or biological (phytoplankton, zooplankton, benthos, algae, phanerogams,
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fishes) elements of the ecosystems. However, there are very few methodologies
integrating all the elements into a unique evaluation of a water body. The need of
these integrative tools to assess the ecosystem quality is very important, both in
terms of scientific and stakeholder point of view. Politicians and managers need to
get simple, pragmatic, but scientifically sound methodologies, suitable to show to
the society the evolution of a zone (estuary, coastal area, etc.), taking into account
human pressures or recovery processes.
These approaches include: (i) multidisciplinarity, in the teams involved in their
implementation; (ii) integration of biotic and abiotic factors; (iii) accurate and
validated methods in determining ecological integrity; and (iv) adequate indicators
to follow the evolution of the monitored ecosystems.
While countries are increasingly using the establishment of marine parks as ways in
which to conserve marine biodiversity, there is awareness at least in Australia based
on terrestrial systems that conservation and management of biodiversity cannot be
restricted to MPA’s but must include areas outside such reserves. An alternative
approach which has been adopted in some areas of Australia is the declaration of
multi-use parks which allow activities such as commercial shipping and fisheries
although regulated.
This contribution reviews the current situation of the integrative ecological
assessment worldwide, by presenting several examples from each of the continents:
Africa, Asia, Australia, Europe and North America.
Key Words:
Ecological integrity, integrative assessment, ecological status, environmental
quality, marine parks
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Introduction
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The marine environment presents high levels of complexity, diverse habitats
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and supports a high level of biodiversity, providing goods and services together with
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different uses which should be undertaken in a sustainable way. However, the
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marine, and particularly estuarine, environment is facing increasing and significant
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impacts, which include physical and chemical transformation and changes in
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biodiversity. Causes include land reclamation, dredging and sediment discharges,
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pollution (hazardous substances, litter, oil-spills, eutrophication, etc.), unsustainable
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exploitation of marine resources (sand extraction, oil and gas exploitation, fishing,
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etc.), unmanaged tourism, introduction of alien species, and climate change. These
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are driven by economic and social pressures for development and access to marine
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resources and activities through commercial fishing, aquaculture, tourism,
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recreation, maritime transport, etc.
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To face these problems, policy-makers world-wide tend to develop strategies
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to protect, conserve and manage the marine environment. Hence, the United Nations
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Convention on Law of the Sea (UNCLOS, 1982) is the international basic legal
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framework that governs the uses of the oceans and seas. UNCLOS establishes an
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obligation to protect and use the resources of the marine environment sustainably as
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does the 1992 Convention on Biological Diversity (CBD, 2000), as highlighted by
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Parsons (2005).
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At a national or regional level, several initiatives have been developed
recently:
(i) in
December 1998, Australia released an Oceans
Policy
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(Commonwealth of Australia, 1999); (ii) the Canadian Parliament passed the
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Oceans Act, which came into force in January 1997, being the Canada's Oceans
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Strategy released in 2002 (Parsons, 2005); (iii) in the USA, the Pew Oceans
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Commission, created in 2000, and the US Commission on Ocean Policy, created by
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the Oceans Act of 2000, reported in 2004 (Granek et al., 2005); (iv) in Europe, the
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Water Framework Directive (WFD), which promotes the protection of continental,
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estuarine and marine waters, was released in 2000 (Borja, 2005), and the European
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Marine Strategy (EMS) Directive, was presented in 2005 (COM, 2005a, 2005b, and
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2005c; Borja, 2006); (v) in South Africa the National Water Act of 1998
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(www.dwaf.gov.za/documents/publications)
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Management Act are presently in the form of the Integrated Coastal Management
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Bill (www.deat.gov.za); and (vi) in the People’s Republic of China (PRC) a
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substantial body of legislation exists to address environmental protection (laws on
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Water (1988/01/21) and Environmental Protection (1989/12/26): Sea Water Quality
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GB 3097-1997, Environmental Quality for Surface Water GB 3838-2002, and
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Provisions for Monitoring of Marine Culture and Propagation Areas (2002/04/01)).
and
the
developing
Coastal
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The objectives of these initiatives are to protect and/or restore the
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corresponding seas, ensuring that human activities are carried out in a sustainable
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manner, providing safe, clean, healthy and productive marine waters. In summary,
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they try to promote the sustainable use of the seas and conserve marine ecosystems.
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Hence, the main objective of these legislations is to achieve marine waters in a good
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environmental or ecological status. Actually, the concept of environmental status
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takes into account the structure, function and processes of the marine ecosystems
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together with natural physiographic, geographic and climatic factors, as well as
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physical and chemical conditions including those resulting from human activities in
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the area concerned.
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Hence, this concept defines quality in an integrative way, by using several
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biological elements, together with physico-chemical and pollution elements. This
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approach will allow assessing the ecological status at the ecosystem level
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(‘ecosystem-based approach’ (EBA) or ‘holistic approach’ methodologies
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(Browman et al., 2004; Nicholson and Jennings, 2004; and Rudd, 2004)), better than
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at species level (i.e. the mussel biomonitoring or Mussel Watch) or just at chemical
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level (i.e. quality objectives) alone. The EBA is defined as: "a strategy for the
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integrated management of land, water and living resources that promotes
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conservation and sustainable use in an equitable way. The application of the EBA
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will help to reach a balance of the conservation, sustainable use, and the fair and
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equitable sharing of the benefits arising out of the utilization of genetic resources”
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(CBD, 2000).
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Following this approach, increasing attention has been paid to the
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development of tools for different physico-chemical or biological (phytoplankton,
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zooplankton, benthos, algae, phanerogams, and fishes) elements of the ecosystems.
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However, there are very few methodologies integrating all the elements into a
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unique evaluation of status and performance of an aquatic system. The need of these
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integrative tools to assess the ecosystem quality is very important, both in terms of
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scientific and stakeholder point of view. The scientific challenge is to develop
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robust simple, pragmatic, but scientifically sound methodologies, which can provide
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communities and decision-makers with tools to define and monitor the evolution,
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current condition and biological performance of marine ecosystems and bioregions.
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These approaches include: (i) multidisciplinarity, in the teams involved in
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their implementation; (ii) integration of biotic and abiotic factors; (iii) accurate and
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validated methods in determining ecological integrity; (iv) accurate and validated
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methods for determining the extent and effect of human uses and impacts; (v)
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adequate indicators to follow the evolution of the monitored ecosystems; and (vi)
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use of protected areas as means of conserving and managing marine environments
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especially coastal areas where the greatest anthropogenic inputs occur.
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The aim of this contribution is to review the current situation of the
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integrative ecological assessment worldwide, by presenting several examples from
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each of the continents: Africa, Asia, Australia, Europe, and North America.
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Current situation in Africa
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Legislative framework
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South Africa is a dry country with an annual national average rainfall of less
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than 500 mm, well below the global average, with expectations of declines
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associated with climate change over the next century. Reviews of environmental
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legislation in South Africa, particularly relating to aquatic resources and the coastal
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zone, are associated with reviews by the Council for the Environment (1989, 1991)
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which began proposing policies for coastal zone management. The status of coastal
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management was subsequently reviewed by Sowman (1993) and this preceded the
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development of a green paper (Department of Environmental Affairs and Tourism,
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1998) focussed on sustainable coastal development, followed by a white paper
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(Department of Environmental Affairs and Tourism, 2000) on the same topic. The
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appearance of these policy documents was paralleled by publications by Glazewski
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(1997) and Glavovic (2000a, 2000b) aimed at converting the policies and concerns
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articulated in the white paper into an integrated coastal management bill and
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ultimately a national Coastal Management Act (www.deat.gov.za) which is
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presently in the process of ratification.
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On the aquatic resources side, which includes the freshwater, estuarine and
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marine environments, the National Environmental Management Act of 199 was
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superceded/complemented by the National Water Act of 1998 (www.dwaf.gov.za/
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documents/publications). This new act represented a radical digression from the
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philosophy inherent in the historical approach to the management of aquatic
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resources in that aquatic environments, particularly fresh water and estuarine
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systems, were granted a legal persona in that the dependence of the functionality of
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these systems on a minimal level of freshwater flow was given a legal status which
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had to be taken into account when any water abstraction was contemplated. The
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arguable premise that aquatic systems, such as wetlands, rivers and estuaries are
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ultimately dependent on minimal levels of freshwater availability, beyond which
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their functionality will be impaired, clearly generates the question as to what this
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level is and how it might be established. This aspect will be dealt with in the next
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section.
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Tools and methodologies used in assessing ecological integrity
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In the present context the emphasis will be on the determination of the
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freshwater requirements of estuaries (DWAF, 1999): Resource Directed Measures
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for Protection of Water, henceforth referred to as the “reserve” although the
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procedure is described as part of a package dealing also with reserve requirements
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of rivers and wetlands (www.dwaf.gov.za/documents/policies/wrpp).
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Some examples of integrative assessment
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An example of the above where an assessment of the current status was
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followed by remedial action and the institution of a monitoring system to check on
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the effectiveness of the measures instituted is provided by the Mhlanga estuary (290
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42’S; 310 6’E) on the northern outskirts of the city of Durban on the east coast of
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South Africa. This small system with an estuarine area of barely 12 ha (Begg, 1978)
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is nevertheless typical of many of the 73 systems which occur along the 570 km of
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the KwaZulu-Natal coastline and further south into the Eastern Cape Province. The
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major physical and chemical features of these systems are determined by the
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seasonal rainfall, and consequently variable river flow, coupled with strong wave
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action and longshore sand transport which typically result in the closure of these
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systems during winter low flow periods. Under these conditions tidal action is lost
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and with it any organisms dependent on an intertidal habitat. Salinities typically fall
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due to sustained low levels of fresh water input and outward seepage through the
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bar, but layering may develop if the bar is low enough for overwash to occur during
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high wave conditions. Water levels behind the bar will rise, depending on the height
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of the bar, and can result in substantial backflooding such that the overall extent of
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the aquatic environment, in terms of water column and benthic habitat, increases
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well beyond that associated with high tides during open mouth periods. Under
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natural conditions this bar would be naturally breached during summer high flow
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periods but historically (Begg, 1984) this pattern has been disrupted by artificial
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breaching to prevent flooding of cultivated land or infrastructure in the backflooded
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areas.
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The Mhlanga estuary has over the last 25 years become one of the better
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known of the smaller KwaZulu-Natal systems by virtue of i.a., general surveys of
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the system carried out in 1980-1981 (Begg, 1984) and a coincident more intensive
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focus on the fish fauna which produced information on trophic relationships within
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the fish community (Whitfield, 1980a), distribution in relation to food resources
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(Whitfield, 1980b) and factors affecting the recruitment of juveniles into the estuary
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(Whitfield, 1980b). Harrison et al. (2000) produced a nationwide assessment of the
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state of South African estuaries based on the geomorphology, ichthyofauna, water
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quality and aesthetics. The latter three parameters were rated on a scale of poor,
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moderate or good. The fish fauna was assessed on the basis of species richness and
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community composition, the water quality on suitability for aquatic life in terms of
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dissolved oxygen, ammonia, faecal coliforms, nitrate nitrogen and ortho-phosphate
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and the aesthetics on a “visual appraisal of the state of development in and around
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the estuary” incorporating i.a. any type of anthropogenic influence, algal blooms,
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odours, noise or invasive plants. The fish fauna and aesthetics of the Mhlanga
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estuary were rated as good but the water quality as poor. The poor water quality
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reflects the vulnerability of these small systems during the closed mouth periods
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when water exchange is minimal and tidal effects non-existent. In 2002-2003 a
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multi-disciplinary
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conditions, nutrient conditions, phytoplankton and microphytobenthos, zooplankton,
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benthos, fish and birds, sponsored by the South African Water research Commission
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was undertaken (Perissinotto et al., 2004) aimed at contributing to the
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implementation of measures for reserve determinations for estuaries. In the local
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context the project focussed on the “responses of the biological communities to flow
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variation and mouth state in temporarily open/closed estuaries” one of which was
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the Mhlanga.
study
incorporating
mouth
dynamics,
physico-chemical
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In summary, the study supported perceptions and interpretations developed
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some 20 years earlier (Begg, 1984), viz. that the broad natural cycle of summer
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breaching and winter closure due to the seasonal rainfall pattern was a major driving
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force in the functioning of these temporarily open/closed systems. Although these
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estuaries became non-tidal and salinities dropped to virtual freshwater levels during
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closed periods, with a consequent limiting effect on benthic invertebrate diversity,
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the fish fauna, which tended to consist largely of juveniles recruited to these nursery
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grounds during open mouth periods or through overwash, appeared able to handle
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these low salinities. Retention and accumulation of water behind the bar also
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resulted in an expanded aquatic and benthic environment relative to that existing
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under high tide conditions. The increased and stable water column permitted the
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development of a phytoplankton and in turn the development of a zooplankton and a
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planktivorous fish fauna while the benthos was able to expand in abundance
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although not in diversity. Optimisation of these processes was dependent on regular
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seasonal cycles of breaching or overtopping, allowing fish or invertebrate migration,
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followed by periods of closure which allowed the accumulation of biomass, both
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plant and animal, before the next exchange. Disruption of this cycle by artificial
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breaching and draining of the estuary during winter when water levels normally
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peak would disrupt this cycle. Additional impacts would be imposed by nutrient
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inputs resulting from agricultural runoff or urban pollution causing algal blooms,
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eutrophication and oxygen depletion.
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In the Mhlanga situation records of mouth behaviour coupled with historical
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observations (Begg, 1978, 1984), calculations of the pristine mean annual and
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monthly runoff and the present situation indicated that outflow of treated water from
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a sewage works situated upstream of the estuary significantly increased the total
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flow into the estuary and the frequency of mouth breaching, resulting in the type of
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impacts described above. At low input levels the variable quality of the treated
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effluent was such as to generate periodic localised low or anoxic conditions
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resulting in fish kills. The increase in water inflow into the river from the sewage
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works resulted from the fact that the water used in the catchment was derived from
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other catchments and resulted in an overall increase in the Mhlanga flow. In this
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situation then the impacts on the estuary arose from the rather unusual situation of
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excess flow rather than the more common problems arising from water abstraction.
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The provisions of the reserve determinations allow for either the
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maintenance of an existing acceptable ecological status or the implementation of
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measures to improve the ecological status of an estuary. In this case the measures
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that have been implemented by the local municipality involve the installation of a
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pipeline to transport the excess water to an adjacent catchment which has been
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subject to significant abstraction as well as improved treatment of the waste water
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from the sewage works. A monitoring operation has been implemented to assess the
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success or otherwise of the reduction in water input as well as a closed circuit
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camera to monitor mouth dynamics including the possibility of anthropogenic
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interference.
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Current situation in Asia
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Legislative framework
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In Asia, China possesses a comprehensive set of laws and regulations
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dealing with coastal areas, including over 25 legislative instruments (Cao and
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Wong, 2007) addressing issues as varied as regulations on dumping (1985, 1992),
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Marine Protected Areas (1994, 1995, 1997), EIA (2002) (Lindhjem et al., 2007),
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and a general implementation of the UNCLOS Convention in 1998 (Keyuan, 2001),
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together with specific dispositions e.g. as regards fisheries (Keyuan, 2003).
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Integrated Coastal Zone Management (ICZM) requires appropriate
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legislation, and additionally benefits from the existence of strong public
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participation and independent coordination (Lau, 2005) thus avoiding the twin
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pitfalls of marginalizing stakeholders and encouraging sectorial management. The
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participation and coordination issues are not easily achieved in the present-day PRC,
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however a pilot structure for ICZM exists in Xiamen (Xue et al., 2004; Peng et al.,
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2006) and is planned for Shanghai (Lau, 2005). Nevertheless, the concept of
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integrated assessment, as set down for instance in the European WFD, does not
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seem to be widely applied.
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Tools and methodologies used in assessing ecological integrity
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A review of available Chinese literature indicates that the application of
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extended tools to assess ecological integrity is incipient (e.g. Xue et al., 2004;
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Huang et al., 2006; Leung, 2006), but that there is a growing national concern in
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shifting from methods based on water chemistry and simple biological diversity
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metrics to more sophisticated approaches which use ecological indicators of
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degradation to provide a more robust assessment. Below two methods are reviewed:
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(i) the coastal eutrophication assessment, compared with the Assessment of Trophic
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Status (ASSETS) model (for details, see. Bricker et al., 2003; Ferreira et al., 2007;
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http://www.eutro.org); and (ii) the integrated Comprehensive Index Assessment
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Method (CIAM), for marine resource assessment.
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Coastal eutrophication assessment
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The application of the ASSETS index to Chinese coastal waters (Xiao et al.,
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2007) provided an opportunity to review the methods currently used in China for
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assessing coastal eutrophication. Historically, this assessment has focused on
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chemical indices, using techniques such as the Nutrient Index Method (NIM), to
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study the effects of system loading by nutrients, and may therefore be considered
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“Phase I” (sensu Cloern, 2001) approaches (Yao, 2005).
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The NIM, proposed by the Chinese National Environmental Monitoring
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Center, is based on a nutrient index (Ni) in seawater (Lin, 1996), calculated using
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Eq.1:
Ni 
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CCOD CTN CTP CChla



S COD STN STP S Chla
(Eq. 1)
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where: CCOD, CTN, CTP and CChla are measured concentrations of Chemical Oxygen
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Demand (COD), total nitrogen, total phosphorus (in mg l-1) and chlorophyll a (in μg
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l-1) in sea water, respectively. SCOD, STN, STP and SChla are standard concentrations of
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COD (3.0 mg l-1), total nitrogen (0.6 mg l-1), total phosphorus (0.03 mg l-1) and
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chlorophyll a in seawater (10 μg l-1), respectively (Lin, 1996). If Ni is greater than 4
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the seawater is considered eutrophic.
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While it is widely used in Chinese coastal systems, research in the past
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decades has identified key differences, to nutrient enrichment, between the
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responses of limnology-originated methods, such as this one, and that of coastal-
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estuarine ecosystems (Cloern, 2001; Bricker et al., 2003; Ferreira et al., 2007).
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Partly, this is because in coastal environments there is often no clear relationship
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between nutrient forcing and eutrophication symptoms—systems with similar
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pressures show widely varying responses. In particular, nutrient concentrations have
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often been shown to be poor indicators of eutrophication symptoms (e.g. Tett et al.,
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2003), since ecosystem responses are modulated by typological factors such as
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morphology, tidal range, natural turbidity and water residence time.
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Marine resources and ecological quality assessment
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Most of the estuaries and coastal inlets and bays are also important fishery
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grounds. The Comprehensive Index Assessment Method (CIAM) has been
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developed and applied to evaluate the ecological quality of the marine fisheries
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environment for the major coastal areas of China (Jia et al., 2003; Jia et al., 2005;
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Ma et al., 2006). CIAM incorporates four assessment modules: seawater quality,
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nutrient level, primary production level, and diet organism richness, the index is the
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mean value of the sub-module indices.
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Seawater quality assessment in CIAM is focused on water pollution status
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evaluation. The main components of coastal pollution in China include organic
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pollution (indicated as COD), eutrophication, total hydrocarbon pollution, and
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heavy metal pollution. The organic pollution status is assessed using the Organic
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Pollution Index method (the A value), while the status of other kinds of pollution is
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assessed using Factorial Analysis (Pi) according to Fishery Water Quality GB
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11607 -1989 and Sea Water Quality GB 3097-1997. The classification of sea water
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quality used in CIAM is given in Table 1. The A value is directly used as Pi during
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the comprehensive ecological quality assessment stage.
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Concentrations of dissolved nitrogen (DIN), phosphate and silicate and their
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ratios are used in the sea water nutrient level assessment using a NIM (the E value).
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The seawater nutrient level is classified as: E=0-0.5, Grade: 1, Nutrient level: Low;
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E=0.5-1, Grade: 2, Nutrient level: Medium; E>1, Grade: 3, Nutrient level:
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Eutrophic. The E value is used as Pi in the CIAM.
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Primary productivity level and diet organism richness are important
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indicators for the fishery environment quality status. Since they vary significantly
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among different areas along the China coast, six grades are used to classify the
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quality status (Table 2). The index level for each item is taken as Pi in the final
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CIAM. The CIAM index is calculated using the following equation:
Ip 
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n
1
n
P
i 1
i
(Eq. 2)
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Ip: comprehensive ecological quality index; Pi: index level assessed for indicator i
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(i.e. indices for seawater quality, nutrient, primary productivity and diet organism
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richness); n: total number of indicators.
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The CIAM of the marine fishery environment is classified into six grades
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according to quality index Ip: excellent, fine, relatively fine, moderate, poor and
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very poor (Table 1).
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Some examples of integrative assessment
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Eutrophication assessment of Jiaozhou Bay, Northeast China
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The ASSETS model was applied to Jiaozhou Bay, together with a
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comparison with the evaluation of eutrophication status using chemical indices. The
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ASSETS method applies a Pressure-State-Response model, focuses on primary and
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secondary symptoms of eutrophication (rather than causative factors such as
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dissolved nutrients), and combines indicators such as algal biomass, harmful algal
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blooms and submerged aquatic vegetation.
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Jiaozhou Bay is located on the west coast of the Yellow Sea (3557-
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3618N, 12006-12021E) with a surface area of 397 km2 and average depth of 7
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m (Editorial Board of “Bays in China”, 1993). It is a semi-enclosed water body,
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connecting to the Yellow Sea through a 2.5 km channel, and has a mean tidal range
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of 2.5-3.0 m. The tides can reach 4.2 m at spring tide, and induce strong turbulent
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mixing, resulting in nearly homogeneous vertical profiles of temperature and
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salinity (Liu et al., 2004).
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The bottom of the bay contains spawning, nursery and feeding grounds for
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fish, and intensive mariculture. Historically, this has focused on the bay scallop
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(Argopecten irradians) and Pacific oyster (Crassostrea gigas), cultivated on long-
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lines. Recently, the long-lines have been removed and the system is used for Manila
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clam (Tapes philippinarum) cultivation, with a production of 200,000 t yr-1. The
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main issue in Jiaozhou Bay is the increase since 1990s of both the frequency and
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magnitude of harmful algal blooms (HABs), although most events are non-toxic
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(Han et al., 2004).
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The volume of Jiaozhou Bay is 1900×106 m3, which, together with a
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nitrogen load into the bay of 30 ton per day (Wang et al., 2006), result in a “High”
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rating for the nutrient component of IF (0.933). Strong tidal mixing and high river
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discharge (8×108 m3 y-1) contribute to moderate flushing and dilution potential
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(Editorial Board of “Bays in China”, 1993). However, the intensive top-down
408
control of the food web has a significant impact on mitigating eutrophic symptoms.
409
The susceptibility component of the IF based on only natural circumstances
410
is considered “Moderate” but when shellfish aquaculture is taken into account, the
411
overall susceptibility is considered to be “Low”. This is one example of the
412
difficulty in universal application of this kind of method, since ASSETS must be
413
potentially adapted to incorporate local societal factors. The combination of “High”
414
nutrient load and “Low” susceptibility gives an overall IF rating of “Moderate Low”.
415
Chlorophyll a is the only indicator with information for the primary
416
symptoms. No information was found for macroalgae, which was therefore
417
classified as “Unknown”. Maximum chlorophyll a values in Jiaozhou Bay did not
418
exceed the threshold indicated in ASSETS for “Medium” eutrophic conditions.
419
ASSETS uses the 90th percentile to provide a typical maximum value for
420
chlorophyll a, and in the bay this value is between 4-5 μg l-1, i.e. in the “Low”
421
category. Therefore, the rating for primary symptoms is “Low” based on chlorophyll
422
a.
423
Discrete data for dissolved oxygen were collected from various sites to cover
424
one annual cycle, as secondary symptom. No information was found for submerged
18
425
aquatic vegetation, but considering the historical scale of kelp aquaculture in the
426
bay, the level for this secondary symptom would be at worst “Low”.
427
Very few values below the ASSETS threshold for biologically stressful
428
dissolved oxygen condition (5 mg l-1) were detected in Jiaozhou Bay. As described
429
earlier, the 10th percentile is applied to provide a more consistent minimum value for
430
dissolved oxygen. In this system, the 10th percentile for annual dissolved oxygen
431
data is between 6-7 mg l-1, indicating no problems for this indicator.
432
According to Han et al. (2004) there were up to 69 harmful algal species
433
observed in Jiaozhou Bay. Toxic blooms are registered episodically, and usually last
434
for only few days, e.g. a Skeletonema costatum bloom was reported to last for five
435
days in July 1998 (Huo et al., 2001). Therefore, the symptom of “nuisance and toxic
436
blooms” is rated as “Low”.
437
In synthesis, the highest level of the three secondary symptoms falls into the
438
“Low” category, and the OEC resulting from the combination of primary and
439
secondary symptoms for this system is “Low”.
440
The estimate based on the current development scenario gives a 9.3%
441
population increase over 20 years (P.R.C. National Bureau of Statistics, 2001). In
442
addition, Qingdao (the main land nutrient source, pop. 8 million) is strongly
443
promoting its tourism industry and less space is available for mariculture in
444
Jiaozhou Bay. Accordingly, the reduced top-down control on primary production
445
could lead to increased eutrophic symptoms. Additionally, Qingdao prepares to host
446
the Olympic Sailing Regattas in 2008, which has focused attention on water quality
19
447
issues and mitigation of eutrophic symptoms. The government has pledged to build
448
more wastewater treatment plants in the near future, and more restrictive pollutant
449
emission regulations are coming into effect (Wang et al., 2006).
450
As a whole, nutrient loads are expected to decrease in spite of the increase in
451
the urban population, and the water quality in Jiaozhou Bay is likely to improve.
452
The Future Outlook can therefore be considered to be “Improve low”. Table 3
453
summarizes the results obtained from the application of ASSETS to Jiaozhou Bay,
454
which resulted in an overall score of “High Status”, indicative of a system without
455
eutrophication problems.
456
The result is better than expected due to top-down control related to
457
intensive shellfish mariculture in this system. This has important implications for
458
successful management of nutrient-related problems, which are not captured by the
459
Chinese Nutrient Index “Phase I” methods, which classify the system as Eutrophic.
460
Furthermore, the Nutrient Index Methods cannot by definition be clear indicators for
461
a large system, because there is no accommodation for spatial differences in impact
462
level within a waterbody. The evaluation of systems by salinity zone, as in
463
ASSETS, contributes to a more accurate evaluation of the system and subsystems,
464
which is necessary to target management efforts. A comparison of various methods
465
for eutrophication assessment is shown in Table 4.
466
The top-down control of the food web in Jiaozhou Bay suggests a feasible
467
way to manage the eutrophication in a coastal system. These strategies for
468
eutrophication control, which have traditionally been used in China for many years,
20
469
are being discussed with respect to practical implementation in the EU and the US
470
(e.g. Lindahl et al., 2005; Ferreira et al., 2007). Paradoxically, the Chinese, U.S. and
471
other governments and scientists currently focus mainly on a bottom-up approach in
472
improving water quality, though there is still plenty of scope to promote top-down
473
control in the food chain. Water quality data collected in Jiaozhou Bay during 1999-
474
2000 were used to estimate the gross removal of algae by Manila clams. On the
475
basis of reported bivalve stocks, these organisms annually remove about 627 t yr-1
476
of chlorophyll a, which (considering a carbon:chlorophyll ratio of 50 and the
477
standard Redfield C:N ratio of 45:7 in mass) corresponds to the removal of almost
478
4,900 t yr-1 of nitrogen, i.e about 1.5 million population equivalents, or 17% of the
479
population of Qingdao, and to about 45% of the estimated 11,000 t yr-1 nitrogen
480
load. Along with economic benefits, the introduction of filter-feeders on a
481
reasonable scale thus allows for cost-effective removal of nutrients and mitigation of
482
eutrophic conditions, which is more environmentally-friendly and sustainable for a
483
coastal system (Shastri and Diwekar, 2006).
484
Marine resources and ecological quality assessment in the South China Sea
485
In order to understand the health status and ecological quality of the fishery
486
environment in the northern South China Sea, a comprehensive and systematic
487
survey program was carried out (including Taiwan bank, East Guangdong, Pearl
488
River estuary, West Guangdong, Southern waters of Hainan, and Beibu Bay), from
489
1997 to 2002. The quality status of the fishery environment in the northern South
490
China Sea was assessed based on the data of seawater quality, sea water nutrient
21
491
structure and nutrient level, primary productivity and diet organism level using
492
CIAM (Jia et al., 2005).
493
The results showed that the overall water quality index was within the
494
criteria limit of Fishery Water Quality GB 11607 - 1989 and the Grade criteria of
495
Sea Water Quality GB 3097 – 1997. The organic pollution index (A value) range
496
was 0.411~0.237, and the nutrient index (E value) range was 0.10~0.34, which
497
indicated the waters were not organically polluted and the nutrient status was low.
498
The primary productivity of the waters, ranging from Grade 5 to Grade 1, with an
499
annual average of Grade 3, was in a “medium” level. For the richness of diet
500
organism, the grade of phytoplankton, zooplankton and benthic organisms were 3, 5
501
and 4 respectively, within a “relatively high” level in general.
502
The comprehensive quality assessment results (Table 5) showed that the
503
quality indices of 9 factors (including DO, DIN, PO43, A, E, primary productivity,
504
phytoplankton, zooplankton and benthos) were all lower than 1.0 and the
505
comprehensive quality index was 0.58, which indicated the ecological quality of the
506
overall area was relatively fine. However, the comprehensive ecological quality
507
index of Pearl River estuary, East Guangdong waters, West Guangdong waters and
508
Beibu Bay were all over 0.60, and the quality status was “moderate”, far worse than
509
the “fine” level, which is a sign of environmental degradation along the coast of
510
Guangdong Province, and means that due to the continuous and rapid growth of
511
industry and economy of Guangdong, especially in the Pearl River Delta, more
512
attention should be paid to environmental protection and ICZM.
22
513
514
Current situation in Australia
515
Legislative framework
516
Australia adopted an Oceans Policy in 1998 and subsequently established a
517
National Oceans Office that initiated a process of marine bioregional planning.
518
These measures reflected requirements and obligations that arose for Australia as a
519
signatory as the United Nations Convention on the Law of the Sea (UNCLOS) came
520
into force in 1994. Subsequently, the Environment Protection and Biodiversity
521
Conservation Act 1999 (EPBC Act) provided the overarching framework for
522
management of Australia’s national and international marine environmental
523
responsibilities.
524
525
In 2005 the Australian Government brought its program of regional marine
planning directly under the scope of the EPBC Act 1999.
526
However, as a federal nation, the situation is complicated by the history of
527
internal jurisdictional responsibilities for the original 3 nautical mile territorial sea
528
that existed prior to the development of arrangements under UNCLOS. In 1975 a
529
decision of the High Court of Australia upheld legislation vesting sovereignty in
530
respect of the territorial sea in the Federal Government. The implications of this
531
were addressed by the Offshore Constitutional Settlement of 1979 whereby
532
legislative competence and proprietary rights over the 3 nautical mile territorial sea
533
were transferred to the States and Territories through the Coastal Waters (State
534
Titles) Act (1980). This revalidated pre-existing state legislation and restored the
23
535
title, responsibilities and rights of states and territories with respect to the seas,
536
seabed and subseabed within the 3 nautical mile territorial sea and internal waters.
537
As a consequence the States and Territories have primary responsibilities
538
for management of marine environments, natural resources and the impacts of
539
human activities within internal waters and the 3 nautical mile territorial sea. The
540
States and Territories have differing approaches to management of marine
541
ecological integrity.
542
Federal responsibilities relate primarily to areas beyond 3 nautical miles
543
although Section 23 (2) of the EPBC Act (1999) gives the Federal Government an
544
overarching capacity to address issues that actions taken outside Commonwealth
545
Marine Areas that have, will have or are likely to have a significant impact upon
546
them.
547
In parallel with the development of a broader political and legislative
548
framework for marine environmental management in the context of UNCLOS,
549
Australia was engaged in the implementation of the Great Barrier Reef Marine Park
550
Act (1975) which provides for the Great Barrier Reef Region to be for conservation
551
and reasonable use.
552
Currently there is no standard set of environmental indicators used across
553
Australia by the states and federal governments. After 2001 the Australian State of
554
the Environment (SOE) reports, an attempt was made at the following Australian
555
and New Zealand Environment and Conservation Council (ANZECC) to obtain
556
some uniformity of indicators across Australia. A core set of 75 indicators was
24
557
established but efforts to reduce this to a smaller core set were unsuccessful.
558
Subsequently for the SOE 2006 process, a data reporting system (DRS)
559
(www.deh.gov.au/soe/DRS) was developed. However a brief review of the various
560
SOE’s for each state government for an evaluation of the status of estuarine and
561
inshore coastal waters, revealed no uniformity or consistent pattern of indicators
562
being used between the states, although virtually all were physical parameters with
563
no biological indicators being used.
564
In 2004 there a comprehensive report prepared on estuarine, coastal and
565
marine indicators for regional NRM monitoring by the CRC for Coastal Zone,
566
Estuary and Waterway Management (give website) and they summarised the issues
567
to be targeted such as inland aquatic ecosystems integrity and estuarine, coastal and
568
marine habitat integrity as well nutrients, turbidity and surface salinity in freshwater
569
aquatic environments, significant native species and ecological communities and
570
invasive species. For each of these items for targeting they provide a detailed
571
summary of useful indicators. However there is no attempt to discuss the actions to
572
be taken by the relevant authorities (of which there are many) to act upon the results
573
obtained from any monitoring undertaken. The situation is confused by the State
574
Federal boundaries. It is unclear as to whether any real progress has been made in
575
co-ordinating these measures and then acting on them.
576
Tools and methodologies used in assessing ecological integrity
577
In parallel with the development of a broader political and legislative
578
framework for marine environmental management in the context of UNCLOS,
25
579
Australia was engaged in the implementation of the Great Barrier Reef Marine Park
580
Act (1975) which provides for the Great Barrier Reef Region to be managed for
581
conservation and reasonable use.
582
With an extent of 350,000 km2 and little physical survey of areas beyond
583
shipping lanes, preparation for ecosystem scale management of the Great Barrier
584
Reef required development of approaches to implement a working understanding of
585
bioregionalisation in a data-lean environment (Kenchington, 1990). The initial tasks
586
included commissioning of a geomorphological classification of reefs and shoals at
587
a scale of 1:250,000. This was followed by detailed surveys of reefs conducted by
588
the Australian Survey Office and subsequently by the development of 1:250,000
589
rectified maps of the Great Barrier Reef region drawing on LANDSAT imagery to
590
infill ground survey data (Kenchington, 1990). These maps were used for expert
591
consultation on the distribution of ecological communities and fisheries resources.
592
Data on the occurrence and distribution of biological communities was extremely
593
patchy with major data sources being reef research stations and expeditionary
594
studies. The community and resource distribution maps resulting from expert
595
consultation were then used to seek comment and amendment during a phase of
596
public consultation prior to development of a draft zoning plan for a section of the
597
Great Barrier Reef Marine Park (Kenchington, 1990).
598
The initial zoning of the Great Barrier Reef took account of seabed
599
communities known from fisheries surveys and production and benthic communities
600
such as Halimeda algal beds and sponge beds but focussed largely on coral reef
26
601
communities. In 1998, the Great Barrier Reef Marine Park Authority started a
602
process of review of zoning in the light of experience, new information and changed
603
circumstances of use and management of the Marine Park (Lawrence et al., 2002).
604
A major element identified in the context of management responsibilities
605
was the need for an adequate network of highly protected (no take) areas
606
representative of all of the bioregions occurring within the Great Barrier Reef World
607
Heritage Area. Using groups of specialists GBRMPA convened a series of
608
facilitated workshops and they identified
609
There are 30 reef and 40 non reefal bioregions which could be characterised
610
by physical and biological features. A decision was made that at least 20% of each
611
of these bioregions was to be declared as a no take zone, after the complete
612
rezoning, many of these bioregions were represented by more than 20% no take.
613
The location of these no take zones took into account economic and social factors to
614
try minimising the impact of these zones on the users of the GBR. Ongoing
615
monitoring during the next few years will attempt to obtain data to support the
616
declaration of these no take zones and identify if additional such zones need to be
617
declared, especially in the light of ongoing climate change.
618
619
Some examples of integrative assessment
620
An example is the Integrated Marine and Coastal Regionalisation for
621
Australia (IMCRA). The methods developed for assessment of the Great Barrier
622
Reef region were an important input to the process of developing a marine
27
623
biogeographic regionalisation of Australia which was required to address
624
commitments in connection with the creation of a national system of representative
625
marine protected areas. A workshop in 1994 (Muldoon, 1995) led to the
626
development of an initial interim biogeographic regionalisation (Thackway and
627
Creswell, 1995), a series of updated versions and most recently in 2006 to an
628
integrated marine and coastal regionalisation of Australia.
629
The IMCRA has developed best contemporary understanding of provincial
630
bioregions based on regionalisation of demersal fish communities (CSIRO, 1996).
631
Nested within this, a meso-scale regionalisation has been developed using finer
632
scale information provided by relevant State and Northern Territory agencies. The
633
third element is a map of sea bed classified into 14 classes of regions of similar
634
geomorphology.
635
636
The IMCRA process is dynamic providing for updates as new data comes to
hand indicating need for revisions.
637
638
Current situation in Europe
639
Legislative framework
640
In recent years the European Union has adopted several Directives for the
641
nature protection (e.g. for Habitats and Species, Wild Birds, or Environmental
642
Impact Assessment). However, increasing pressures and impacts within European
643
estuaries and coasts, have lead to the approval of a series of laws focus on water
644
management (e.g the Water Framework Directive, the Recommendation on
28
645
Integrated Coastal Zone Management (ICZM), the Directive on the Marine Strategy,
646
and the Maritime Policy) (for details on these Directives many references can be
647
consulted, such as Borja et al. (2004a), Borja (2005, 2006), or a recent special issue
648
in Marine Pollution Bulletin (2007: 55(1/6)). All these Directives emphasise the
649
increasing need of protecting European coastal and estuarine ecosystems and the
650
move towards marine integrative management.
651
Hence, the main objective of the WFD is to achieve a Good Ecological
652
Status, for all European water bodies, by 2015. On the other hand, the EMS requires
653
achieve a Good Environmental Status, for all European seas, by 2021.
654
655
Tools and methodologies used in assessing ecological integrity
656
The achieving of the abovementioned objectives requires the development of
657
tools and methodologies suitable to assess the environmental quality in an
658
integrative way. The integration of several elements of the ecosystem, including
659
physico-chemical and biological elements (phytoplankton, zooplankton, benthos,
660
algae, phanerogams, fishes), is essential in the assessment.
661
The European scientific community is developing many different
662
methodologies to assess the quality of each of the elements separately, within the
663
WFD. Some of these methodologies can be consulted in volume 55 (issues 1-6) of
664
Marine Pollution Bulletin, together with the report “WFD intercalibration technical
665
report.
666
www.xxxxxxxxxxxx). Conversely, the efforts addressed until now for the EMS
Part
3
–
Coastal
and
Transitional
Waters”
(March,
2007.
29
667
implementation have been focused in fisheries management, under the EBA
668
(Browman et al., 2004; Nicholson and Jennings, 2004; Rice et al., 2005; Frid et al.,
669
2006).
670
However, very few studies have been published in integrating all physico-
671
chemical and biological elements in a unique assessment of the ecosystem status
672
(Borja et al., 2004a).
673
674
Some examples of integrative assessment
675
The approach of Borja et al. (2004a) was expanded to include physico-
676
chemical (Bald et al., 2005), chemical (Borja et al., 2004b; Borja and Heinrich,
677
2005; Rodríguez et al., 2006), and benthic communities (Borja et al., 2000; Muxika
678
et al., 2007).
679
The WFD established two different quality statuses: chemical and ecological
680
status. The chemical status is based upon metal and organic compound
681
concentrations, and is determined by comparing monitored concentrations with
682
quality objectives (QO) (if concentrations are under QO, the chemical status is met;
683
if concentrations are over QO, the chemical status is not met). The WFD only
684
mentions water quality in assessing the chemical status, but some authors include
685
sediment and biomonitoring quality (Borja et al., 2004b, 2006; Borja and Heinrich,
686
2005) or just sediment (Crane, 2003) (Table 6).
687
Conversely, the ecological status integrates physico-chemical, chemical and
688
biological indicators. The physico-chemical indicators used in this assessment are
30
689
those supporting the biological elements (thermal conditions, salinity, oxygen,
690
nutrients, transparency). The physico-chemical status is assessed by means of a
691
Factorial Analysis (FA). Hence, the projection of each sampling station, to the line
692
connecting reference conditions of ‘high’ and ‘bad’ status, is calculated in the new
693
3-dimensional space defined by the FA (see Bald et al. (2005), for details).
694
Consequently, those stations located near the high reference would represent a
695
‘high’ physico-chemical status, and stations located near the bad reference, would
696
be classified as in ‘bad’ physico-chemical status. Intermediate stations would be
697
classified in ‘good’, ‘moderate’ or ‘poor’ status.
698
Pollutant concentrations are also used in assessing the ecological status, but
699
only to determine ‘high’, ‘good’ and ‘moderate’ status (see below). In this particular
700
case, the WFD defines ‘high’ status, when concentrations of pollutants remain
701
within the range normally associated with undisturbed conditions (i.e., below the
702
background level). The concentrations between background levels and QO are in
703
accordance with the WFD ‘good status’ definition, while ‘moderate’ status can be
704
considered when concentrations are over QO (for details in this assessment, see
705
Rodríguez et al. (2006)).
706
The metrics used in the methodologies implemented for the biological
707
quality assessment, within the WFD, are very diverse (Table 7). These
708
methodologies include multimetric and multivariate approaches (Borja et al., 2004a;
709
Muxika et al., 2007). These authors provide the way to assess the quality of each of
710
the individual biological elements (i.e. phytoplankton, macroalgae, benthos or
31
711
fishes). However, it is also necessary to integrate these individual results in a unique
712
quality value (Borja et al., 2004a).
713
Following some interpretations, the classification of the ecological status in
714
the WFD should be based upon the worst of the values in the biological elements.
715
Hence, if the phytoplankton has a moderate value and the remainder of the elements
716
has a high status, the global classification should be moderate ecological status.
717
Taking into account the variability of some of the biological elements, and the
718
absence of accurate methodologies in assessing the biological status, Borja et al.
719
(2004a) propose the weighting of those elements, such as benthos, with contrasted
720
and intercalibrated (Borja et al., 2007) methodologies. Hence, a decision tree
721
permits to derive a more accurate global classification, including also physico-
722
chemical and chemical elements (Table 8).
723
Most of the methodologies used within the WFD, determines the quality at
724
the sampling station level. However, the Directive requires integrating quality at the
725
water body level. Hence, one of the possible ways to achieve this is by using Table
726
9. This example illustrates a water body with 4 sampling stations, each of them
727
representative of a certain surface, within the water body. When obtaining the status
728
for each station, this result can be substituted by an equivalent value, which allows
729
weighting the global status, depending on the representativeness of each of the
730
sampling stations. The same approach can be used in a previous step, when
731
integrating values of biological or physico-chemical elements, as shown in Table 8.
732
32
733
Current situation in North America
734
Legislative framework
735
In the United States (US), recognition of short and possible long-term
736
impacts led to legislation to monitor conditions and study the causes of
737
eutrophication, in order to provide a basis for protecting water bodies from further
738
degradation and for appropriate management measures that will improve water
739
quality at impaired sites. The main legislations for the US are the Clean Water Act
740
(CWA) of 1972, Air Pollution Prevention and Control Act of 1977, Coastal Zone
741
Management Act of 1972, Harmful Algal Bloom and Hypoxia Research and Control
742
Act of 1998. The responsibilities and interactions among US agencies are shown in
743
Figure 1.
744
National monitoring and assessment of water quality and nutrient inputs in
745
the US is shared by several federal agencies, primarily the Environmental Protection
746
Agency (EPA) and the National Oceanic and Atmospheric Administration (NOAA).
747
The EPA is charged with regulating most aspects of water quality (e.g. development
748
of nutrient criteria, 305(b) and 303(d) reporting requirements and total maximum
749
daily load (TMDL) development) per the federal CWA (USEPA, 2003).
750
Responsibility for implementing water quality standards and criteria, and for
751
monitoring to assess attainment, is generally delegated to state water management
752
authorities. States and tribes are thus required to report periodically to the EPA on
753
water quality conditions within their states, and to develop plans to remedy
754
impairments when they occur. EPA and NOAA support regulatory decisions by
33
755
providing research and assessment results and share some management
756
responsibilities (e.g. Coastal Zone Management Act Reauthorization Amendments
757
[CZARA] Section 6217 coastal non point pollution control program; CZMA, 1996).
758
The CWA establishes that, wherever possible, water quality must provide for
759
the protection and propagation of fish, shellfish, and wildlife, and recreation in and
760
on the water and/or protection of the physical, chemical, and biological integrity of
761
those waters. In adopting water quality standards, states and tribes designate uses for
762
their waters in consideration of these CWA goals, and establish water quality
763
criteria (that contain sufficient factors) to protect integrity and those uses. In 1998,
764
EPA developed the National Strategy for the Development of Regional Nutrient
765
Criteria (USEPA, 1998). This strategy presented EPA’s intention to develop
766
technical guidance manuals for four types of waters (lakes and reservoirs, rivers and
767
streams, estuaries and coastal waters, and wetlands). The approaches described in
768
these manuals have been applied by EPA and resulted in the publication of 26
769
ecoregional nutrient criteria documents, under CWA section 304(a) for lakes and
770
reservoirs, rivers and streams, and wetlands, to assist states and tribes in establishing
771
numeric nutrient criteria to protect uses when adopting water quality standards. To
772
date, there have been no such 304(a) criteria established for any U.S. estuary,
773
however, the technical guidance manual for estuaries and coastal waters was
774
published in fall 2001 (USEPA, 2001). It presents EPA’s recommended approaches
775
for developing nutrient criteria values for estuaries. In this document there was
34
776
ambiguity regarding whether nutrient concentrations or nutrient loading are to be
777
considered optimal in the context of estuarine eutrophication management.
778
779
Tools and methodologies used in assessing ecological integrity
780
There are several methods for evaluating ecological integrity or condition
781
status of coastal waters and they have different issues of focus. Here we discuss and
782
give results for the application of two methods for eutrophication assessment, the
783
National Estuarine Eutrophication Assessment/Assessment of Estuarine Trophic
784
Status conducted by the National Oceanic and Atmospheric Administration and the
785
Environmental Protection Agency’s National Coastal Assessment. The EPA NCA
786
includes indices for evaluation of additional pollutants and ecological compartments
787
including a Benthic Index which here is compared to the IBI developed elsewhere
788
(Dan…..you have expertise in this which I do not, could you add some discussion
789
here? Thanks!). All of these methods are designed to evaluate conditions and some
790
methods also address the causes for observed impairment with the intent to inform
791
management.
792
793
NOAA’s Estuarine Eutrophication Assessment and the Assessment of Estuarine
794
Trophic Status (NEEA/ASSETS)
795
NOAA’s nationwide assessment of 141 US systems used assessment of
796
individual systems, synthesized to regional and national levels, to show that nutrient
797
related water quality problems occur on a national basis (Bricker et al., 1999;
35
798
NOAA, 1996, 1997a, b, c, 1998). The recent update examines changes that have
799
occurred since the early 1990s (Bricker et al., 2007). The NEEA Program
800
established working groups to improve accuracy and applicability of the
801
methodology with the intent of being complementary to the National Research
802
Program for Nutrient Pollution in Coastal Waters (Howarth et al., 2003), interactive
803
with the EPA NCA, and to support efforts by US states and the European
804
Commission (EC) member states to fulfill requirements of the CWA section 305(b)
805
and the EU WFD (eg. Coast, 2003 and OSPAR, 2002), respectively.
806
The method and ongoing modifications are described here in brief. A full
807
description of the Pressure-State-Response method can be found in Bricker et al.
808
(1999, 2003, 2007), Ferreira et al. (2007), Scavia and Bricker (2006) and at
809
http://www.eutro.org and http://www.eutro.us .
810
811
Pressure–Influencing Factors (IF): Pressure is determined by a matrix combines the
812
magnitude of inputs from the watershed with some measure of the system’s ability
813
to dilute or flush the nutrient inputs (i.e., susceptibility). A model compares
814
anthropogenic nutrient loading with natural background concentrations and also
815
factors in potential nutrient inputs from oceanic sources providing insight to whether
816
management measures would be successful. Where monitoring data is inadequate,
817
model results are used to estimate nutrient loads and sources from watersheds and to
818
examine the primary causative factors in the development of problems and potential
36
819
benefits of BMP application (eg. USGS SPARROW model, Smith et al., 1997, and
820
WATERSN model, Castro et al., 2001; Whitall et al., 2003, 2004).
821
822
State– Overall Eutrophic Condition (OEC): Five variables are used to determine
823
condition. These are divided into two groups: 1) chlorophyll a (Chl) and
824
macroalgae, called primary symptoms, indicate early stages of eutrophication; and
825
2) low dissolved oxygen (DO), losses of submerged aquatic vegetation (SAV), and
826
occurrence of nuisance and/or toxic algal blooms (HABs), called secondary
827
symptoms, are indications of well-developed problems. An area-weighted-estuary-
828
wide value for each variable was determined based on concentration, spatial
829
coverage, and frequency of occurrence of extreme or problem conditions. The
830
primary symptom expression level is determined by averaging the values for Chl
831
and macroalgae, while for secondary symptoms the highest score (worst impact) of
832
the three is selected for precautionary reasons. These values are combined in a
833
matrix to determine an overall ranking of eutrophic conditions, or impact level, for
834
the estuary which fall into one of five categories; High, Moderate High, Moderate,
835
Moderate Low or Low.
836
Chl and DO scores are determined using statistical criteria (percentile 90 for
837
Chl and percentile 10 for DO using annual data; Bricker et al., 2003) while
838
macroalgae, HABs and SAV loss are determined heuristically. Improvements have
839
been proposed for macroalgae and SAV assessments based on a comparison of
840
potential colonization area and observed colonized area. Threshold levels for
37
841
specific HABs have also been considered, such as the index for brown tides
842
developed for the Maryland Inland Bays by Gastrich and Wazniak (2002).
843
844
Response–Future Outlook (FO): Response, the expected condition at some time in
845
the future, is determined by combining susceptibility of the system with expected
846
changes in nutrient loads. Predictions of future nutrient loading (increase, decrease,
847
unchanged) are based on predicted changes in population and watershed uses,
848
mitigated by planned management actions. Note that this is not the traditional use of
849
response which usually refers to the behavioral or management response designed to
850
mitigate and prevent further problem occurrences. A more traditional Response
851
component is being developed to complement the existing Response component
852
(Ferreira et al., 2005)
853
854
ASSETS Synthesis: The three components are then combined into a single rating for
855
the estuary resulting in a rating of: Bad, Poor, Moderate, Good or High.
856
857
858
859
Modificatios to the NEEA/ASSETS:
Modifications include development of a type classification, type-specific
indicator thresholds, and a human use indicator (Bricker et al., 2007).
860
Type classification: Addition of a classification by type is expected to
861
improve the assessment accuracy and management effectiveness. Typology groups
862
embayments within a small number of classes based on similarity of physical
38
863
characteristics such as morphology, freshwater discharge and salinity stratification,
864
using both qualitative and quantitative criteria. The aim is to use type specific
865
indicators that are manageable and can be used to characterize status and provide
866
“early warning” of impending type-specific problems. The number of types should
867
be kept as small as possible, and a subset of well-studied systems used to represent
868
each type to reduce monitoring costs while maximizing prospects for successful
869
management of systems that are not well studied.
870
The EPA has developed a classification using a similar clustering approach,
871
however, additional pollutant stressors are included in the EPA classification
872
framework (eg., suspended sediments, toxics, habitat alteration) as are other types of
873
water bodies (i.e. the Great Lakes and Great Lakes coastal riverine wetlands;
874
Burgess et al., 2004). Despite the differences, the EPA classification resulted in 11
875
groupings compared to the 10 NEEA groups (Bricker et al., 2007; Kurtz et al.,
876
2006). The EPA and NOAA are currently collaborating on testing and
877
intercalibrating the two classifications.
878
Human use impacts: Despite its importance, few previous studies have
879
looked at the social and economic costs of eutrophication, and those that do focus on
880
lost productivity because indirect and non-use costs are so difficult to measure.
881
Given the complex nature of eutrophication, there are a variety of potential human-
882
use impacts, (e.g. commercial and recreational fishing, fish consumption,
883
swimming, boating, aesthetics, tourism) that can be examined, however, adequate
884
data is not available for most of these activities (Bricker et al., 1999, EPA 2005).
39
885
Recreational fishing is important in most estuaries and one that is often directly
886
impacted by eutrophication. Fishing data can be combined with water quality
887
monitoring data and analyzed to determine whether recreational fishing catch rates
888
are related to eutrophic conditions within particular estuarine systems (e.g. Lipton
889
and Hicks, 1999, 2003; Bricker et al., 2006). When a significant relationship is
890
found, recreational fishing catch rates, with appropriate adjustments for other
891
influencing factors, can be used as an indicator of human use impairment due to
892
eutrophication. With additional data and analysis, a dollar value estimate of lost
893
economic welfare can be estimated directly using techniques such as travel cost and
894
random utility models (Herriges and Kling, 1999). Alternatively, with a large
895
number of recreational fishing value studies available in the literature, an
896
approximation of lost economic value can also be determined using benefits transfer
897
(Walsh et al., 1992).
898
899
EPA National Coastal Assessment
900
The US EPA's National Coastal Assessment (NCA) Program surveys the
901
condition of the nation's coastal resources through nationally integrated,
902
comprehensive monitoring in coastal states. The Program is implemented through a
903
federal – state partnership and is designed to answer broad-scale questions on
904
environmental conditions to supplement section 305(b) of the CWA, which requires
905
EPA to report periodically on the condition of the nation’s waters (USEPA, 2003).
906
All samples are collected at stations that are randomly selected using the EPA
40
907
Environmental Monitoring and Assesment Program’s (EMAP) probablilstic
908
sampling framework. Data were sampled once per year during a summer index
909
period (June to October) which typically represents the time period of greatest
910
observed nutrient-related impacts (USEPA, 2001a).
911
The National Coastal Condition Reports (NCCR1, NCCR2; USEPA, 2001,
912
2004), developed and reported regional and national results for five primary indices
913
using data available from the NCA and other national coastal programs: water
914
quality index, sediment quality index, benthic index, coastal habitat index, and fish
915
tissue contaminants index. These indices were selected because of the availability of
916
relatively consistent data sets for these indicators for most of the country. They do
917
not address all characteristics of estuaries and coastal waters that are valued by
918
society, but they provide information on both ecological condition and human use of
919
estuaries. The Water Quality Index (WQI) is the indicator describing nutrient related
920
conditions and will be the one discussed here for comparison to the
921
NEEA/ASSETS. It is a combination of data on the status of 5 indicators: dissolved
922
inorganic nitrogen (DIN), dissolved inorganic phosphorus (DIP), Chl, water clarity
923
(Secchi depth) and DO. A rating is developed for each of the 5 indicators. The WQI
924
is intended to characterize acutely degraded conditions within coastal regions during
925
the index period. The aim is to develop a national picture through a probabilistic
926
statistical approach that is not expected to capture site-specific details. If a site is
927
designated as fair or good, it may be characterized by poor conditions for short time
928
periods but outside of the index period.
41
929
The NCA Benthic Index is a set of regionally based (Northeast, Southeast,
930
and Gulf coasts) benthic indices of environmental condition for estuaries that reflect
931
changes in diversity and population size of indicator species to distinguish degraded
932
benthic habitats from undegraded benthic habitats (Engle et al., 1994; Weisberg et
933
al., 1997; Engle and Summers, 1999; Van Dolah et al., 1999). These indices reflect
934
changes in benthic community diversity and the abundance of pollution-tolerant and
935
pollution-sensitive species. A high benthic index rating for benthos means that
936
samples taken from an estuary’s sediments contain a wide variety of species, a low
937
proportion of pollution-tolerant species, and a high proportion of pollution-sensitive
938
species. A low benthic index rating indicates that the benthic communities are less
939
diverse than expected, are populated by more pollution-tolerant species than
940
expected, and contain fewer pollution-sensitive species than expected. The benthic
941
condition data presented throughout this report were collected by the NCA Program
942
unless otherwise noted. Indices vary with region because species assemblages
943
depend on prevailing temperatures, salinities, and the silt-clay content of sediments.
944
(Dan – I stole this from the NCCR2 introductory chapter…feel free to modify!)
945
The NCA complements the CWA Sections 305(b) and 303(d) state reporting
946
requirements, which identifies waterbodies with eutrophication indicators that do
947
not meet adopted state water quality criteria for designated uses. For waterbodies
948
that do not meet the criteria (i.e. impaired waters) special state actions are required,
949
specifically inclusion on the Section 303(d) list of impaired waters, which
950
establishes protocols, such as development of total maximum daily load (TMDL)
42
951
analyses, that must be taken to correct pollution caused impairments (USEPA,
952
2003).
953
Each of the five indicators are assessed for each individual site based on
954
existing criteria, guidelines, or the interpretation of scientific literature. The five
955
indicator ratings are given equal weight and combined to give an overall rating for
956
the site. Then a regional rating for each indicator is developed based on the
957
condition of individual sites within the region. For example, in order for a region to
958
be rated poor with regard to the dissolved oxygen indicator, more than 15% of the
959
coastal area in the region must have dissolved oxygen measured at less than 2 mg.
960
L-1. A particular strength of this program is that the regional criteria boundaries (i.e.,
961
percentages used to rate each regional condition indicator) were developed by EPA
962
in conjunction with NOAA and other agencies, environmental managers, resource
963
experts, and the knowledgeable public. This type of collaboration, designed to
964
produce more accurate assessments should be encouraged and continued as type
965
classification and load response relationships are developed for the US and
966
waterbodies globally.
967
Other Benthic Index/Indices……Dan – this is a place holder for
the
968
benthic description of which I have little expertise, feel free to move etc it’s just so
969
it has a place
970
Some examples of integrative assessment
971
972
The results presented here are from applications of the NEEA/ASSETS and
the EPA National Coastal Assessment.
43
973
974
NOAA’s Estuarine Eutrophication Assessment and the Assessment of Estuarine
975
Trophic Status (NEEA/ASSETS)
976
The NEEA/ASSETS method has been used to evaluate individual systems,
977
results that can then be summarized regionally and nationally to provide a broader
978
picture of conditions. The results that follow are from the recent National Estuarine
979
Eutrophication Assessment (Bricker et al., 2007).
980
981
Influencing Factors: The majority of systems assessed (36 of 64) had high
982
influencing factor ratings, indicating that these estuaries receive a large amount of
983
nitrogen from human related activities compared to their capacity to dilute or flush
984
nutrients (Figure 2). High nitrogen loads were largely attributed to the influence of
985
expanding and dense coastal human populations. Influencing factor ratings were
986
high along the Gulf of Mexico and mid – south Atlantic coast, low in the North
987
Atlantic, and mostly unknown in the Pacific region. Low influencing factors in the
988
North Atlantic region are likely due to relatively low nitrogen loads and strong tidal
989
flushing. In contrast, high ratings in the mid-Atlantic region likely result from high
990
nitrogen loads relative to susceptibility. The high geographic variability of nitrogen
991
loading and susceptibility indicates a need for locally tailored management action.
992
993
Overall Eutrophic Condition: Eutrophication is a widespread problem throughout
994
most of the regions (Figure 3; for more detailed regional results (see Chapter 4 in
44
995
Bricker et al., 2007). The majority of estuaries assessed (64 of 99 systems) had
996
eutrophic conditions rated as moderate to high. These systems represent 78% of the
997
assessed estuarine surface area. The largest concentration of highly eutrophic
998
systems is located around Chesapeake Bay, an area that also has high influencing
999
factor ratings and the highest population density of any region. There were 35
1000
systems (21% of assessed surface area) which exhibited low to moderate low overall
1001
eutrophic conditions. More than half of these estuaries were located in the Gulf of
1002
Mexico and Pacific Coast regions. Overall eutrophic condition (OEC) and symptom
1003
expressions were geographically variable though systems with high OEC ratings
1004
were mainly located in the mid-Atlantic region (Figure 3). There were differences in
1005
eutrophic status among estuaries in close proximity with the net effect of this
1006
variability being no national pattern of overall eutrophic conditions or symptom
1007
expressions. The most commonly occurring eutrophic symptom was high spatial
1008
coverage and high frequency of elevated chlorophyll a levels (Figure 4a). Most
1009
estuaries also exhibited at least one other moderate to high symptom expression in
1010
addition to chlorophyll a (e.g. nuisance/toxic blooms; Figure 4b).
1011
The most frequently noted caused of eutrophic impairments were agricultural
1012
activities (crops and animal operations), urban runoff, wastewater treatment plants
1013
and atmospheric deposition. These are the major sources of nutrients that are
1014
included in the total loads used to determine influencing factors. There are patterns
1015
that exist between influencing factors and eutrophic conditions. Of the 15 estuaries
1016
with high OEC, 13 had high susceptibility scores. In contrast, of the 35 systems with
45
1017
moderate low or low OEC scores, 31 had low or moderate susceptibility. Eleven of
1018
these 31 systems had a low or moderate low OEC rating despite high nutrient loads.
1019
These systems seem able to naturally suppress eutrophication.
1020
Data confidence and reliability (DCR) varied among systems. The general
1021
trend showed moderate DCR ratings in systems with high overall eutrophic
1022
conditions but low DCR ratings in systems with moderate or low OEC ratings. Most
1023
of the systems with high confidence are located in the North and mid-Atlantic
1024
regions, while those with low confidence are located in the Gulf of Mexico and
1025
Pacific Coast regions.
1026
1027
Changes in eutrophic condition since the 1990s
1028
Comparison of eutrophic conditions assessed from the early 1990s to 2004
1029
indicates similar levels of eutrophication (Figure 5) despite population increases in
1030
coastal counties nationally of 13% between 1990 and 2003 (Crossett et al., 2004).
1031
Among the systems where data were available for comparison (58), conditions in 32
1032
had remained the same (77% assessed surface area) while conditions in 13 systems
1033
had improved (9% of assessed surface area), and in another 13 systems conditions
1034
had worsened (14% assessed surface area). In the early 1990s, 69% of assessed
1035
systems (72% of assessed area) had moderate to high eutrophic conditions compared
1036
to 65% of assessed systems (78% of assessed area) in 2004 (Figure5). However,
1037
these results must be viewed with caution, because the total number of systems for
1038
which conditions are unknown has increased from 17 in the 1999 assessment to 42
46
1039
in this assessment. Similar to the early 1990s, the Pacific Coast region had the least
1040
robust data and the lowest assessment confidence.
1041
1042
Future Outlook: Worsening conditions by 2020 were predicted for 65% of estuaries
1043
and improvements predicted for 20% of estuaries for which the analysis was
1044
performed (Figure 6). This is largely based on population estimates which, on a
1045
national basis, suggest that coastal populations will increase by 12% during the
1046
same time frame (Crossett, pers. comm..). This is a bleak outlook for the nation’s
1047
estuaries, however, future outlook was not determined for 67 systems, illustrating
1048
uncertainty in the future outlook conclusions. There are presently 12 systems with a
1049
moderate low to low overall eutrophic condition but moderate to high susceptibility
1050
and a worsening future outlook. Although many systems are predicted to worsen as
1051
population in the coastal zone increases, the systems with low eutrophic conditions
1052
and moderate to high susceptibility should be a priority for protective management
1053
because they are the most at risk. The potential for changes in nutrient loads and
1054
hydrology due to climate change should also be explored when considering the
1055
future estuarine health.
1056
1057
ASSETS: The assignment of an ASSETS rating requires data for all three
1058
components and there were adequate data for determination of an ASSETS rating
1059
for only 48 systems. Only one system was rated as high quality (Connecticut River),
1060
while five were rated as good (Biscayne Bay, Pensacola Bay, Blue Hill Bay, Sabine
47
1061
Lake, Boston Harbor). Eighteen were rated as moderate and 24 systems were rated
1062
as poor or bad. The single rating of ASSETS allows simple comparisons between
1063
and among systems. It has been applied at a national (this study) and international
1064
level (http://www.eutro.org). The intent is to share lessons learned and encourage
1065
more pro-active approaches to the protection and maintenance of estuarine health
1066
here in the US and globally.
1067
1068
A Human use Indicator for Barnegat Bay: Barnegat Bay is an excellent candidate
1069
for the application of recreational fish catches as a human-use indicator. Surrounded
1070
by a large population center, Barnegat Bay sees a lot of recreational fishing activity.
1071
Barnegat Bay is also frequented by a variety of recreational species targeted by
1072
fishermen. Through the Marine Recreational Fisheries Statistics Survey (MRFSS),
1073
the National Marine Fisheries Service regularly conducts surveys of recreational
1074
fishing activity and success in most U.S. estuarine systems. Salinity, temperature,
1075
and dissolved oxygen data for Barnegat Bay sampling stations were averaged by
1076
month and year and then matched to the month and year of the fishing trip from the
1077
MRFSS data.
1078
For Barnegat Bay, summer flounder, the most sought after species, is a good
1079
indicator of the human use impacts of eutrophication. The solid line in Figure 7
1080
shows the average actual catch of summer flounder in a month for the period from
1081
1997-2002. The statistical model was then used to predict summer flounder catches
1082
under different water quality conditions. Specifically, an upper limit on chlorophyll
48
1083
a concentrations was set so that they could not exceed the sample averages of 7.12
1084
μg L-1, and a lower limit on dissolved oxygen of 6.51 mg L-1. The dashed line in
1085
Figure 7 represents the predicted summer flounder catches under these improved
1086
water quality conditions, and the distance between the two lines is the impairment
1087
due to eutrophication. In some months, the limits are rarely exceeded and there is no
1088
difference in expected catches. Overall, the average catch of summer flounder is
1089
reduced from the predicted average of 1.25 fish per trip to 0.92 fish per trip, a 26%
1090
reduction.
1091
1092
EPA National Coastal Assessment
1093
The NCCR2 summarized results by region and show that overall the
1094
condition of estuaries in the US is fair. Figure 8 shows the results for the combined
1095
indicators for all regions, though each indicator and region can also be seen
1096
individually. Only one of the five indicators of estuarine condition received a poor
1097
overall rating, the coastal habitat index. The water quality index, the equivalent to
1098
the overall eutrophication condition component of the NEEA/ASSETS, and the fish
1099
tissue contaminants index received a fair rating, and the benthic index and sediment
1100
quality index were rated fair to poor (Figure 8 summarizes U.S. estuarine condition).
1101
These ratings are based on samples collected at 2,073 estuarine sites in the
1102
conterminous 48 states between 1997 and 2000 (about 90% of the samples were
1103
collected in 1999 and 2000). Of the five summary indicators (water quality index,
1104
sediment index, benthic index, coastal habitat index, and fish tissue contaminants
49
1105
index), only the fish tissue contaminants index was rated good for any region of the
1106
United States. The water quality index is rated fair throughout the estuaries of the
1107
United States, although estuarine waters in the Northeast Coast region appear to
1108
have poorer water quality conditions than those in other regions of the country. The
1109
sediment index is poor in Northeast Coast and Puerto Rico estuaries and in the Great
1110
Lakes; borderline fair in West Coast estuaries; fair in the Gulf Coast estuaries; and
1111
borderline good in Southeast Coast estuaries. The benthic index shows that
1112
conditions are poor in the Northeast Coast and Puerto Rico, borderline fair in the
1113
Gulf Coast and Great Lakes, and fair in the Southeast Coast and West Coast.
1114
Condition as measured by fish tissue contaminants is poor in Northeast Coast and
1115
West Coast estuaries and fair to good in the remainder of the country. More
1116
specifically, 21% of estuarine area in the United States (excluding the Great Lakes)
1117
is unimpaired for human and/or aquatic life uses (Figure 9). About 28% of estuarine
1118
area is impaired for aquatic life use, 22% is impaired for human use, and an
1119
additional 44% is threatened for both uses. Impaired aquatic life use was indicated
1120
by lower-than-expected biodiversity, increased abundance of pollution-tolerant
1121
species, decreased abundance of pollution-sensitive species, poor water quality
1122
condition, poor sediment quality, and coastal wetland losses. Impaired human use
1123
was defined as exceedances of fish tissue contaminant risk-based guidelines for
1124
consumption (based on four 8-ounce meals per month). Threatened use is equivalent
1125
to fair overall condition for any of the indicators.
1126
50
1127
Other Benthic Index/Indices……another place holder for results that can then be
1128
compared to the benthic index in figure 8.
1129
1130
Discussion
1131
As illustrated above, several legislations were at present created in USA and
1132
Canada, Australia, European Union, South Africa, and China intending to assess the
1133
ecological quality or integrity, within estuarine and coastal systems. Most of these
1134
legislations tend at present to define environmental water quality in an integrative
1135
way, requiring the ecological status to be assessed at the ecosystem level. For this
1136
reason, there is an increasing interest in developing assessment tools for different
1137
physico-chemical or biological ecosystems’ elements, although very few
1138
methodologies integrate all these elements into a unique evaluation of a water body.
1139
In practical terms, from the management point of view, decision makers need to get
1140
simple but scientifically well grounded methodologies, susceptible of showing to the
1141
general public the evolution of a zone (estuary, coastal area, etc.), taking into
1142
account human pressures or recovery processes. In this scope, a major scientific
1143
challenge is to develop suitable tools, which may allow defining adequately the
1144
current condition of marine ecosystems and bioregions in terms of biological
1145
performance and monitor their changes through time.
1146
Among such tools, ecological indicators are widely used to supply synoptic
1147
information about the state of ecosystems. Most often they address ecosystem’s
1148
structure and/or functioning accounting for a certain aspect or component, for
51
1149
instance nutrient concentrations, water flows, macro-invertebrates and/or vertebrates
1150
diversity, plants diversity, plants productivity, erosion symptoms and, sometimes,
1151
ecological integrity at a system’s level. In general, the main attribute of a good
1152
ecological indicator is to combine numerous environmental factors in a single
1153
value,, which might be useful in terms of management and for making ecological
1154
concepts compliant with the general public understanding. But the application of
1155
ecological indicators is not exempt of criticisms, the first of which is that the
1156
aggregation of results in oversimplification of the ecosystem under observation and
1157
are not efficient in capturing qualitative modifications resulting from the emergence
1158
of new characteristics arising from self organisation processes (e.g. in a global
1159
climate change scenario). Moreover, problems arise from the fact that indicators
1160
account not only for numerous specific system characteristics, but also other kinds
1161
of factors, e.g. physical, biological, ecological, socio-economic etc. Thus, indicators
1162
must be utilised following the right criteria and in situations that are consistent with
1163
its intended use and scope; otherwise they may drive to confusing data
1164
interpretations.
1165
In view of management needs, what might be the characteristics of a good
1166
ecological indicator?, What kind of information, regarding ecosystem responses, can
1167
be obtained from the different types of biological and physicochemical data usually
1168
taken into account in evaluating the state of coastal areas and transitional waters?
1169
The analysis of the current situation in different continents shows that these are still
1170
pertinent questions. In fact, coastal shallow water ecosystems are extraordinary
52
1171
valuable in terms of services and goods, correspondingly subject to strong human
1172
impact, and at the same time extremely vulnerable to climate change. Consequently,
1173
appropriate, accurate, and efficient information on ecosystem status and trend
1174
constitute a prerequisite of sustainable development of marine ecosystems and
1175
resources, including not only environmental protection but also economic growth
1176
and social welfare (see for instance the Lisbon Agenda).
1177
Measures and indicators presently proposed, which focus on the living part
1178
of aquatic systems and its “supporting” hydro-morphological, chemical and
1179
physicochemical elements to evaluate aquatic ecosystem health and fitness, are
1180
reasonably able to describe the current status of an ecosystem in terms of actual
1181
criteria as outlined in the European Water Framework Directive (WFD,
1182
2000/60/EC). In this sense, for the time being, they fulfil what appear to be needs of
1183
managers and decision-makers.
1184
But with regard to coastal and transitional waters, biological elements are
1185
limited to composition, abundance (and biomass) of phytoplankton, other aquatic
1186
flora, benthic invertebrate fauna, and fishes. The measures used to quantify these
1187
biological
1188
distributions/gradients, ratios, biodiversity indices and classification schemes
1189
(Indicators and Methods for the Ecological Status Assessment under the Water
1190
Framework Directive, EUR 22314EN). Modelling approaches, for instance, are still
1191
rare and limited to sub-systems.
elements
can
be
categorized
broadly
in
tools
describing
53
1192
Therefore, measures used today may provide “snapshots” of given
1193
ecosystem structural properties, but provide lithe or no information about ecosystem
1194
functioning. In fact, the causal links between the measured quantities/qualities and
1195
underlying ecosystem functioning remains largely uncertain or even unknown
1196
(introduce here a figure – scenarios analysis - refer here consequences in terms of
1197
the Ecosystems Services and Goods Approach). Will this be sufficient in the near
1198
future? Will it be good enough to achieve for instance the objectives of the
1199
European future maritime policy, which indicates strong demand for higher
1200
integrated and more holistic approaches towards the sustainable development of the
1201
European marine heritage (a typical back-cast scenario approach)? And what about
1202
the magnitude and speed of anticipated changes related to global change combined
1203
with the increasing intensity and multitude of marine environment use by man?
1204
In order to fulfil future needs, environmental science must complement the
1205
“static” look at structural ecosystem properties through an approach towards the
1206
ecosystem function and dynamics, which will provide a sounder and reliable basis
1207
for successful management strategies. Common reductionistic approaches can only
1208
partially cope with ecosystem complexity that arises from its large number of
1209
components, interactions and its spatio-temporal dynamics. Inevitably, we must
1210
recognize that the whole behaves differently from the sum of its parts, and thus
1211
neither examining a small subsystem nor reduction to simple relationships are
1212
sufficient approaches to understand ecosystem functioning. What happens is that
54
1213
specific qualities/features/properties emerge at the ecosystem level, and must be
1214
related to ecosystem functioning.
1215
Trophic interactions (who eats whom) of organisms, although governed and
1216
modulated by external boundary conditions, appear as the most prominent and
1217
significant type of ecological relationships. A way to access the ecosystem level is
1218
looking upon it as a network of such trophic species-to-species interactions.
1219
Through holistic approaches, for instance Ecosystem Network Analysis (ENA),
1220
properties of both ecosystem network structure and network flow (of matter and
1221
energy) can be explored with respect to ecosystem functioning (e.g. overall system
1222
stability and resilience) and general principles might be derived in the relationship
1223
between ecosystem network, ecosystem functioning, and ecosystem services and
1224
goods, based on a comparison of many ecosystems using a standardized approach.
1225
This is a possible vision of tomorrow.
1226
1227
Conclusions and recommendations
1228
(Taking into account the discussion, we can propose several recommendations for
1229
future developments)
1230
1231
55
1232
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64
Table 1. Classification of sea water quality used in CIAM, including organic
pollution assessment, TPH (Total Petroleum Hydrocarbon), and heavy metal
pollution (adapted from Jia et al., 2003); and comprehensive ecological
quality grade of marine fishery environment (Jia et al., 2003).
Organic pollution
assessment
TPH and heavy
metal pollution
assessment
Comprehensive
ecological quality
grade of marine
fishery
environment
Quality index (A value)
<0
0~1
1~2
2~3
3~4
>4
Pi
<0.4
0.4~0.6
0.6~0.8
0.8~1.0
1.0~2.0
>2.0
Index range
0.2
0.2~0.4
0.4~0.6
0.6~0.8
0.8~1.0
>1
Grade
1
2
3
4
5
6
Grade
1
2
3
4
5
6
Grade
1
2
3
4
5
6
Quality Assessment
Excellent
Clean
Relatively clean
Slight pollution
Medium pollution
Serious pollution
Quality Assessment
Background
Clean
Relatively clean
Slight pollution
Pollution
Serious pollution
Quality status
Excellent
Fine
Relatively fine
Moderate
Poor
Very poor
65
Table 2. Grade of primary productivity and diet organism richness (Jia et al., 2003).
Item
Status
Index level
Primary productivity
(mg C mg-1 day-1)
Phytoplankton
(104 ind m-3)
Zooplankton (mg m-3)
Benthos (g m-2)
Grade
1
Low
2
Relatively low
3
Medium
4
Relatively high
5
High
6
Super high
>1.0
<200
1.0~0.8
200~300
0.8~0.6
300~400
0.6~0.4
400~500
0.4~0.2
500~600
<0.2
>600
<20
20~50
50~75
75~100
100~200
>200
<10
10~30
30~50
50~75
75~100
>100
<5
5~10
10~25
25~50
50~100
>100
66
Table 3. ASSETS application to Jiaozhou Bay.
Index Method
IFa
Susceptibility
Indicator
Dilution
potential
Flushing
potential
Nutrient inputs
PSMc
OECb
d
SSM
FOe
Future nutrient
pressure
Chlorophyll a
Macroalgae
Dissolved
Oxygen
SAV loss
Nuisance and
toxic blooms
Level of
expression
Moderate
Moderate
High
Low
No problem
Low
Index result
ASSETS
score
Low
(due to intense
shellfish
aquaculture)
High
Low
Low
Low
Decrease
Improve low
IF – Influencing Factors; bOEC – Overall Eutrophic Condition index; cPSM – Primary
Symptoms Method; dSSM – Secondary Symptoms Method; eFO – Future Outlook index
a
67
Table 4. Summary of comparison among “Phase I/II” methods (adapted after
Bricker et al., 2006).
Methods
Nutrient
Index I
Nutrient
Index II
OSPAR
COMPP
EPA NCA
ASSETS
Indicator criteria/
thresholds
Modified after
Not specified
Japanese criteria
Modified after
Not specified
Japanese criteria
Growing season, Individually/regionally
winter for
determined reference
nutrients
condition
Determined from
Summer index
American national
periods
studies
Determined from
Annual cycle
American national
studies
Temporal focus
Combination method
Sum of four ratios
Ratio of three indicators to
their threshold values
Integration of scores for
four categories
Ratio of indicators:
good/fair indicators to
poor/missing data
Average of primary and
highest secondary are
combined by matrix
68
Table 5. Quality indices of the fishery environment of northern South China Sea (according to Jia et al., 2005).
Waters
DIN
PO43--P
A
E
DO
Primary
productivity
Phytoplankton
Zooplankton
Benthos
Northern South China Sea
0.27
0.55
0.23
0.34
0.72
0.58
0.53
0.88
0.78
Taiwan bank waters
0.29
0.52
0.20
0.28
0.67
0.50
0.37
0.75
0.68
East Guangdong waters
0.31
0.61
0.21
0.30
0.78
0.54
0.79
0.88
0.81
Pearl River estuary
0.26
0.58
0.23
0.37
0.74
0.68
0.55
1.00
0.91
West Guangdong waters
0.28
0.54
0.25
0.47
0.72
0.58
0.58
0.88
0.77
Southern waters of
Hainan
0.24
0.50
0.20
0.30
0.73
0.64
0.36
0.87
0.82
Beibu Bay
0.24
0.45
0.20
0.30
0.66
0.72
1.00
0.94
0.94
Comprehensive index
0.58
Relatively fine
0.51
Relatively fine
0.61
Moderate
0.64
Moderate
0.60
Relatively fine
0.55
Relatively fine
0.66
Moderate
69
Table 6. Decision tree when integrating water, sediments and biomonitors in assessing
chemical status, within the Water Framework Directive. Note: a variable ‘meet’
when the concentration is under the quality objectives established by the
Directive (Table modified and adapted from Borja et al. (2006)).
WATER
All variables meet
SEDIMENTS
All variables meet
1 variable does not meet
>= 2 variables do not meet
All variables meet
1 variable does not meet
1 variable does not meet
>= 2 variables do not meet
>= 2 variables do not meet
BIOMONITORS
All variables meet
No data
>=1 variable does not meet
STATUS
Meet
Meet
Not Meet
Meet
Meet
Meet
Not Meet
Not Meet
Not Meet
70
Table 7. Metrics used in assessing the biological elements quality, within the Water
Framework Directive, after Borja et al. (2004a). AMBI: AZTI’s Marine Biotic
Index (Borja et al., 2000); spp: species.
Phytoplankton
Chlorophyll a
Species composition
Number of blooms
Macroalgae
Richness
Cover of opportunistic and
sensitive spp.
Ratio green algae /other
spp.
Benthos
Richness
Diversity
AMBI
Fishes
Richness
Abundance and
percentage of resident spp.
Trophic composition
Flat fish percentage
Pollution indicator spp.
Invasive spp.
Fish health
71
Table 8. Decision tree in assessing the integrative ecological status, within the Water Framework Directive (modified from Borja et al., 2004a).
BQE: Biological Quality Element; BQ: Biological Quality; Ph-Ch: Physico-Chemical; H: Hydromorphological.
BQEs
BQ status
Ph-Ch
conditions
H conditions
Ecological
Status
Do all BQEs meet High
status?
Yes
High
Then
- Do Ph-Ch conditions
meet High status?
- Are chemical
concentrations
<background?
Yes
Do H conditions meet
High status?
Yes
High
Not
Not
- Does each BQE meet High
or Good quality?
- Does Benthos meet High
quality and from the
remainder 1 or 2 meet
Moderate quality? (1)
- Do Benthos and another
BQE meet Good quality and
the remainder Moderate? (2)
- Coast: Does Benthos meet
Good quality and other
Moderate? (3)
Yes
Good
Then
- Do Ph-Ch conditions meet
Good status?
- Are chemical quality
objectives meet?
Not
Not
- Does one BQE meet
Moderate quality (except
1,2,3)
- Does Benthos meet High
or Good quality and no one
has Bad quality?
Not
- Other combinations
- Does Benthos meet
Poor quality?
Not
- Do all BQEs meet
Bad quality?
- Does Benthos
meet Bad quality?
Yes
Moderate
Yes
Poor
Yes
Bad
Then
Then
Then
Moderate
Poor
Bad
Yes
Not
Then
Good
72
Table 9. Example in integrating ecological status of several sampling stations (St) into a
unique value, for the whole water body.
Ecological Status
Equivalence (E)
Surface (km2)
Rate (per one) (R)
TOTAL (E x R)
GLOBAL STATUS
Equivalence (E)
10
8
6
4
2
St.1
St.2
St.3 St.4
Poor Moderate Good Good
4
6
8
8
0.3
0.5
0.7
1.0
0.12
0.20
0.28 0.40
0.48
1.20
2.24 3.20
Ecological Status
High
Good
Moderate
Poor
Bad
TOTAL
2.5
1.0
7.12
Good
TOTAL (E x R)
8.4 – 10
6.8 – 8.39
5.2 – 6.79
3.6 – 5.19
2.0 – 3.59
73
Figure captions
Figure 1. Roles, responsibilities and interactions of US Agencies for estuarine and
coastal environments (EPA, NOAA, States).
Figure 2: Influencing factors of eutrophication on a national scale.
Figure 3: Overall eutrophic condition on a national scale.
Figure 4: Symptom expressions on a national scale for a) chlorophyll a and b)
Nuisance/toxic blooms (HABs).
Figure 5: Number of estuaries in each eutrophication category in the early 1990s (1999
assessment; Bricker et al., 1999) and early 2000s (Bricker et al., 2007).
Figure 6: Future outlook on eutrophic condition on a national basis.
Figure 7: Barnegat Bay monthly average summer flounder actual catch per recreational
fishing trip (solid line), and predicted catch rates under improved water quality
(WQ) conditions (dashed line).
Figure 8: Overall national and regional coastal condition between 1997 and 2000
(USEPA, 2004).
Figure 9: National estuarine condition (USEPA, 2004)
74
Fig. 1
Components of the US regulation
of nutrients for management of
eutrophication effects in estuaries
N load reductions of
Point and Non Point sources
BMP Implementation
If use is not attained
(criteria are exceeded)
water body is listed
(303d)
(305b)
reporting
Load Response Mode
Place based
Spatially explicit
Nutrient Load-Limits
TMDL
process initiated
Load Response Translator
State numeric response (or condition) criteria
Derived from reference conditions for each class of estuary or
designated use (for many water bodies these do not exist)
Response (or condition) Monitoring, Assessment, Modeling
NOAA
NEEA/ASSETS
Typology
Human Use Indicator
EPA
NCA
N Models
States &Tribes
305(b) reporting
303(d) listing
75
Fig 2
76
Fig 3
77
Fig 4
78
Fig 5
79
Fig 6
80
Fig 7
81
Fig 8
82
Fig 9
83