EPA STRIVE Programme 2007-2013
Management Strategies for the Protection of
High Status Water Bodies. A Literature Review
2010-W-DS-3
STRIVE Report
University of Dublin, Trinity College
Authors
Kenneth Irvine and Emer Ní Chuanigh
Environmental Protection Agency
ENVIRONMENTAL PROTECTION AGENCY
An Ghníomhaireacht um Chaomhnú Comhshaoil
PO Box 3000, Johnstown Castle, Co.Wexford, Ireland
Telephone: +353 53 916 0600 Fax: +353 53 916 0699
Email: [email protected] Website: www.epa.ie
ACKNOWLEDGEMENTS
This report is published as part of the Science, Technology, Research and Innovation
for the Environment (STRIVE) Programme 2007-2013. The programme is financed by
the Irish Government under the National Development Plan 2007-2013. It is
administered on behalf of the Department of the Environment, Heritage and Local
Government by the Environmental Protection Agency which has the statutory function
of co-ordinating and promoting environmental research.
DISCLAIMER
Although every effort has been made to ensure the accuracy of the material contained
in this publication, complete accuracy cannot be guaranteed. Neither the
Environmental Protection Agency nor the author(s) accept any responsibility
whatsoever for loss or damage occasioned or claimed to have been occasioned, in part
or in full, as a consequence of any person acting, or refraining from acting, as a result
of a matter contained in this publication. All or part of this publication may be
reproduced without further permission, provided the source is acknowledged.
The EPA STRIVE Programme addresses the need for research in Ireland to inform
policymakers and other stakeholders on a range of questions in relation to
environmental protection. These reports are intended as contributions to the necessary
debate on the protection of the environment.
EPA STRIVE PROGRAMME 2007-2013
Published by the Environmental Protection Agency, Ireland 4
PRINTED ON RECYCLED PAPER
Details of Project Partners
Kennenth Irvine
School of Natural Sciences, Trinity College, Dublin 2
Tel.: 01 8961366
E-mail: [email protected]
i
Executive Summary
The Water Framework Directive (WFD; 2000/60/EC) requires EU Member States to categorise their
water bodies across a 5-point ecological status scale of high, good, moderate, poor and bad. It also
requires that Member States identify water bodies that have been minimally impaired by
anthropogenic pressure. These are the reference sites from which all other sites are compared in
order to estimate an Ecological Quality Ratio (EQR) based on observed state compared with
reference. So that water bodies with similar natural geological and landscape settings are compared
with each other, an early stage of the WFD was agreement on a water body typology, which enabled
identification of type-specific reference sites. While there is ongoing debate on the determination of
reference state, it provides a baseline against which monitored sites can be compared. Within the 5point classification scale of the WFD, reference sites represent the upper end of the scale of high
status sites.
Owing to historical low intensity land use and low density human population Ireland retains a large
number of high status water bodies. However, long term monitoring of rivers by the EPA has shown a
dramatic and continuous decline of these sites over the last 20 years. High status rivers equate to an
EPA river monitoring score of Q4/5 or 5. Since 1987, high status sites declined from almost 30% of
those sampled to 17% for the period 2006-8. So, while Ireland still retains a large number of high
status sites this monotonic decline in their number is cause for considerable concern. Such extensive
and long-term monitoring is not available for other water bodies, although less than 30% of 35
putative reference lakes identified by the EPA were confirmed as such by palaeolimnology. The
network of high status water bodies are clustered and negatively related to intensive agriculture. Land
use intensification is associated with impact on water resources globally. The dramatic decline of
high status sites suggests, however, effects from small-scale intensification and other, quite localised,
impacts. The most general effect on Irish freshwaters is from nutrient enrichment, and low quantities
of nutrients entering a waterbody can have a significant and negative impact. Other effects can arise
from increased sediment load, alterations in drainage and chemical pollution, including acidification
from conifer forestry in areas with low buffering to changes in pH.
Across Europe there has been a focus on achieving good ecological status, and while decline across
status classes is in breach of the WFD, there has been a general lack of attention to the mechanisms
to protect high status water bodies. A reliance on protected areas designated under the Habitats
Directive (92/43/EEC) to protect high status waters is unlikely to be effective, as the targets and
mechanisms to harmonise objectives with the WFD are not effective. In Ireland the Habitats Directive
has, in any case, failed to adequately protect designated sites, and only about 35% of high status
sites coincide with the candidate Special Areas of Conservation (SAC) network.
Specific national
legislation, The European Communities Environmental Objectives (Surface waters) Regulations S.I.
No 272 of 2009, designed to effect the requirements of the WFD requires that there is no deterioration
from high status water bodies to a lower classification, but with the exception of European
Communities Environmental Objectives (Freshwater Pearl Mussel) Regulations 2009 (S.I. No 296 of
ii
2009), the there are no specific mechanisms to protect high status water bodies. The administrative
procedures for River Basin Management under the European Communities (Water Policy)
Regulations (S.I. No. 722 of 2003, and subsequent amendments) rely heavily on a local authority
lead in implementation of the WFD, but this is hindered through lack of resources and, probably more
so, by a widespread fragmentation of water governance. Updated planning legislation under The
Planning and Development (Amendment) Act 2010 strengthens the relationship with the WFD,
providing a clearer requirement for local authorities to consider potential impacts on high status water
bodies. While better planning for development including one-off housing and associated water
treatment is required it does not address low level and localised impact from land use.
Impacts from agriculture on water quality are regulated by the European Communities (Good
Agricultural Practice for Protection of Waters) Regulations S.I . No 610 of 2010 (following a series of
other Regulations dating back to 2006).
These focus on cross compliance with the WFD and,
especially for the establishment of a Nitrates Action Programme in compliance with the Nitrates
Directive (91/676/EEC).
The Nitrates Action Programme is designed so that modern farming is
compatible with WFD compliance for good status. It is not designed to protect high status water
bodies. Indeed it is likely to lead to increased pressure on these sites because inter alia there will be a
need for more extensive spreading of animal waste and a general premise of a phosphorus soil
content commensurate with optimal (i.e. maximal) agricultural production.
A policy to maximise
agricultural production is not compatible with the protection of high status sites or water bodies.
Recent agricultural policy initiatives, under Harvest 2020, to increase dairy production by 50%
accentuates potential impact on all waters, but especially those at high status if animal waste is
allowed to be exported across catchments.
Harvest 2020 was not subject to a Strategic
Environmental Impact Assessment, which may be in breach of Directive 2001/41/EC.
As well as the Habitats and Nitrates Directives, nine others are listed in Annex VI Part A which are to
provide Basic Measures for implementing the WFD. The Birds Directive (79/409/EEC) links with the
Habitats Directive in providing for the Natura 2000 network of sites. Site conditions for protection of
bird numbers are not necessarily synonymous with those of wider ecological quality. Indeed many
aquatic bird populations, especially those associated with estuaries may benefit from moderate
nutrient enrichment. Licensing under Directive 96/61/EC on Integrated Pollution Prevention and
Control accounts for quality, and vulnerability, of receiving waters and, along with licensing under
Urban Waste Water Treatment Directive (91/271/EEC), are likely to be to too stringent for any future
industrial or waste water emissions to high stats water bodies.
The Sewerage Sludge Directive
(86/278/EEC) under the Waste Management Act, 1996. S.I. No. 148 of 1998, and its amendment of
2001, limits the heavy metal content of sludge spread on land and needs to have regard to pH and
nutrient content of receiving soils. In this way it links with the Nitrates Regulations of 2010. Local
authorities have developed sludge operational management plans for the management of sludge
arising from all sources including waste water treatment plants, septic tanks, industry and agriculture.
The licensing condition, and location of spreadlands should be clearly identified and available to local
authorities and the EPA.
iii
The Environmental Impact Assessment Directive (85/778/EEC) as amended by Directive 98/83/EC,
links with the new Planning Act 2010. There is an onus on requiring Environmental Impact
Assessment for developments, including forestry and some agri-development projects that may
impact on high status water bodies. This requires greater integration between local authority and the
River Basin Management Plans (RBMPs). It also requires increased awareness with planning
departments of potential impact of small developments on sensitive sites and a more integrated use
of information held by different State bodies. The remaining Directives listed in Annex VI Part A of
the WFD have generally less significance for protection of high status sites than those discussed
above. The Bathing Water Directive (76/160/EEC), as amended by Directive 2006/7/EC, is focussed
on protection of the public from faecal bacteria contamination. Failure to reach a bathwater standard
can, however, indicate sources of pollution, and act as a check for high status designation. The
Drinking Water Directive (76/160/EEC, as amended by Directive 98/83/EC), Plant Protection Products
Directive (91/414/EEC), and Major Accidents (Seveso) Directive(96/82/EC)
have no particular
aspects that relate specifically to high status protection. Drinking water extracted from high status
waterbodies . Water extracted from high status sites would be expected to low in nutrients and
pathogens, therefore, requiring a low level of treatment. The Plant Protection Products Directive
(91/414/EEC) requires Member State authorisation for plant protection products (PPPs) to provide a
safeguard for human health and the environment. Harmful substances are subject to a maximum
allowable concentration (MAC), listed in Table 11 of the European Communities Environmental
Objectives (Surface waters) Regulations S.I. No
272 of 2009. The Major Accidents (Seveso)
Directive (96/82/EC) requires provision to be made for emergencies, including unplanned emissions,
for major industrial facilities. In Ireland these would tend be of potential risk more for coastal than
inland sites.
In summary, the existing Directives that are there to provide Basic Measures for implementation of the
WFD fail to provide a satisfactory safeguard for high status water bodies. Supplementary Measures
as described in Part B of Annex VI of the WFD can be developed, but only one measure that relates
specifically to high status sites has been proposed. This, for the pearl mussel, provides a useful
precedent for the process required for other possible measures that are needed for protection of high
status water bodies.
If the current WFD driven Regulations and existing Directives do not provide adequate protection for
high status sites, as evinced by their steady decline, it suggests new approaches are required to meet
national expectations and international obligations to protect the best quality aquatic systems in the
country. The INSPIRE Directive 2007/2/EC, establishing an Infrastructure for Spatial Information in
the European Community should help support a more consolidated and integrated approach for water
policy in general, as it should be self-evident that use of spatial information held on Geographical
Information Systems (GIS) is a national resource that should be freely available and interchangeable
across all relevant State-funded bodies. The better use of information is unlikely, however, to address
the highly fragmented structure of the country's water governance.
iv
The main and most widespread policy driver of environmental degradation in Europe is the Common
Agriculture Policy (CAP), and it is its reform that provides the greatest opportunity for high status
protection and, where feasible, restoration. To date agri-environmental schemes funded through the
CAP have had a disappointing effect, and widespread decline of biodiversity across all habitats
continues. Europe has failed to meet international obligations under the Convention of Biological
Diversity and internal policy to halt the decline of biodiversity. There is little evidence that the Rural
Environmental Protection Scheme (REPS) had a positive impact on biodiversity in general. An
impression that REPS improved water quality cannot be quantified as its effects were not monitored.
This general failure of demonstrating a positive return for the investment Ireland received for this has
drawn criticism from the OECD. REPS has now been replaced by the agri-environment options
scheme (AEOS). The only water - protection measure in the AEOS is the limiting of animal access to
water courses, and provision of drinking troughs. The AEOS funding is also prioritised for Natura
habitat and/or Non-Natura Commonage. The AEOS scheme reflects, therefore, two important factors
relevant to farming that may affect high status sites. First, there is a low priority for fiscal support for
farms outside Natura 2000 sites, meaning that most high status sites are unlikely to benefit from the
AEOS unless there is a generally low uptake of the scheme. Second, it reflects the failure of the cSAC
management network to protect habitats because of insufficient direct investment into the
management of these sites. The prioritisation of Natura 2000 sites under AEOS has the potential to
increase pressure on high status sites outside that network.
General rhetoric for the next round of CAP reform in 2013 is that it provides an opportunity to provide
greater habitat protection in agricultural landscapes. This is only likely to be realised if more money of
the pillar 1 (which funds the direct payments) of the CAP is used for enhancing environmental
protection. This can also help reduce widespread contradictions and confusion in agricultural and
environmental policy. More carefully targeted use of CAP pillar 2 (used for rural development) can
provide fiscal incentives for high status water body protection.
For this to have any chance of
developing policies to protect high status sites it requires urgent and immediate dialogue between the
EPA, its parent Government Department and the recently reconfigured Department of Agriculture,
Food and the Marine.
Planning and potential CAP reform are also important for forestry and its relationship with high status
sites. Forestry can impact water resources and there is a legacy in Ireland of inappropriate settings
and management of commercial forestry. New forest guidelines, and a new Forestry Act pending
should provide greater integration with the goals of the WFD. There are also possibilities for more
positive impacts of forestry through restructuring of grants to promote environmentally sensitive
forestry. As with agriculture in general, there is a need for better liaison with the EPA and local
authorities and more comprehensive use of shared GIS to guide planning. However, the current
approach is focussed on compatibility with achieving good status of water bodies, and more
innovative approaches are needed to support and promote protection of high status sites, many of
which are in upland areas and prone to impact from low to moderate disturbance. Consideration is
needed to the banning of new plantations that may impact high status water bodies and, for maturing
v
forests, harvesting limited sized coups, with strict adherence to best practice guidelines. The premise
that felled forest is replanted requires a re-evaluation. There are opportunities for enhancing
protection of high status sites through promotion and fiscal support for riparian buffer strips, including
wet forest which currently is not considered commercially viable, utilising existing mechanisms
allowed through the CAP, and integrating new forest areas within the landscape with the goal to
attenuate nutrient mobility.
Within an appropriate policy framework there are mechanisms that can improve the protection of high
status water bodies. Without such a framework, mitigation and preventative measures remain largely
theoretical and discursive. Such a policy framework also requires mechanisms for active participation
and financial incentive for stakeholders. This requires schemes targeted to high status sites, and
those where restoration back to high status may involve low cost. Preventing deterioration of high
quality sites is almost certainly a more cost effective strategy than large-scale restoration of seriously
impacted ones, although in the long-term current policy is to rely on the Nitrates Action Programme to
fulfil that requirement.
Examples from Ireland and abroad provide useful case studies which can be used to protect high
status water bodies. These are most effective when targeted to local situations, involving close liaison
and building trust between advisory services and stakeholders and providing enabling finances.
Costs, however, can be modest when compared with current agricultural subsides, especially under
pillar 1 of the CAP. Change to management occurs when stakeholders are aware of a problem, and
have the knowledge and resources to address it. In Ireland valuable lessons are to be learnt from the
BurrenLIFE project and the Lough Melvin catchment management programme. In both of these
programmes there was a concerted effort for intense engagement with local farmers, including field
visits and discussion groups. Multi-criteria decision support techniques were shown to be important.
This reflects international experience. The BurrenLIFE project involved a diverse partnership including
farming and conservation interests. It also used CAP funds supported by those from government to
establish low intensity farming with the specific objective for enhancing grassland biodiversity. A key
aspect was demonstrating the local interest in traditional low-impact farming required for the
maintenance of the Burren species rich grasslands.
The two major threats to this are land
intensification on the one hand and land abandonment on the other. Utilising CAP funds the project
has been extended into a second phase, with a target of 100 participating farmers.
The Lough Melvin catchment programme identified the importance of small patches of nutrient rich
land as a contributing factor to nutrient export, and explored a range of management and fiscal
measures to mitigate impact. This included landuse auctions where farmers bid for available funds
based on plans for environmentally beneficial land use.
Currently there is no policy framework to
capitalise on these types of initiatives, which have been used successfully in the US and trialled in
Australia. Ireland’s centralised structure to agri-environment schemes, compared with many other EU
states, is not conducive to “bottom up” participation, and the fact that the Melvin initiative occurred
outside the existing governance structures meant is difficult to follow up on the study’s
recommendations.
vi
Regulation by itself, as done for example through the Nitrates Regulation (see Section 3.3.1), can
also only go so far in achieving environmental objectives, even if the underlying principles are sound.
For the protection of high status sites a need for more local and incentive based schemes are
essential. A strong public and stakeholder participatory process provides a foundation for a structured
approach to site protection. It is only through local engagement that the appropriateness of local
management options can be assessed. These sort of mitigation strategies are used widely in the
U.S., for example in the protection of ‘Outstanding National Resource Waters’ (ONRWs) under the
Clean Water Act (CWA)1974. Under the CWA, maintenance of water quality centres around
designated use as a public resource. Once a designated use is established for a water body, the
State develops water quality criteria considering a water’s assimilative capacity for different levels of
pollution, defined as total maximum daily loads (TMDLs). The use of a TMDL approach s
recommended for Irish water bodies and can be of particular value for protection of those at high
status. Current water quality nutrient standards for phosphorus in Irish rivers are likely incompatible
with the maintenance of standards in receiving lakes.
A common strategy to prevent impact to ONRWs in the U.S. is the use of buffer strips. The minimum
buffer width for sensitive stream mitigation projects is 50ft (ca 15 m). This compares with buffer strips
of 2 m under Irish Good Agricultural Practice Regulations of 2010, illustrating a need for a different
approach of for status water bodies. It is, however, recognised that the effectiveness of riparian buffer
strips can be highly variable and dependent on design and local conditions. They are also used
effectively in a suite of measures that protect the source water of New York City, that also involves
management agreements with farmers for low intensity farming. Other commonly used measures
internationally include, stock holdings, managing hot-spots of sediment and nutrient emissions and
attenuating water movement through local wetland creation. Such a suite of measures are used in
New Zealand to reduce nutrient emissions from, like Ireland, a predominantly grassland agriculture.
There is sufficient knowledge to establish locally effective strategies for the protection of high status
water bodies. For high status waters the key to success is adoption of locally relevant strategies and
identification and monitoring for possible small spatial-scale impacts. This can be supported with
local knowledge and interest. The U.K. River Trusts provide a useful model. This includes liaison with
landowners and development of site-specific and cost-effective, management planning. There are a
number of similar bodies in Ireland, although tending to be less formal or financially secure. There is,
however, the potential for development of expertise and professionalism within these bodies. They
could provide a local conduit for implementing effective management.
In conclusion, protection of high status waters lacks an effective policy framework and requires a
concerted interest, and development of effective protection mechanisms across all the relevant State
bodies, working closely with the River Basin Management Plans. High status sites should be afforded
the same level of protection as protected habitats, and there is, therefore, an obvious need to
consider further the relationships between the Habitats Directive and the WFD. More ambitious
planning would consider establishing a connected network of high status sites across the country.
This could also link with policies for rural development, international obligations under the Ramsar
vii
Convention and provide wider conservation benefit. For this to occur requires more effective
administrative structures under the WFD, greater resolution of common purpose between branches of
government charged with the protection farming interests and those of the environment. It also
requires locally focussed stakeholder engagement and a redirection of funding targeted to be fit for
purpose. Relying on current policies and structures to protect high status water bodies is unlikely to
be effective.
viii
“Protection against losses needs to be seen to have a value in the same way as
exploitation of goods” (Newson, 2010).
Section 1: Monitoring and determination of high status sites
1.1 Introduction
In 1971 An Foras Forbatha (the forerunners of today’s EPA) started a national monitoring programme
of Irish rivers, using a method based primarily on invertebrate communities. The traditions of this
method can be traced to the “saprobic” index used in central Europe from early in the 20th century to
assess effects of sewerage outfalls on river health, and developed for more general use by the U.K.
Trent River Board (Woodiwiss, 1964) and the more widely adopted BMWP scores (Chesters, 1980), a
forerunner of the U.K. RIVPACs (Wright et al. 1998, 2000; Clarke et al., 2003) developed from the
1980s.
The Irish assessment scored rivers on the basis of a Quality score, the Q value, with
maximum scores of Q5 representing excellent river water quality, progressively declining to Q1 with
an invertebrate community dominated by species highly resistant to depleted oxygen concentrations.
Much later the Q-value scoring system was demonstrated for its positive association with fish
communities (Champ et al., 2009). Since the start of river monitoring in 1971 the channel length
assessed was gradually increased to the current baseline of 13 200 km by 1994. This extensive data
series, following the same methodology over 30 years and often involving the same personnel,
provides an excellent barometer of the condition of Irish rivers and a foundation for river monitoring
under the Water Framework Directive (WFD; 2000/60/EC). Under the WFD, high ecological status
equates to a Q-value of 4/5 or 5.
While the Q-value network has been modified slightly to accommodate the type-specific monitoring
requirements of the WFD (McGarrigle and Lucy, 2009; EPA, 2010), this has not detracted from the
value of a long-term data set on the quality of Irish rivers. From the early 1970s to the late 1990s,
successive three year reporting periods showed an overall decline in the quality of Irish rivers (EPA,
2002). Since the early 1990s, while there has been a reduction in the decline of what the EPA term
“unpolluted” (EPA, 2010), the decline of the best quality (high status) river channels has continued.
The percentage of high quality river sites almost halved between 1987 and 2008 and there was a
seven-fold decrease in rivers attaining a Q5 (Lucy, 2009). In each survey period since 1987 the
decline in high status sites has continued, from almost 30 per cent of the total sampled in the 19871990 period to less than 17 per cent in 2006-2008.
Similar high quality and long-term monitoring data for other water bodies is not available. The
monitoring of lakes prior to the WFD was very sporadic and, in the main, focussed on those lakes for
which
there
were
perceived
water
quality
problems
(reviewed
in
Irvine
et
al,
2001).
Palaeolimnological investigations suggest variable timing of onset of water quality impact across Irish
1
lakes (Taylor et al., 2006). The decline in lake water has, however, been acknowledged since the
1970s. In 1977, C.Ó. hEcocha, Chairman of the National Science Council opened a national
conference on lakes, stating “Time is not on our side in a country in which the quality of the water of
many of our lakes has disimproved dramatically in a short number of years” (Downey and Ní Uid,
1977).
Historical record of water quality in other surface waters is even more sporadic. About half of Irish
estuaries are considered to be unpolluted (i.e of good or high status), with improvements noted in
recent years (EPA, 2010). Prior to the WFD, monitoring of estuaries and coastal waters was
conducted by a number of agencies in fulfilment of national legislation and the OSPAR Convention for
the Protection of the Marine Environment of the North-East Atlantic (1992) with little integration
towards a national programme, and involving 12 agencies (Irvine et al., 2002; EPA, 2003; Hartnett et
al, 2011). In contrast to the WFD focus, there was no requirement to monitor biological elements. This
now provides a fundamental difficulty for estimating benchmark conditions in transitional waters
(Hartnett et al., 2011). Furthermore, although the monitoring of estuaries is based on a salinity-based
typology, salinity can mimic a response to pollution. A WFD-compliant network of transitional and
coastal waters is in its infancy.
Monitoring standing water of turloughs, defined as water dependent habitats by the WFD, has also
been highly sporadic but major impact from drainage (Coxon, 1987; Drew & Coxon, 1988) and
nutrients is evident (Kilroy et al., 2001). No definition of reference condition, and hence ecological
status, has yet been agreed for these sites, which are assessed under the favourable conservation
status concept of the EU Habitats Directive. There is, however, no requirement under the WFD to set
environmental objectives for these water bodies. The general, and particularly vague EU REFCOND
guidance (European Commission, 2003a) is that these sites which are dependent on groundwater
bodies or are protected areas (all turloughs are within the former, and these designated as candidate
Special Areas of Conservation (cSAC) within the latter) “will benefit from WFD obligations to protect
and restore the status of water”.
The discussion above provides the context for this literature review, which sets out to 1) review the
relevant legislation and policies related to aquatic habitat protection and management of, especially,
high status waters; 2) review the determination and monitoring of WFD defined reference and high
status sites in Ireland and across EU Member States; and 3) assess consultative procedures that
support protection of high quality sites, including case studies from outside Ireland. Objectives 1 and
2 depend on the understanding of what is meant in the WFD by high status, and the extent that
policies protect those sites. So far, the focus of implementation of the WFD has been (under Article 4
of the WFD) that all waterbodies meet at least good status by 2015.
The WFD environmental
objective that prohibits decline of class of a water body has received far less attention.
We,
therefore, review what is meant by high status and how that is determined, before moving on to the
review of how national legislation and policies protect high status sites, including cross-compliance
with other policies.
2
1.2 High Status Waterbodies
1.2.1 Definition and Importance
Under the WFD, high status water bodies have “totally, or nearly totally, undisturbed conditions” for
each biological, physical, chemical and hydromorphological quality element. The final version of the
REFCOND guidance document for surface waters (page 17) that reference condition equals high
ecological status (CIS, 2003) has been adopted by a number of workers (Dalton et al. (2009),
although more recent discussions, that have formed part of the EU Intercalibration process, make a
distinction between “reference condition” and high status (McGarrigle and Lucey, 2009); Pado et al.,
2010). This allows setting an anchor point for reference states from which a departure from this
baseline, estimated as an Ecological Quality Ratio (EQR) can be estimated, following the logic that all
sites lie along a continuum of quality irrespective of the status class in which they are classified.
Reference sites are de facto high status, but not all high status sites will be in reference condition.
Strict screening of proposed reference sites across Europe is now proposed in terms of land use and
water chemistry. The EQRs for these sites normalise the estimation of metrics, and has been
successful in aligning EQRs across widely different water body types. While distinguishing between
reference and other high status sites has led to a more robust classification process, the relevance of
this for management is a moot point. For protection and management high status and reference
condition should be considered synonymous. Both represent the best quality sites attainable and are
extremely vulnerable to small magnitude anthropogenic pressures.
Identifying reference conditions has, however, presented fundamental difficulties that can influence
status classification (Kelly-Quinn et al., 2009). While the EU-wide Iintercalibration process has
attempted to provide a consistent approach, the setting of reference conditions remains a significant
challenge (Erba et al., 2009; Nõges et al, 2009).There are also important, and largely unresolved
discussions of what is “natural” (Boon et al., 2010) or the effect that invasive alien species (Stokes et
al., 2004; Maguire et al., 2005) have on status class. This is a particularly difficult conundrum, eliciting
divided opinions between national agencies charged with nature conservation and those with the
implementation of the WFD. This is discussed further in Section 3.2 below.
The lack of suitable reference sites in some regions (e.g Bennion et al., 2004; Borja et al., 2007) has
led to selection of ‘least impacted’ or ‘best available’ as reference sites. This is not the same as
reference state (Irvine et al., 2006), although initially a preferred option by many Member States
(McGarrigle and Lucey, 2009), and one that also influenced the decision making process in the
Republic of Ireland (Kelly Quinn et al., 2005; 2009; Dodkins et al.,2005). The RIVPACs scheme for
assessment of rivers in the U.K. is based on the best available sites, rather than the more stringent
concept implicit in the WFD (Clarke et al., 2003). The WFD Article 5 Characterisation reports prepared
by the local authorities and the EPA (Government of Ireland, 2005) identified broad pressures and
risks of water bodies to fail to meet the environmental objectives. Identifying reference sites in rivers
3
was based on sites attaining the highest quality of the EPA Q-values.
The initial selection of
reference sites in Ireland was based on expert opinion and existing data rather than a pressure
analysis as used, for example, in the EU intercalibration for stream invertebrates (Erba et al., 2009).
There is a risk that
identification of “minimally impacted” (as required by the WFD; European
Commission, 2003a) based on the biological communities rather than an assessment of minimal
pressure, risks a circular, or self-fulfilling, premise based on long-standing perceptions rather than
analysis of the risk from possible catchment impacts. This is important because understanding the
relationship between pressures and biological community structure (the dose-response relationship)
provides the knowledge required for protecting and restoring sites. While the density of river
monitoring points in Ireland is one of the highest in Europe (M.McGarrigle, EPA, pers com), not all
sites can be monitored. To provide a comprehensive coverage of status across all rivers requires
extrapolating from monitored to unmonitored sites using simple algorithms that relate water body type
to a pressure analysis for diffuse nutrients (Donohue et al., 2006). Status of unmonitored sites is
based on that of the nearest monitored site within a similar water body type. In essence status of
unmonitored sites is based on overall similarity and physical distance. Methods for extrapolation are
still under development, but are likely to be based on amalgamation of unmonitored sites in water
management units (WMUs). For lakes, verification of putative reference sites was done through
palaeolimnology (Leira et al., 2006), but requires much more work to provide a comprehensive
coverage. This will be done through GIS supported modelling (M. McGarrigle, EPA, pers com.).
Extrapolating status from land use can suffer from low predictability, so will require additional testing
of models though monitoring. WFD-compliant classification systems and models which incorporate
multiple, and potentially interacting pressures compounds the problem, and require further
development (Garcia et al., 2006; O’Toole and Irvine, 2006; Donohue et al., 2009; Rask et al., 2011).
Although consistency in the designation and definition of the term ‘reference condition’ is clearly
desirable, variations on the theme globally are widespread (Vendonschot, 2000; Stoddard et al.,
2006; Gibbons et al., 2008): including “totally or nearly totally undisturbed conditions” (European
Commission, 2003a), “best available” (Clarke et al., 2003), “least disturbed’ (Reynoldson et al.,1997,
USEPA 2002a; Davies, 2000; Wigand et al., 2010); “best attainable" (Harrison & Whitfield, 2006);
sites within catchments with low pressures (Lougheed et al., 2007), “relatively undisturbed biological
communities” (Logan & Taffs, 2010); departure from full ecological integrity (Fennessy et al.
2004);and “historical condition” (Young and Sanzone 2002; Nijboer et al., 2004). Setting a particular
date for such impact is, however, ill advised (European Commission, 2003a; Taylor et al., 2006).
Verification of putative reference lakes, using diatom records in sediment cores, showed that 11 out of
35 candidate Irish reference lakes were verified as being in reference condition (Liera et al., 2006). In
the west of Ireland major cycles of impact were related to increased human population in the century
leading up to the Great Irish Famine followed by recovery and more recent decline over the last 40
years associated mainly with increased agricultural intensity of cattle farming (Donohue et al, 2010).
Degradation of lakes in Demark has been associated with the introduction of the plough in the middle
ages (Johansson et al., 2005).
4
Estimating reference state and incorporating uncertainty in both the initial estimate and departure
from it, are poorly understood and controversial (Moss et al., 2003; Howarth, 2006; Taylor et al.,
2006).
Type-specific reference (Baily et al., 1998; 2004) is likely an inherently flawed concept
because it assumes concordance of biotic communities among similarly-typed water bodies, or
influenced by natural changes within a regional reference network (Bates Prins and Smith 2007).
Having comparable, reliable, reference conditions is, however, pivotal in estimating deviation of
ecological conditions from ‘reference’, so as to calculate the EQR, the metric on which ecological
status under the WFD is determined (European Commission, 2003a; Nijboer et al., 2004). Error in
classifying a site can, therefore, arise from a failure to reliably define the expected state (Oberdorff et
al., 2001). Natural variability of biotic communities compound the problem further, as similarity of
community composition, within and across identified reference sites and within water body types, are
nested along spatial (Little, 2008), biogeographical (Moog et al., 2004; Borja et al., 2009a), and
typological gradients (Hering et al., 2010). Water body “types” are identified under the WFD to allow
comparison across water bodies with similar physical and chemical attributes. As water bodies lie
across multidimensional continua, it is clearly an artificial construct that represents a compromise
between the practical need to keep the number of types as low as possible and maintaining the power
to discriminate between natural variability and anthropogenic impacts (Kelly-Quinn et al., 2009).
A further challenge is identifying ecological boundaries between successive status classes (Ellis et al.,
2006). This is accentuated when trying to separate reference from high status sites (Wallin et al.,
2003; Erba et al., 2009). Addressing uncertainty in class classification and in the methods used to
assess ecological status requires considerable further work to obtain sufficient precision using fixed
boundary values (Carstensen, 2007; Carstensen and Henriksen, 2009; Hering et al., 2010), and is the
subject of ongoing work (www.wiser.eu).
The identification of reference sites require validation through monitoring using ecological criteria
(Economou, 2002; Nijboer et al., 2004; Chaves et al., 2006; Erba et al., 2009). Initial views and
validation can be quite different for some sites (Liera et al., 2006). While they are essential for
validation of reference conditions, the proliferation of indices and metrics over the past decade has
further confused rather than simplified the issue, with inconsistencies across regions (Borja et al.,
2009a,b; Noges et al., 2009). In general, the development of classification techniques has also
focused on metrics based on the composition of a variety of biological groups, rather than attributes of
ecological function (see Section 2.2).
The determination of reference state that WFD classification depends is, therefore, inexact, and
subject to ongoing discussion and resolution. It may be that only site-specific reference state
(Carvalho et al., 2009; Jyväsjärvi J et al., 2009.) is a valid concept, so that a site is only compared
with itself over temporal scales, rather than with a network of similar sites over spatial scales. If
monitoring fails to account for temporal or spatial variability (Irvine, 2004), or there are idiosyncratic
response of biological communities in water body types, the very concept of reference condition may
be unworkable (Hartnett et al., 2011). While this view has strong ecological merit, and indeed reverts
to a view of water bodies prior to the WFD (Moss et al., 1994) that, because of site-specific variation
5
in dose-response relationships, each waterbody responds to pressures uniquely (Yarrow and Marin,
2007). Variation in natural background conditions can also mimic anthropogenic disturbance (Fabris
et al., 2009; Hartnett et al., 2011). Site-specific assessment requires long-term monitoring, or robust
site-specific models (Clarke et al., 2003; Pont et al., 2006; Cardoso et al., 2007; Aroviita et al.,
2009a,b; Carvalho et al., 2009). While this may be the logical solution to the uncertainty within types,
it is not feasible for most water bodies. It may also not be WFD-compliant (Hering et al., 2010). A
wider discussion on the problems that the WFD legal standing has for assessing ecology is given in
Howarth et al (2006).
The key point is to recognise, however, is that while the reference state
concept may indeed be flawed and comparisons across sites are either difficult or inherently
approximate, the existence of high status sites prior to major anthropogenic impact is self-evident.
Within the constraints of resources and knowledge it is, therefore, important to identify minimally
impacted sites, and these require particular protection. The ongoing trend of the degradation and loss
of these sites in Irish rivers demonstrated by long-term monitoring (EPA, 2009a), and in lakes
demonstrated by palaeolimnological studies (Leira et al., 2006; Taylor et al., 2006; Hobbs et al., 2005;
Donohue et al., 2010) highlights that need. The EPA have an ongoing programme to assess
catchment activities in catchments with high status waters, or those which were at high status until
recently. This should provide greater knowledge of local and more widespread impacts on these
waters. Generic land use models have limited value for detection of localised impact (Newson, 2010).
The need to stem the degradation of high status sites merits high priority, not least because
preventing, or addressing, small,, impacts is a feasible option, and likely much more cost effective
than large scale restoration to good status for sites at moderate status or worse.
The importance of the decline of high status sites is not confined to a breach of a European Directive,
but is of fundamental significance for maintenance of biodiversity, ecological integrity and as refugia
of species from a widely impacted landscape (Aroviita et al. 2009a,b; Bradley et al., 2003; Hering et
al., 2010).Such refugia are likely crucial for recolonisation of restored sites, as the target for good
status through implementation of the WFD is realised (Meyer et al., 2007; see also Section 2.2).
Habitat variability, or patchiness, at local scales tends to be greatest at low levels of impact. The
same principle applies at regional scales across a range of taxa groups (Donohue et al., 2010). A
network of high status sites provides a mechanism for the preservation of European aquatic
biodiversity, and as a possible buffer to impacts of climate change (Hering, 2010). This is also crucial
to meet European and global ambitions to halt biodiversity decline (Secretariat of the Convention of
Biological Diversity, 2001; EC, 2011).
Globally, aquatic ecosystems are the most impacted habitats
by human activities and continue to decline at an alarming rate (Groombridge et al., 1998; Millenium
Ecosystem Assessment, 2005). More than half of the world’s wetlands and two-thirds of European
wetlands may have been destroyed in the last century (CEC, 1995; Ramsar Convention Bureau,
1996), promoted by policies that encouraged drainage and land reclamation (Pursglove, 1988; Green
et al., 2002). Many others are severely damaged through land-use activities, particularly through
nutrient enrichment (Smith et al., 2006). The importance of wetlands is increasing encapsulated in
the benefits they offer for ecosystem services (Covich et al., 2004; TEEB, 2009). Recent initiatives in
the Netherlands, U.K. and Germany that aim to provide flood mitigation measures, with conservation
6
enhancement are reviewed in Williams et al. (in press). Such programmes restore wetland functions
and have high applicability to a wider landscape approach for the protection of wetlands.
1.2.2. Vulnerability and spatial networks
Vulnerability of wetlands has been long recognised, and wetlands were the first major ecosystem to
be protected by an international treaty, the Convention on Wetlands of International Importance
especially as Waterfowl Habitat, held in Ramsar, Iran, in 1971. The Ramsar convention entered into
force in 1975 (Matthews, 1993; Ramsar Convention Bureau, 1996), and by 22 November 2001, 130
States had become Contracting Parties. As of 10th June 2011, 1933 sites have been declared as
Wetlands of International Importance. While the Ramsar convention does not feature prominently in
discussion of the WFD or Irish habitat protection, its underlying philosophy is the creation of an
interconnected network of aquatic habitats, originally recognised for their relevance for bird
migrations. Spatial pattern and connectivity explains many features of the chemistry and biology of
water bodies. For rivers, longitudinal features are summarised in the River Continuum theory (Vanote
et al., 1980). For standing waters landscape position explains significant aspects of their limnology
(Sorrano et al., 1999; Riera et a., 2000; Kernan et al. 2009), with high relevance for lake typology.
Connectivity of small bodies of waters has been shown to be important for regional biodiversity (Biggs
et al., 2005; Jeffries, 2005) and community structure of invertebrates in Irish lakes has been shown
repeatedly to be spatially nested (Little, 2008; White and Irvine, 2003; Donohue and Irvine,
unpublished data). The rationale for a connected network of high quality sites (Amezaga 2000;
Amezaga et al., 2002) applies equally to the WFD (Hering et al., 2010), fits well with wider
considerations of extensification of land use to support conservation objectives (Lütz and Bastian,
2002; Berger et al., 2006; Von Haaren and Reich, 2006), and should be incorporated into river basin
management plans (Kettunen et al., 2007). In Ireland and elsewhere many high quality sites are
found in the headwaters of rivers, containing ecological communities of fundamental biodiversity
importance. The same principle applies to small standing waters. Their protection is well justified,
involving a wider consideration of ecological quality outside of the typological and geographic
constraints of the WFD (Meyer et al., 2007; Kelly-Quinn et al., 2009). Small water bodies can be
important for regional biodiversity (Bradley et al., 2003;De Meester et al., 2005;Oertli et al., 2005;
Sondergaard et al., 2005), and provide refugia from which re-colonisation of larger water bodies can
occur following restoration (Hering et al., 2010).
A recent ruling by the Court of Justice of the
European Communities (ECJ) in relation to Ireland’s failure to implement Environmental Impact
Assessment effectively (see below, Section 3.7) highlights the importance of small wetlands. In 2007,
high quality ponds were added to the list of UK Biodiversity Action Plan Priority Habitats.
Facilitating a wider countryside approach to the protection, or restoration, of a network of high status
sites is the The European Landscape Convention (see www.coe.int ). This was ratified by Ireland in
March 2002, and requires an integrated approach to landscape planning and management across all
areas of government policy formulation and implementation. In common with WFD implementation,
the Landscape Convention requires a process of public participation and awareness, training and
7
education; the improvement of damaged landscapes; and the integration of landscape in all relevant
policies. There is certainly scope for a greater integration of landscape protection (Heritage Counci,
2006; Fáilte Ireland, 2007) and WFD objectives.
This also links into the possibilities to integrate protection of high status sites with Rural Development
funds, as discussed in Section 4.5. The revised Planning and Development (Amendment) Act, 2010
incorporates a definition of landscape in accordance with the European Landscape Convention. The
Heritage Council (2010) provides examples of how the European Landscape Convention can be
moulded to suit the requirements of individual member states. Examples of particular interest for
Ireland may be the approach of the French Regional Parks ( http://eng.parcsnationaux.fr; Guihéneuf,
2009), The Catalan Landscape Observatory (www.catpaisatge.net/eng/index.php) and the Canadian
Heritage Rivers System (www.chrs.ca).
The French Regional Parks provided a pioneering model for a voluntary charter which has been
proposed by the Heritage Council to protect landscapes in the Burren region in Co. Clare. The
French Charter, which provides for the establishment of the Natural Regional Parks (Parc Naturel
Regional; PNR), was published in 1967 in response to insufficient environmental protection and
shortfalls in existing landscape legislation. Local councils in the regions and counties collaborated
with central government to develop small agricultural areas for biodiversity conservation. A PNR must
adhere to principles promoting sustainable development. PNRs aim to make agricultural and forestry
practices compatible with the conservation of natural environments by establishing agri-environmental
contracts. This approach has led to the establishment of the 46 Regional National Parks.
Similar
provisions are available to Irish Local Authorities under Section 204 of the Planning and Development
Act 2000 to establish Landscape Conservation Areas but has been under-utilised.
The first
Landscape Conservation Area may be established in the Tara/ Skyrne Valley in Co. Meath
(http://www.heritagecouncil.ie/landscape/news/view-article/article/).
The Catalan Landscape Observatory aims to increase the knowledge of landscapes among Catalan
society and to support implementation of the European Landscape Convention by facilitating
communication among the Catalan Government, local authorities, universities, professional groups
and Catalan society in general. The Landscape Observatory is organised as a consortium and is
included in the Act for the protection, management and planning of the landscape in Catalonia
(Resolution PTO/3386/2004).
This Act defines ‘Landscape Catalogues’ as “documents of a
descriptive and prospective nature which identify the types of landscapes in Catalonia, their values
and state of preservation, and propose the quality objectives to be met”. The Landscape Catalogues
integrate landscape into planning, promoting awareness on the diversity and value of landscapes.
The Canadian Heritage Rivers System is a non-statutory model that works through regional
cooperation to support community involvement in the protection of rivers, designated on the basis of
their importance for local heritage and recreation. While largely set up to provide a cooperative
framework to support protection of large rivers, and involvement of indigenous communities, the
8
principles could be adapted for Irish landscapes. Natural values are an essential component of the
Canadian system.
The European Landscape Convention places a large focus on how the public perceives and
evaluates landscapes. This is likely a key factor in securing stakeholder participation in protecting
high status water bodies (see Section 5). The designation of ‘high status’ waters according to purely
scientific criteria under the WFD may not inspire a cultural affinity or understanding towards the
protection of these waters. A broader approach to the protection of whole landscapes may, however,
invoke a high degree of community support. Certainly this was born out by public interest in a openconference on the protection of the Irish Western lakes (Huxley and Irvine, 2008). A proposal to
introduce a Landscape Ireland Act is intended to introduce new participative approaches for
communities for the management of landscapes, (Heritage Council, 2010).
This could facilitate
protection of entire landscapes, including wetlands and small water bodies which are not currently
classified under the WFD. Following a similar approach to that of the Countryside Agency and
Scottish National Heritage (www.naturalengland.org.uk/Images/lcatopicpaper3_tcm6-8173.pdf), a
Landscape Characterisation Assessment has been developed by the U.K.
Department of
Environment, Food and Rural Affairs (DEFRA) for some river corridors, and for the 22
Environmentally Sensitive Areas (ESAs) focussing on environmental impact by agriculture.
The distribution of high status sites in Ireland is spatially patchy and concentrated in the west of the
country (Figure 1). This reflects intensity of land use. The EU Habitats Directive designates sites on
the basis of being representative of habitats in the Member States. The identification of high status
sites under the WFD is, in contrast, a reflection of the current status quo. The WFD date for setting
status was 2009, which implies that degradation of these sites prior to this date is acceptable.
However, for national preservation of aquatic biodiversity, there would be consideration of restoration
back to high status sites, in order to effect a national and interconnected network of type-specific high
status sites, akin to the philosophy of the Habitats Directive. This provides a fundamental challenge
for landscape management and interaction with other policies, especially those relating to agriculture
and rural development (See Section 3.3 and Section 4.5). There is only moderate, and chance,
overlap between high status sites and candidate Special Areas of Conservation (cSACs). The
protection of high status water bodies in Ireland can be viewed legitimately as both the preservation of
Irish heritage and an international responsibility. National economic interests can frustrate meeting
such obligations. Objectives that reconcile these with the global biodiversity crisis (Abell, 2002)
requires a new paradigm of thinking.
9
Figure 1. High status surface water bodies nationally as identified in the RBMPs 2009-2015
10
Across Europe, a network of ‘high status sites’ as key areas that protect aquatic biodiversity is
advocated by Hering et al (2010). In France, about 400 sites characterized by a low level of human
pressure and good biological quality provide a permanent reference monitoring network. The
European Environment Information and Observation Networking improving Europe's environment
(EIONET; www.eionet.europa.eu), a network of the European Environment Agency could provide
this, possibly linked to
Long-Term Ecosystem Research sites (LTER: http://www.lter-europe.net)
(Hering et al., 2010).
Key points: High status water bodies:
High status water bodies are especially vulnerable to low levels of anthropogenic impact, and long
term monitoring of rivers by the Irish EPA shows a continued decline of these sites.
Palaeolimnological data shows a variable timing in the on-set of impact. High status is difficult to
distinguish from reference state, and should be considered synonymous with it for management. A
regional connected network of high status sites merits consideration, to include strategies for
restoration of sites that could be, but are currently less than, high status. Defining high status is a
challenge across the EU, but irrespective of difficulties in identifying type-specific reference
conditions, there is a need to instigate policies to protect and enhance the network. Aquatic sites are,
globally, the most vulnerable and impacted habitat from anthropogenic pressure.
Ireland, as a
signatory of the Ramsar Convention, and through the Millennium Ecosystem Goals, has international
obligations to protect its aquatic sites, of which the high status sites represent the best quality. This
obligation, fundamental to the implementation of the WFD, is further supported through other
legislation such as provided by Environmental Impact Assessment.
Section 2. Legislation and Policies for the Protection of
High Status Sites.
2.1 Introduction
The WFD has inter alia a legal requirement that by 2015 surface waters in the EU Member States
achieve at least good ecological status, and that deterioration from one status class to another is
prevented. Deterioration of ecological status within a status class is not a legal requirment per se, and
difficult to verify, but a clear and obvious requirement for environmental protection of the high status
sites is minimising potential impacts. Similarly, while the very essence of the WFD is based on the
acheivement of at least good status for all water bodies by 2015, there is no legal provision for any
restoration to high status unless covered by legislation under the Habitats Directive (92/43/EEC) in
achieving favourable conservation status (Withrington, 2005) and where this, by default, achieves
11
high status under the WFD (see below, Section 3.2). Under the WFD, the mechanism to effect
environmental objectives are the Programmes of Measures (POMs), outlined in Article 11 of the
Directive. In order to develop the POMs, it is necessary to identify likely pressures and impacts on
waterbodies. This was done through the Article 5 Characterisation report (Government of Ireland,
2005), drawing on the methods produced at European level by the IMPRESS working group under
the Common Implementation Strategy (European Commission 2003b). The initial pressure and
impact assessment in Ireland was focussed on the risk of water bodies failing to achieve good status.
Consideration was not given to risk of failing to meet high status, although all high status sites were
deemed to be at risk of failing to meet the environmental objective (pers. Comm, M. McGarrigle,
EPA). Many high status sites are subject to localised small scale, but extensive, pressues, such as
local pollution and drainage. These are not necessarily documented in the Significant Water
Management Issues (SWMI) reports done for each River Basin District as part of the drafting of the
RBD management reports (wfdireland.ie). Detailed assessment of potential impact at the site level
for many high status water bodies is, therefore, limited but essential.
In the Republic of Ireland, good status refers to the achievement of at least mandatory standards
prescribed in national legislation tranposing 11 key Directives (listed in Annex VI part A of the WFD)
relevant to water protection i.e. The Bathing Water Directive (2006/7/EC);The Birds Directive
(79/409/EEC); The Habitats Directive (92/43/EEC); The Drinking Water Directive (98/83/EC); The
Major Accidents (Seveso) Directive (96/82/EC); The Environmental Impact Assessment Directive
(85/337/EEC) as amended by Directive 2003/35/EC; The Sewage Sludge Directive (86/278/EEC);The
Urban Waste-water Treatment Directive (91/271/EEC); The Plant Protection Products Directive (EC
No 1107/2009); The Nitrates Directive (91/676/EEC); and the Integrated Pollution Prevention Control
Directive (2008/7/EC).
The assumption is that compliance with the 11 Directives listed in Annex VI part A of the WFD
provide the minimum measures required to meet the environmental objectives for good status. In
previous Government discussions of WFD implementation, these were referred to as Basic Measures
of Article 11 of the WFD. Supplementary Measures were considered those that were needed in
addition to the basic measures to meet the environmental objectives of the Directive, and referred to
in Annex VI, part B of the WFD. The distinction between Basic and Supplementary Measures seems
to have disappeared recenty from the official WFD language in the Republic of Ireland, and is not
found in the River Basin Management Plans (RBMPs). However, the option to use additional
measures to protect aquatic systemsover and above those provided by the transposed WFD
legislation and the 11 associated Directivesremains an option. This is of crucial importance for the
protection of high status. It is also important that a) there is no fundamental conflict between the
objectives of those Directives listed in Annex VI part A, and any subsequent amendments, and those
of the WFD and b) that sufficient attention is paid to water quality objectives and, more so, cross
compliance with ecological standards. The key issues for this review are, therefore, to what extent
specific WFD and other legislation provides protection of sites that have been identified as high
status, and is cross compliant with the WFD. This Section reviews these questions at three scales:
12
1. WFD–specific legislation in Ireland and if this is fit-for-purpose for protection of high
status sites;
2. protection afforded to high status sites through the Directives listed in Annex VI part A of
the WFD, some of which have been amended since the publication of the Directive; and
3. potential impacts on high status sites arising from other policies and practices.
2.2. The likely effectiveness of transposed WFD legislation for protection of High Status
Sites.
The WFD was transposed initially into Irish legislation by the European Communities (Water Policy)
Regulations (S.I. No. 722 of 2003). These dealt primarily with administrative arrangement of the WFD
and the formation and operation of the River Basin Districts (RBDs). These Regulations are
summarised in Irvine and O’Brien (2009) in so far as they relate to the working of the River Basin
Advisory Councils and procedures and experience of stakeholder involvement. It is now generally
accepted that the administrative arrangements for the WFD in Ireland have not been sufficiently
effective for water governance in Ireland which, in common with environmental protection policies in
general, are fragmented and in need of a major overall (DECLG, 2011). In Schedule 1 of The Surface
Water Regulations. S.I. No 272 of 2009 (see bleow), 23 relevant public bodies are listed.
That
fragmentation, the working of the River Basin Advisory Councils, and the redirection of focus from
specialised River Basin Managment project teams to the local Authorities, which may lack resources
and expertise in the area, provide underlying difficulties for the implementation of the WFD. The
protection of high status sites is one important facet of this. Furthermore, the Advisory Councils,
disssolved for local elections in 2009, have not been reconstituted, in contravention of S.I. No 413 of
2005 amended Article 16 of S.I. No. 722 of 2003 .
The European Communities Environmental Objectives (Surface waters) Regulations S.I. No 272 of
2009, transposes the requirements of Articles 6 and 9 of the WFD into Irish legislation, to provide
measures for the protection of surface waters whose status is deemed to be high or good (or good
ecological potential for Heavily Modified or Artificial waters), and for restoration of waters deemed to
be at less than good status. The regulations allow (under Article 43) to redress upward trends of
pollution, including within-status trends, that would likely result in deterioration in status over time.
This is effected through advice to relevant public authorities. For protected areas, the Regulations
allow for the designation of less than good status where compliance with other European legislation
has not been met (Article 49). This lies at the heart of difficulties in harmonising the WFD with the
Habitas Directive, the protection of high status waters, and designation of favourable conservation
status of cSACs.
With the exception of the European Communities Environmental Objectives
(Freshwater Pearl Mussel) Regulations 2009 (S.I. 296 of 2009) there are no POMs to address specific
water managment problems in areas protected for nature conservation. This is discussed further in
Section 3.2 below. There is also no mention in the European Communities Good Agricultural Practice
13
for Protection of Waters S.I. 610 of 2010 of strategies to protect sites designated under the Habitats
Directive.
The Surface Water Regulations S.I. No 272 of 2009 provide, under Schedule 2, a series of measures
that implement Community legislation for the protection of waters. The high rate of decline of high
status sites demonstrates the failure of exisiting polices to be effective. The Article 5 Pressures and
Impacts analysis as detailed in Government of Ireland (2005), identified in the SWMI reports and
summarised in RBMPs (2010) have identified the key pressureson aquatic habitats to be: 1. diffuse
pollution sources particularly from land use; and 2. morphological alterations particularly associated
with rivers, impoundments of lakes, channel drainage, and activities associated with ports in
transitional and coastal waters.
The prevalance of diffuse pollution is ubiquitous across the country, affecting both
surface and
groundwaters. Waste Water Treatment Plants (WWTPs), septic tanks (unsewered on-site waste
water treatment systems (OSWWTS) and priority substances have also been identified as important
and widespread pressures on water bodies. Hydromorphological alterations, such as river
channelization and water abstraction are identified as the second most prevalent risk of water bodies
failing to meet their environmental objectives (Government of Ireland, 2005), but impacts of this are
unquantified. Widespread land drainage provides a major pressure in accelerating movement of water
and nutrient transport from land to waterbodies. Sediment loss arising from a variety of urban and
rural activities also impact water bodies, but an overall view of the extent and importance of this is
very scant, although clearly has been, and continues to be, of potential major importance in peatlands
degraded by overgrazing (Bleasdale, 1998), and in forestry following tree harvesting (see Section
4.6).
For the monitoring of surface waters the WFD requires that techniques are developed that enable the
estimation of an EQR. For many of the biotic elements listed in Annex V of the WFD, these
techniques are still under development across Europe (Hering et al., 2010). Schedule 5 of S.I. No
272 of 2009 provides the details for calculating EQRs for the high-good and good-moderate boundary
for those elements where national agreement has been reached on the appropriateness and reliability
of such techniques, albeit that these may be subject to future refinements, including the estimation of
confidence limits around status class boundaries. The technical details, or algorithms, of how EQRs
are calculated for each element are not included in the Schedule, but are contained in Standard
Operating Procedures used by the agencies charged with these estimates.
Techniques for
developing EQRs for different elements across water bodies can differ, but the assumption is that
overall classification is robust. Consistency across water body classification, especially where water
bodies are physically connected, is important for the development of strategies for integrated
monitoring and protection.
The relevance of an integrated assessment process can be illustrated by looking at the supporting
nutrient chemistry for rivers and comparing that with a general view of nutrient state in lakes. Under
Schedule 5 of S.I. No 272 of 2009, high status of rivers is < 0.025 mg l -1 of unfiltered molybdate
14
reactive phosphorus (MRP) as a median value or < 0.045 mg l-1 as a 95 percentile. These values are
equivalent to 25 and 45 µg l-1 of MRP, respectively. Phosphorus standards for lakes are not yet
established, but it is possible to make some estimates of how this equates to lake nutrient state,
based on a long traditional of limnology and reference to the OECD (1982) classification scheme for
lakes, the basics which were used by the EPA, and its forerunners, for many years.
Donohue et al.
(2006), in work done by the EPA, considered that “Reference or natural concentrations for soluble
reactive phosphorus in rivers are typically in the range 0-10 μg l-1”. Moorkens (2006) used a median
value of 5 μg l-1MRP, as background concentrations, based on an analysis of records over time for
freshwater pearl mussel populations with recruiting juveniles. Gibson
et al. (1995) reported that
background concentrations of phosphorus relating to rainfall and, therefore, minimal human impact in
the catchment, for upland rivers were in the order of 15 μg l-1 . This is also the target concentration
for the upper catchment lakes that supply source water for New York City (Dell et al., 2009, and see
Section 6.4). Using information from diatoms in sediment Foy et al (2003) estimated a reference TP in
Lough Neagh of 19 μg l-1. Taylor et al. (2006) estimated lowest values of diatom-inferred TP from
sediment cores in a number of Irish lakes to be < 20 μg l -1, with many between 15-20 μg l-1. Anderson
(1997) estimated diatom-inferred TP concentration in six small rural lakes in Counties Down, Armagh
and Tyrone to range from 6-58 μg l-1, all from before 1900. The interpretation of diatom-inferred (i.e.
modelled) TP needs to be treated with some caution (Rippey et al., 1997), and in Ireland significant
impact in many areas is likely to have occurred prior to 1850 owing to high human
population
densities and tillage (Donohue et al., 2010). All of the 11 of the 35 candidate lakes that Leira et al.
(2006) confirmed to be in reference state had mean TP concentrations (EPA data) < 20 μg l-1.
Using
a modern network of putative Northern Irish reference lakes, Rippey et al. (unpublished data,
University Ulster) estimated mean TP to range from 5-25 μg l-1.
Background concentrations of phosphorus in calcareous lakes and turloughs, with their capacity for
attenuation and sedimentation of phosphorus (Otsuki and Wetzel 1972, Søndergaard et al., 2003),
are likely <10 µg TP l-1 (Hobbs et al., 2005; Donohue et al., 2010; Peirera, 2010). Kilroy (2001),
using national data from 1995 to 1997, found a median value of 17 μg P l-1of unfiltered molybdate
reactive P in groundwaters, but that one quarter of the data were higher than 30 μg MRP l − 1, which
under the previous EPA lakes classification system (based on OECD, 1982) would be considered to
be eutrophic if found in surface waters as TP.
To estimate the potential effect of nutrient standards for rivers on lakes, data was complied from Irvine
et al. (2001) that estimated residence time for 28 lakes of varying depth and surface area. Residence
time was estimated by division of annual hydrolic load, estimated from weather data, divided by lake
volume. Based on the mean input P concentrations from the river standrads, annual mean lake P
concentration in each of the 28 lakes was modelled from their residence time (τw):
PLake = 1.55 [Pinput/1+√τw]0.82
from the equation originally formulated by OECD (1982), and used widely in lake nutrient assessment.
(Note a modification of the OECD equation estimated by Foy (1982) for Irish lakes, and used by Irvine
15
et al., 2000, provides slightly higher estimates of in-lake TP). The model was run using both the highgood (median of 25 µg l-1 of MRP) and good-moderate boundary (35 µg l-1 of MRP) used in
Schedule 5 of S.I. No 272 of 2009. An assumption was made that mean concentration of MRP as
estimated by the EPA standard practice of using unfiltered water samples equated to lake total
phosphorus (TP). This is likely to be a conservative conversion, as the EPA analytical techniques for
MRP will underestimate TP (Irvine et al., 2002). Fig. 2. shows the outcome of lake phosphorus of this
simple model for inputs from rivers for both the high-good and good-moderate thresholds. For all
lakes except those with long residence times, approaching 4 years, mean modelled TP for inputs
from high status rivers exceeds the OECD (1982) thresholds for oligotropic waters.
Higher
concentrations are, naturally, estimated for the good status rivers, with the modelled lake
concentrations appoaching those considered by OECD (1982) to be eutropic for lakes with low
residence times. This suggests a miss-match between the current river nutrient standards and the
protection of high status lakes, suggesting that river concentratuions at the respective upper
boundaries of high and good status are commensuate with a degradation of lake water quality as
estimated by OECD (1982).
30.0
Eutrophic
TP in lake
25.0
20.0
15.0
Mesotophi
10.0
c
5.0
Oligotrophic
0.0
0
2
4
6
8
Tw (years)
Fig 2. Modelled mean concentrations of TP (µg l -1) in lakes (based on average concentration of phosphorus in
inflowing water to a range of lakes of different residence times, for concentrations of 25 µg l -1 (diamonds, lower)
and 35 µg l-1 of MRP (squares, upper). Trophic classification boundaries for mean annual lake concentration of
TP are given.
It is worth emphasising that this simple modelling exercise is done to illustrate that the process used
for establishing the concentrations of phosphorus in rivers appears not to have considered the impact
this may have on downstream lakes. However, this likelihood is also accepted by the EPA
(M.McGarrigle, EPA, pers com), such that lakes receiving water from rivers may need stricter
standards than are in the current Regulations S.I.272 of 2009.
Laoadings from generally low
impacted catchments, such as to the west of Lough Mask in Co.Mayo
can have P loads
approximating 0.1 kg ha-1 year-1 (Donnelly, 2001). There is, therefore, a strong need to reassess the
rationale for P mangement entering high quality lakes. A similar mismatch may apply to nutrient loads
16
to transitional waters. The model illustrated in Fig 2 is also based on average concentrations and a
number of assumptions in converting the estimates of MRP in rivers to TP in lakes, averaged over a
year. It has been well established that average point sampling in rivers, as done by the EPA, is not
an effective method to estimate total diffuse loads to a river because high concentrations in rivers are
strongly skewed with high rainfall events (Lennox et al., 1997; Morgan et al., 2000). Diffuse loading,
arising from agriculture is the major input to Irish water bodies (Foy et al. 1995; Allott et al., 1998;
Lennox et al., 1998; Jennings et al., 2002; Bartley and Johnston, 2005), with an association between
increased pasture in a catchmet and declining ecological status, using WFD criteria (McGarrigle,
2009). There is, therefore, a strong need to move from monitoring a limited number of individual
samples for nutrients over an annual cycle in rivers to measurements at high temporal resolution, or
use of validated models, to enable estimates of phosphorus loads. This is used in many U.S. nutrient
management programmes, adopting protocols for estimating Total Maximum Daily Loads (TMDLs) as
required under Section 303d of the U.S. Clean Water Act 1972 (USEPA 2008). TMDLs are required
for all waters that do not meet water quality standards, and as a management tool to maintain quality
standards. The principle of TDMLs is that they represent an assimilative capacity (Havens and
Schelske, 2001) of the receiving waters with respect to the specific pollutants and which is compatible
with designated use for e.g drinking water, fishing and recreation. TMDLs are site specific and must
include the total of all point and diffuse loads, incorporate a margin of error, and account for seasonal
and spatial variability of load and impact. A TMDL implementation plan is analogous to the WFD
Programmes of Measures, and frequently uses modelling to determine effectiveness of control
measures (Ambose et al., 1996; Irvine et al., 2005). Articles 7 & 9 of SI No 272 of 2009 require the
review of discharge licenses to support the environmental objectives.
A TDLM approach would be
useful for determining chemical standards for high status (and other) sites in Ireland, but requires
suffcient data, or robust models, on point and diffuse loads.
Under the WFD, many small water bodies (lakes <1 ha and first-order streams) do not require
assessment. First and second order streams are located in the headwaters of catchments and
comprise the majority (up to 70%) of the Irish national river network. They are particularly vulnerable
to localised impacts and many are important for salmonid spawning. The small Stream Risk Score
(SSRS), developed by the Irish EPA can support the POMs for such sites, by providing high spatial
resolution data (Ní Chatháin, 2006).
Owing to current lack of method development, S.I. No 272 of 2009 does not include all the elements
listed in Annex V of the WFD. These will be included pending developmemt of techinques.
The
development of ecological assessement has, across Europe, focused on what might be termed the
structure of communities (what is present and in how much abundance), rather than any attributes of
function (such as measures of production and communities interactions). For high status sites the
principles of nutrient parsimony, naturalness, food web structure, hydrological connectivity and
size,outlined by Moss (2008) should apply. Intact ecosystems are typified by nutrient retention and
lack of measurable freely available nutrients (Likens et al., 1971; Raven et al., 2005). Nutrients in
ecological assessment of the WFD are a supporting element, with the main emphasis on biotic
17
elements. The current protocols developed for the ecological assessment is not, therefore, the same
as a focus on ecology, which is inherently about the interactions among the biotic and abiotic
components. Understanding the ecology within high status sites would appear to be a self-evident
requirement in order to identify impact of, and mitigation from, pressures. Additionally, for lakes, and
possibly transitional waters and larger rivers, the assessment of zooplankton, and their ecological
role, is an obvious omission (Caroni and Irvine, 2010; Jeppesen et al., 2011). The philosophy within
the WFD is that a number of elements are used to assess ecological status has merit, but assumes
equal sensitivity across all elements to defined pressures. It also requires that status is determined
on the basis of the most impacted, the so-called “one-out all-out“ principle. The probablity of
misclassification increases with increasing number of variables or parameters
included in an
assessment. The difficulties inherent in the “one-out, all out“ principle were well demonstrated in an
extensive pan-European project ECOFRAME (Moss et al., 2003), and its consequences for
classification and making type 1 (detecting an impact when there is none) and type 2 (failing to detect
an impact) errors are discussed further in Paavola et al. (2003) and Hering et al (2010).
The probability of making a type 1 or 2 error, identified with uncertainty bands around class
boundaries (Ellis 2006) is not addressed in S.I. No 272 of 2009, or indeed in general across Europe
(Hering et al., 2010). Estimating uncertainty in classification remains a major challenge for
implementation of the WFD (Sigel et al., 2010) and environmental policy in general (Handmer et al.,
2001; Harremoes, 2003; Morgan and Henrion, 1990; Pahl-Wostl, 2002, 2007; Newson, 2010). This
relates not only to statistical probabilities, but also to stakeholder perceptions and priorities (See
Section 5). For example, reviews of reliability of RIVPACS (Walley and Fontama, 1998ab, 2000), the
main tool for UK river assessment has been critical of its approach and high degree of
misclassifications (see also summary in Irvine et al., 2005). The developmentof RIVPACS took over
20 years and cost several million pounds sterling. It is, therefore, not surprising that the classification
tools across other elements are still developing, and it is important to acknowledge and accept current
uncertainty in water body classifications. It might also be worth considering the view of Opdam et al.
(2010) that “The claim that certainty has to be provided by science is unrealistic, because policy
causes a good deal of uncertainty affecting how science can operate“. These and other complexities
identified during the first stage of WFD implementation are unlikely to have been envisaged by the
authors of the Directive (Hering, 2010).
The lack of protection of high status sites in Ireland is well illustrated in Kavanagh and Bree’s (2009)
summary of measures of protection. In Table 4 of that article there is list of Supplementary Measures,
summarised from Mayers (2008). In a list of six Supplementary Measures there is not one that, under
current policies, is able to be backed with tangible management. The list comprises what might be
termed policies rather than measures.
Key points: WFD Regulations:
The WFD and its transposition into Irish law have revealed a number of difficulties with the process of
classification, for which further consideration and devlopement is required. The consistency among
18
biotic elements and inherent uncertainty surrounding classification boundaries present significant
challenges for robust classification. A more sophisticated approach to understanding the ecology,
rather than the construct of ecological classification, is needed to support protection and management
of the highest quality water bodies in Ireland. The environmental objectives to generally achieve good
status, and the no-deterioration requirement for high status water bodies,favours the current status
quo for high status sites, rather than a programme of protection that includes targeted management of
high status water bodies. The Surface Water Regulations S.I. No
272 of 2009 outline that no
degradation from one status class to another is allowed, but there is
Programems of Measures
to
little in way of specific
support the protection of high status waterbodies.
Site-specific
management and techniques such as those contained in the Small Stream Risk Score (SSRS)
assessment provide a possible framework to at least identify localised pressures that may impact high
status waters.
This will require a greater emphasis on specific, and local, rather than generic
measures for management. This could include adopting assessment similar to the USEPA TMDLs for
individual sites. Small standing waters and headwater streams are important ecological resources and
should be included in policies designed for high status waters.
Section 3. The Effectiveness of the Associated Directives
listed in Annex VI for the Protection of High Status Sites.
3.1 Introduction
The Directives listed in Annex VI Part A of the WFD are to provide a series of Programmes of
Measures (POMs) to support the objectives of the WFD. This Section reviews their effectiveness in
protecting high status water bodies. For the purpose of this review, the Birds Directive (79/409/EEC)
and Habitats Directive (92/43/EEC) are discussed together, as they are both linked with the Natura
2000 network under an overall theme of nature conservation.
3.2 The Birds Directive (79/409/EEC) and Habitats Directive (92/43/EEC)
The Birds Directive (79/409/EEC) provides a framework for the conservation and management of wild
birds, requiring the designation of Special Protection Areas (SPAs).
Directive (92/43/EEC) on
Conservation of Natural Habitats and of Wild Fauna and Flora, requires Member States to designate
sites identified as Special Areas of Conservation (SACs) which, together with the SPAs, form the
Natura 2000 network of protected sites. Because the SAC list submitted by the Republic of Ireland
has not yet been formally approved by the E.C., these are officially still candidate SACs (cSACs).
The link that the Birds and Habitats
Directive have with the WFD has presented considerable
discussion, including whether or not they are mutually supportive (Hatton-Ellis, 2008; Cowx et al.,
2009; Irvine 2009), or if the WFD can provide significant benefits for water dependent protected
19
areas. The EC recently prepared a document on the links between the two Directives, and “frequently
asked questions” (EC, 2010). Overall the document reiterates various statements from the Directives,
but useful guidance is somewhat lacking. This likely reflects the post-priori, rather than planned a
priori, linking of the Directives. The document did clarify that favourable conservation status under the
Habitats Directive is likely to be the more stringent than good status under the WFD. This, however,
avoids the issue of the relationship between high status and favourable conservation status. Under
S.I. 272 of 2009 if a protected area does not attain favourable conservation status, because of a
failure to meet water quality or hydromorphological standards then it cannot be classified as good
status under the WFD. This has merit in that it ensures that restoration of those sites are required for
WFD compliance, although it is not clear if, in such an eventuality, a site would then be classed as
high status, or if de facto all sites at favourable conservation status are synonymous with high status.
There is also a risk, however, that that downgrading cSAC sites to moderate status means that
chemical targets for restoration or of management (such as licensing) may not be sufficiently
stringent.
Legal advice sought by the U.K. conservation agencies highlights one other important
aspect of this debate (Withrington, 2005). Article 4.1(c) of the WFD requires POMs to “achieve
compliance with any standards or objectives” in Community legislation under which the protected area
was designated. This means that POMs need to be applied under the WFD in order to meet the
objectives of Natura 2000 sites, even if those sites are not identified as water bodies under the WFD.
Withrington (2005) provides the examples of restoration of river and coastal SACs as valuable
consequences of this decision.
For protection of high status water bodies in the Republic of Ireland, two key messages arise from this
discussion. The first is that the respective agencies need to coordinate their ideas and plans to
maximise the synergies between the Directives and minimise the conflicts. This coordination needs to
extend to sharing of data and use of compatible, and readily available, IT systems, including GIS. This
reinforces the recent recommendations arising from a review of the EPA (DECLG, 2011). The second
is that the use of Supplementary Measures, such as those indicated in Annex VI, part B of the WFD
needs to be considered for WFD compliance. In this regard the Irish Government may want to
rehabilitate the terms Basic and Supplementary Measures to the nation’s dictionary of WFD
vocabulary.
The WFD incorporates obligations for nature conservation explicitly in Article 8, including the need for
monitoring and management over and above basic requirements. Of the 689 high status river water
bodies identified in the RBMPs (2010), 31% are within cSACs. High status sites provide a
comprehensive documentary of the best quality sites across Europe. The mechanism by which these
sites were identified, means that there was no defined strategic link between high status sites and
protected nature conservation areas under the Habitats and Birds Directive.
In Ireland, the implementation of the Habitats Directive has been far short of providing the required
protection for lakes and rivers, with a number of difficulties that arise for a variety of reasons (Irvine et
al 2007). Under Article 11 of the WFD, the Birds and Habitats Directive are two key measures for
WFD compliance within their relevant remits. Curtis et al. (2009) provide an overview of the
20
requirements of species and habitats that are covered by the Register of Protected areas (Curtis et
al., 2006) required by Article 6 of the WFD, and the gaps in information required for their protection.
They found, with a few exceptions, a general lack of data, or data at inappropriate scales. They
suggest an “urgent necessity” for detailed survey of species and habitats distributions and
requirements. While management requirements for protected species and habitats have been
compiled (Curtis et al., 2009), there are few detailed action plans that are adequately backed-up by
legislation or effective policies.
Freshwater
Pearl
Mussel
The notable exception to this are the measures to protect the
(Margaritifera
margaritifera)
under
the
European
Communities
Environmental Objectives (Freshwater Pearl Mussel) Regulations 2009 (S.I. 296 of 2009).
S.I. 296 of 2009 set environmental quality objectives (EQO) for Freshwater Pearl Mussel habitat, and
requires the production of sub-basin plans with measures to a) set specific objectives and targets; b)
investigate sources of pressures leading to the unfavourable conservation status ofthe pearl
mussel;c) establish a programme, including a timeframe, for the reduction of pressures giving rise to
unfavourable conservation status; and d) provide for monitoring to evaluate the effectiveness of
measures and progress made towards restoring favourable conservation status. Of the 93 known
populations of pearl mussels in the Republic of Ireland, some of which include two or more rivers in
close enough proximity to make them one single population, 27 populations have been designated
within 19 SACs designated for Margaritifera margaritifera. While references conditions have not been
set for the pearl mussel (Curtis et al. 2009), these are likely to be more stringent than agreed for high
status rivers under S.I. No 272 of 2009. This then raises questions about the classification of high
status rivers in general, suggesting that for rivers for which pearl mussels are currently or previously
found, supporting chemical and habitat (i.e siltation) conditions should be condusive to the survival
and reproduction of pearl musssels. It, furthermore, suggests that any pearl mussel river outside the
SAC network cannot be designated as high status if it does not have viable popluations of the
mollusc.
The pearl mussel is either doomed to extinction in Ireland, or severely contrained in its national
distribution as only one population is viable, and hence considered to be at favourable conservation
status and, therefore, at high status (as required by S.I. No 296 of 2009, Schedule 4). For the pearl
mussel it is probably a case of too little too late (Moorkens, 2010), but even if these measures do not
safeguard the pearl mussel they will provide a level of protection greater than that currenty provided
for by the Surface Water Regulations S.I. No 272 of 2009 in waters in which the mussel occurs. The
pearl mussel plans also provides a precedent for a process that recognises that some sites (including
many cSACs at less than favourable conservation status) require additional measures, as envisaged
in Article 11 and Annex VI part B of the WFD. The need for such additional (or Supplementary)
measures apply to a number of other species and habitats that require particular measures that are
currenly not in place. A prime example is Atlantic salmon (Salmo salar), which requires river water
with summer Q values of 5–4, and typified by low nutrient concentrations (Curtis et al. 2009).
The
EPA report on water quality for the period 2007-2009 (EPA, 2010) showed that only 10 of 132 sites
21
surveyed for fish in 2008-2009 were at high status for fish biology, and about 50% were at good
status (EPA, 2010).
For other species that would be indicators of high status sites, but which are not listed in Annex 11 of
the Habitats Directive, there is no statutory protection or species action plan. A good example is the
Arctic Char (Salvelinus alpinus) athough they are often associated with other species or habitats that
are protected (Igoe and Hamar, 2004).
It is the view of the National Park and Wildlife Service
(NPWS) that such association “should provide some protection“ (C. O’Keeffe, NPWS, pers. comm.,
quoted in Igoe and Hammar, 2004). In contrast, in Northern Ireland there is an action plan for the
species (Environment and Heritage Service, 2007), indicating a greater acceptance of the
conservation needs of important species that may not be covered by the Habitats Directive.
The WFD obligations for nature conservation in Article 8 requires targeted monitoring that supports
management of protected areas, and which supplements basic monitoring and management. Under
the WFD, these are those SACs and SPAs which are aquatic, or water dependent. The latter include
wetlands such as fens, marshes, bogs and turloughs. All of these are also subject to ongoing and
serious pressures, arising from land alterations, nutrient loads, development and a lack of a coherent
national approach to their protection. A relationship between ecological status, as defined by the
normative definitions of Annex V of the WFD, and “conservation status” defined under Article 1 of the
Habitats Directive appears intuitive, and while management objectives to achieve the respective
environmental targets should clearly be complementary (Cowx et al., 2009), they are not necessarily
synonymous. In some cases they may be conflicting. Some ducks, for example, may benefit from
nutrient enrichment of a water body. Generally, however, it would be expected that high status under
the WFD is compatible with favourable conservation status under the Habitats Directive
(http://www.wfdireland.ie/docs/4_HabitatsAssessment/). The clear implication is that high status is
suitable for species such as the freshwater pearl mussel in rivers, or charophytes in lakes, requiring
unpolluted water and intact habitat; although there is no obvious way to translate the WFD EQR to the
Habitats Directive favourable conservation status. This lack of clarity in aligning quality among the
Natura 2000 network with classification under the WFD provides potential confusion for overall site
designations, monitoring and management. Unlike the agreed water body typology of the WFD, this
has not been applied to the habitat types that provide the foundation of the Habitats Directive SAC
network, and the inadequacy of both the lake and river descriptors in the Habitats Directive is widely
acknowledged (James et al.,2003; JNCC, 2007; Irvine, 2009).
National vegetation classification schemes for lakes developed for the WFD bear only some general
resemblance to those used in the Habitats Directive. A survey of vegetation and water chemistry of
617 lakes in Northern Ireland identified 16 lake groups (Wolfe-Murphy et al., 1992). A scheme for the
UK developed by Palmer (1992) and recently revised using records from 3447 sites in the British Isles
(Duigan et al., 2006), identified eleven plant communities, bearing only some concordance with those
listed in the Habitats Directive. So, in essence the very terminology and approach to site description
between the Habitats Directive and WFD are inconsistent and, therefore, it is not surprising that there
is confusion in trying to harmonise the classification and management for, respectively, favourable
22
conservation and high status water bodies. Basic monitoring under the WFD is not designed to
evaluate conservation status, and hence it cannot be assumed that management for high status
waters under the WFD is commensurate with that for favourable conservation status under the
Habitats Directive. Furthermore, cSACs selected as the best representatives of a particular habitat,
may not equate to WFD concepts of reference state. Inclusion of water dependent habitats merely
compounds this problem. Furthermore, there is no consistent approach to invasive species. Current
EPA policy in lakes that contain zebra mussel cannot be designated as high status but there are a
number of other invasive aliens for which there is no policy on how these affect status class. This is
an area in need of further discussion and action. Management plans have only been prepared for a
small number of cSACs .
While a suite of existing water-related national legislation (see Irvine 2005) can support
implementation of the WFD and Habitats Directive, the protection of cSACs has relied mainly on
prevention of potentially damaging activities through a list of “notifiable actions” that can only be
carried out by consent from the NPWS. These notifiable actions overlap with other legislation leading
to possible confusion as to the responsible authority for these activities, and they are not wide enough
to protect against some forms of damage (e.g. agricultural activities that are acceptable under current
Nitrates regulations but are damaging to highly sensitive water bodies (see below, Section 3.3).
NPWS propose to replace notifiable actions with “activities requiring consent” (ARCs), which will form
the basis for protection against damaging activities under new EU Habitats and species regulation
due to be enacted in 2011. Some site and species-specific ARCs have, in the meanwhile, been put in
place through the NPWS Farm Plan Scheme, where farmers sign up to 5 year contracts that
incorporate prescriptions for agricultural management.
The clear and striking need for conservation objectives, and management, to be aligned and
screened for cross-compliance with WFD objectives, provides an incentive to link these plans to WFD
objectives for high status sites. The opportunity is there for the NPWS to play a key role in agreeing
POMs for both cSACs and high status sites. In support of this the RBDs have already compiled
extensive action points that cover monitoring to administrative arrangements for WFD measures
aligning the Habitats Directive and WFD (Mayes, 2008). The use of an ARCs type, or even an
aligned, system for protection of high status sites seems an obvious consideration. The new Planning
and Development Act (2010) provides a framework to more closely integrate planning with the
implementation of the WFD and Habitats Directive implementation, and is discussed below in Section
4.2.
Safeguarding the interests of nature conservation and ecological quality in high status sites requires
common purpose among the respective agencies.
Formal liaison and data exchange among
agencies with an interest in ecological quality of cSACs is a basic prerequisite, with screening for
compatibility and cross compliance of objectives, monitoring and management under Habitats
Directives and the WFD. Where targets for the WFD may not be synonymous with those of the
Habitats Directive this requires thoughtful resolution.
There is no evidence that relying on the
objectives of nature conservation to be met by the machinery of the WFD will be successful. Under
23
the Habitats Directive the record for producing and enforcing appropriate nutrient management plans
to safeguard SACs from nutrient enrichment has been very lacking (Irvine et al., 2007; Hobbs et al.,
2005). There is an urgent need for the completion of all SAC management plans that are either
compliant with WFD objectives, or which identify potential conflicts. In his paper on the links between
the WFD and Natura sites Withrington (2005) makes the observation that “It should be borne in mind
that the Water Framework Directive is a legal instrument and will be enforced as such. It is to be
hoped, however, that progress can be made in the UK without the need for clarification from the
European Court of Justice”. This applies equally well to the Republic of Ireland.
Key points: Natura 2000 sites and interaction with WFD
While management requirements for protected species and habitats have been compiled, there are
few detailed action plans that are adequately backed-up by legislation or effective policies. The
notable exception to this are the measures to protect the freshwater pearl mussel (Margaritifera
margaritifera) under the European Communities Environmental Objectives (Freshwater Pearl Mussel)
Regulations 2009 (S.I. 296 of 2009). While references conditions have not been set for the pearl
mussel these are likely to be more stringent than agreed for high status rivers under S.I. No 272 of
2009. The pearl mussel plans will provide a level of protection greater than that currenty provided for
by the Surface Water Regulations S.I. No 272 of 2009 in waters in which the mussel occurs. The
pearl mussel plans also provide a precedent for a process that recognises that some sites require
additional measures, as envisaged in Article 11 and Annex VI part B of the WFD. The need for such
additional (or Supplementary) measures apply to a number of other species and habitats, including for
threatened species not listed in Annex II of the Habitats Directive. Both the Natura 2000 and the
WFD are concerned wth the protection of ecological integrity, although the terminology and objectives
are not exactly alligned. In many cases the maintenance of high status sites under the WFD will be
compatible and supportive of nature conservation objectives, but it is important that common
objectives and, especially, differences in objectves are understood and any conflicts resolved among
the relevant agencies. The implementation of protection of habitats and species under the Habitats
Directive has not been satisfactory and there is potential for greater cooperation between the NPWS
and the EPA, as the main statutory bodies charged with coordination and implementation of habitats
and WFD policies and management. A wider approach to the protection of high status sites and their
biodiversity importance has high merit and in keeping with obligations under the Ramsar convention,
and a wider countryside approach conducive to the protection of high status waterbodies.
Consideration could be given to extending statutory protection under the Habitats Directive to high
status water bodies currently not in the cSAC network, and to restoration of water bodies to reach
high status in order to provide a more comprehensive and effective network of high quality aquatic
habitats, including those that may fall below the reporting thresholds for the WFD. The logic of the
Habitats Directive is that there is a requirement for a representative number of cSACs across the
landscape. A similar logic applied to high status water bodies under the WFD would instigate a
gradual process of a better geographical target of high status sites across the country. A higher profile
for high status waterbodies is suggested, with a closer operational and policy connection with the
24
Habitats Directive. The use of an Activities Requiring Consent (ARC) type of system for protection of
high status sites seems an obvious consideration. This can, where appropriate, be facilitated through
the new Planning and Development Act (2010).
3.3.1 The Nitrates Directive (91/676/EEC).
The primary pressure on Irish inland surface waters is emissions of diffuse nutrients from agriculture
(Foy et al. 1995; Allott et al., 1998; Lennox et al., 1998; Jennings et al., 2002; Bartley and Johnston,
2005), with other localised sources, including from farmyards and septic tanks (Arnscheidt et al.,
2007; Edwards and Withers, 2007), accentuating the problem. Ireland has a high percentage of land
cover in intensive agricultural use (see Bignal and McCraken, 1996, modified in Berger et al., 2006),
dominated by grassland-based farming. Grazed pasture can be the major source of P emissions to
surface and ground waters (McDowell et al., 2007; Bourke et al., 2008). The main threat to Irish
estuaries (the Transitional waters of the WFD) is considered to be from nutrient enrichment, although
physical modifications, especially down the east coast are also widespread (Hartnett et al., 2011).
The primary legislative measure to reduce agricultural pressure on water is the Nitrates Directive
(91/676/EEC). A judgement by the European Court of Justice (Case C-258/00, Commission v France)
found that eutrophic waters are required to be addressed under the Nitrates Directive even if the main
source of that eutrophication is from phosphorus (P); hence, the inclusion of P management in the
Nitrates Regulations. The delay in producing a “Nitrates Action Programme”, initially due by 19 th
December 1995 has been described by Flynn (2006) as a sorry saga, reflecting “a lack of political will
in the face of predictable and understandable opposition from the farming organisations”. A Nitrates
Action programme was only eventually enacted following legal proceedings and the threat of fines
from the E.U, and transposed eventually into Irish law in 2005, 19 years late, by the European
Communities (Good Agricultural Practice for Protection of Waters). Subsequent amendments have
been made on an almost annual basis, in 2006, 2007, 2009 and 2010, the latter through S.I . No 610
of 2010 (DEHLG/DAFF. 2010), commonly known as the Nitrates Regulations. These regulations have
been particularly controversial. On one hand the agricultural industry sees them as unduly restrictive.
In contrast, many ecologists see them lacking in providing sufficient protection for surface and ground
waters. For protection of high status waters, it was acknowledged by the DHELG that the Regulations
are not designed for the protection of high status waters (Nitrates Review Expert Advisory Group,
2010). Indeed there is a distinct likelihood that the Regulations will increase pressures from
agriculturally derived nutrients on those sites, for reason summarised below.
Work by Teagasc (Daly et al., 2002) associated measurements of soil test phosphorus (STP) with
river concentrations of MRP, and made a clear and statistically significant distinction between risk of
diffuse emissions of P between peatland dominated soils compared with well drained, predominantly
brown earth, soils. It is clear from Fig 2 of Daly et al. (2002) that high production grassland on
peatland soils, even at low overall percentage of the total catchment, is associated with MRP values
exceeding the river nutrient standard S.I. 272 of 2009 for good status rivers. There are very limited
25
options for further land spreading of P in peatland catchments without impacting water quality. Most
high status rivers are in catchments dominated by peatland. Hence, even the minimal allowance of P
addition to grasslands on peat soils (defined in the Nitrates Regulations as organic content >20%)
provided in Schedule 2 (Table 13 and 15) risks increasing P loss to water, promoting unsustainable
long-term land use and presenting high risk of impact to high status sites.
This relates to the well
known low capacity for peat soils to retain P (Renou-Wilson and Farrel, 2007; Van Beek et al., 2007).
In peatland catchments with high status waterbodies, it is effectively a mechanism to impact those
systems contrary to the objectives of the WFD to prevent further deterioration of ecological quality.
The scale of potential impact of dissolved phosphate can be illustrated by a simple calculation. A
single kg of P when present as PO4 will pollute 40 000 000 litres as a mean of 0.025 mg P l-1,
(equivalent to 25 μg l-1) which is the river standard for high status. One m 3 of slurry is estimated to
contain 0.8 kg P, and a dairy cow is estimated to excrete 13 kg P yr -1; two dairy cows ha-1 are
estimated to be equivalent to an input of 170 kg N ha-1 yr-1 (Table 6 of S.I. No 610 of 2010).
That the Nitrates regulations are insufficient to protect high status sites and surface water, or water
dependent, cSACs relates to the allowance for additional fertiliser additions to Index 1-3 soils (see
Tables 13 and 15 of Regulations), and soil Index 4 soils in circumstances where capacity on other
soils is used up.
In all catchments this provides a mechanism for greater extensive nutrient
enrichment of soils, and concomitant risk of P runoff. In particular, this should be avoided where there
are sensitive receiving waters, including high status sites and water-dependent cSACs. The adequacy
of soil index 3 as a reasonable threshold to be used for the protection of surface and groundwaters is
unproven, and indeed there is evidence to suggest a threshold of less than 5 mg l-1 Morgan’s P is
required for high and good quality water (e.g. Jordan et al., 2000; Sibbesen and Sharpley, 1997; both
of these studies related soil P loss to Olsen’s STP). Report of a 30 year study by Teagasc (Tunney et
al., 2010) demonstrated maximum beef production occurred at 4.1-6.4 mg l-1 Morgan’s P, a
concentration that Torpey and Morgan (1999) considered commensurate with river concentrations of
ca 35 μg l-1 MRP, which is the good:moderate boundary under the Surface Water Regulations S.I. 272
of 2009. A Morgan’s P of 4.1- to 6.4 mg l-1 lies across the boundary of Soil Index 2-3. The upper
boundary of Soil Index 3 equates to 8 mg P l-1. It is particularly relevant that much of the evidence
quoted above comes from Irish State agency funded work or their own research, and it is clear that
the recommended target of Soil Index 3 provides a policy allowance for the eutrophication of surface
waters. As this is counter to the evidence, it can only reflect political compromise in the setting of
these boundaries.
Results of long term monitoring in Northern Ireland demonstrated that while a “maintenance” fertilizer
application of 8.5 kg P ha-1 resulted in no net P surpluses it was, nevertheless, associated with
nutrient run-off far above commonly accepted thresholds relating to eutrophication of receiving waters
(Watson and Mathews, 2008). The allowable additions of P to grassland given in Schedule 2 (Tables
13 and 15) of the Nitrates Regulations are well above any sensible agronomic need, which are likely
to be less than 20 kg ha per year when soil P is low (Ryan and Finn, 1976; cited in Tunney et al.,
1997), and less at moderate soil P concentrations.
26
The Nitrates Regulations are designed to reduce risk of nutrient enrichment from farming, but
effectively allow for an increase across all catchments to Soil Index 3 (re Article 15 and 16 and
associated Schedule 2 Tables 11, 13 and 15). Furthermore this is an average concentration across a
farm holding, which in an enabling mechanism to increase local sources to Index 4 and above. The
general premise that soils can reach Soil Index 3 before these pose a risk to water quality is
misguided, as it allows for further P emissions to water (Donohue et al., 2005; Styles et al., 2006).
Where farms which have received derogation (European Commission, 2007) for higher nutrient
loadings than the 170 kg N yr-1 allowed under the Nitrates Directive are within or close to SACs, this
provides a high risk to the favourable conservation status of those sites. It appear that there is no
requirement, or practice, for DAFF to consult with NPWS in approving derogation farms, or to
consider risks this may pose on receiving waters in the proximity of those farms. If so, this seems a
clear example of a failure of cross compliance and dialogue across Government Departments.
There would appear to be a fundamental conundrum inherent within the Nitrates Regulations such
that while many soils have high P content which should negate any rational argument for additional
application of P as a fertiliser, there is still the need to dispose of animal waste. While land spreading
is the main option, and given the densities of cattle and current economics related to intensive
farming, there is a fundamental difficulty in maintaining current agricultural intensity (never mind
targets for increasing it, DAFF 2010), protecting high quality ecosystems, and achieving WFD
compliance. Until these conflicts are accepted and full cooperation between DAFF (now reconfigured
as the Department of Agriculture, Food and the Marine (DAFM)) and the, newly configured,
Department of Environment, Community and Local Government (DECLG) to work together to resolve
these problems, pressures on water quality and, where they occur in agricultural landscapes, high
status sites, will continue. This will lead to a likely further loss of those sites and, therefore, failure of
Ireland to comply with the WFD. Indeed there is risk that pressure on the high status sites will
increase as a direct result of the Nitrates Action Plan. Quantifying that risk, in order to make a well
founded judgement on the extent that agricultural pressure, albeit at low-moderate intensity, may
impact high status sites requires information on land use at appropriate scales. Mechanisms to
overcome political or practical obstacles to this are required urgently.
There is some progress
towards this through the liaison between DAFM, and the local authorities in relation to the inspections
and data availability for compliance with the Nitrates Regulation, but further development of these
types of arrangements is required if decisions on land use and protection for high status sites are to
be evidence based.
Improved storage facilities for slurry in the Nitrates Regulations does not reduce net P content of
slurry per se, only the timing of disposal and the temporal risks to surface waters. It might, ironically,
in some cases add to the problem in inland waters because winter disposal of slurry can be subject to
high flushing rates through surface waters. Under Schedule 4 of Nitrates Regulation 2010, slurry
cannot be spread until after mid-late January depending on the region.
There is, then, a large
motivation to dispose of as much slurry as permissible. Rainfall occurring soon after this will pose a
high runoff risk to surface and ground waters. Spreading large amounts of slurry in early spring could,
27
because of high growth potential of algae, available silicates in the water (that support diatoms), and
likelihood of high rainfall that flush nutrient to surface waters, accentuate spring increases of algae in
standing waters. In Northern Ireland, February has been shown to be a month of high risk of nutrient
loads to surface waters (B.Foy, AFBI, pers comm.) and there is likely to be a review by the E.C. of the
closed season for slurry spreading.
In light of work by the U.K. Agricultural Development and
Advisory Service (ADAS), that has shown consistently high pollution in February and March, the U.K.
government are reviewing the closed period.
The essential wider point is that given current soil P concentrations, there is little logical argument for
further additions of P to grassland for agronomic production. Therefore, the Nitrates Regulations are
less about water quality than the maintenance, or expansion, of agricultural practice (see Minister
statement, Section 3.3.2). Under Schedule 3 (Table 15) grasslands of Soil Index 3 are allowed a
loading of 20 kg P per annum. Table 6 of Schedule 3 suggests an average dairy cow excretes 13 kg
P p.a. There is, therefore, an equivalence of 1.54 dairy cows commensurate with the maximum
loading for an Index 3 soil (note also the view, above, that 170 kg N yr -1 is the equivalent intensity of
about 2 dairy cows ha-1). Work done by Irvine et al (2000) associated cattle densities of >1 ha -1 with
risk of eutrophication of lakes (ca 30 µg TP l-1). Recently, similar values have been reported from
Northern Ireland (R.H. Foy, AFBI, pers comm.). Intensity of grassland farming commensurate with
high status water bodies needs to be considerably less.
While the possibility, allowed within the Nitrates Regulations, for increasing P loads to catchments
containing high status sites provides the fundamental cautionary tale, this is accentuated by a number
of specific provisions:
Under Article 13 (1), if the storage of livestock manure on a holding is less than the capacity specified
in Article 9, 10, 11 or 12, as appropriate, manure can be transferred to a person authorised under and
in accordance with the Waste Management Acts 1996 to 2003 or the Environmental Protection
Agency Acts 1992 and 2003 to undertake the collection, recovery or disposal of the manure. This
increases risk of nutrient loss to water at the catchment scale, and is against the principles of
sustainable land management in allowing land outside the holding to be used for the spreading of
organic waste. It is even more worrying if waste is permitted to be spread outside the relevant
catchment. Spreading of waste from a holding to other land should not be permitted if this might affect
protected areas, as defined in Annex IV of the WFD, or high status sites.
Article 16 (Table 13) allows for fertilisation of P of 35, 25, and 15 kg ha -1for P index of 1,2 and 3,
respectively. This provides for gradual enrichment of soils which are currently un-enriched and hence,
greater net potential for P emissions to water. Furthermore:
a) the provision that soil tests are
required only once every 6 years (see Article 16 (2) (c)) could very easily lead to enhanced loss of P
to water because of interim build up of soil P; and b) that P addition to peatland soils (comprising
>20% organic matter) at a rate equivalent to a soil index 3 soil provides a high risk of P loss, because
of low capacity of peat to hold P. Addition of P to peatlands soils should be prohibited unless
demonstrated to have minimal potential impact. These should apply especially (under the WFD) to
28
protected areas of peatlands or those areas with drainage waters to high status sites. Furthermore,
under Article 16 (2), the P index for soil shall be deemed to be phosphorus index 3 unless a soil test
indicates that a different phosphorus index is appropriate in relation to that soil.
However, P load of
soils with a lower index can receive more P. This should not be the case in catchments of high status
water bodies, as there is substantial evidence that low amounts of P can impact those sites. In N.
Ireland, under the Phosphorus (Use in Agriculture) (Northern Ireland) Regulations 2006 farmers are
required to demonstrate, through soil tests every 3-5 years, the need for application of inorganic with
a soil P test, and to deduct P applied in manure in calculating inorganic loading (Defra, 2010). This
puts to onus on demonstrating, rather than assuming, a need.
Under Article 17 (1), chemical fertiliser shall not be applied to land within 2 m of a surface
watercourse. Although buffer strips are a common mitigation strategy for rivers (Haycock et al., 1997;
Newson, 2010), and shown to attenuate nutrient loads from land to water, their effectiveness can be
highly variable and dependent on design and local conditions (Vought et al., 1993; Schmitt et al.,
1999; Wenger, 1999; Polykav et al., 2005; Liu et al., 2008; Collins et al., 2009; Hoffmann et al., 2009).
This has posed a difficulty for policy makers. Common sense would suggest, however, that 2 m is an
insufficient boundary. The general premise that the wider the boundary from a water course the better
would seem sensible for nutrient buffering.
Buffer strips provide a non-linear mitigation of nutrient and sediment mitigation, with an initial steep
slope, then tending towards an asymptote, with about a mean 80% retention rate at 30 m, this varied
with slope
(Mayer et al., 2007; Collins et al., 2009).
Reviewing effectiveness of buffer strips,
Desbonnet et al. (1994), indicated an efficient width of vegetated buffers for sediment sediment
removal was 25 m. A width >10 m was often associated with > 80% removal of sediment (Dillaha et
al., 1988; 1989), but that for long-term benefit >30 m was recommended (Castelle et al., 1994).
Buffer strips are less effective for nitrate removal, but removal rates of between about 50 and 95% for
ca 9 m riparian buffers strips have been reported (Peterjohn and Correll, 1984; Magette et al., 1989),
although Wenger (1999) in his review of buffers concluded that 15 m is probably necessary for most
buffers to reduce nitrogen concentrations effectively.
The lower efficiency of buffer strips to prevent
movement of nitrate compared with sediment relates to its high solubility and mobility. Removal
through microbial mediated conversion to nitrous oxide or ammonium is a feature of the nitrogen
cycle, with most efficient denitrification associated with low redox potential in wetlands (Hanson et al.,
1994; Howarth et al., 1996; Fennessy and Cronk, 1997; Valiela et al., 2000). Overall, both grass and
forested buffers have been shown to reduce nitrogen and phosphorus as long as there is sufficient
width, raging from ca 10-20 m (Wenger, 1999). Similar widths as reported for nitrogen have been
shown to substantially reduce phosphorus transport, although riparian buffers are typically effective at
short-term control of sediment-bound phosphorus but may have low net dissolved phosphorus
retention (Lowrance, 1998). Phosphorus trapped in buffer strips may also be leached out over time
(Osborne and Kovacic 1993, Mander 1997), as phosphorus is only immobilised, not removed from the
system, unless through riparian zone harvest, especially once the buffer is saturated.
29
New Zealand, which has quite similar grassland farming as Ireland advocates 3-5m buffer strips near
water courses and lakes (see Section 7.5). In North Carolina, U.S., minimum buffer strips for sensitive
streams is ca 15m (see Section 7.2), following general guidance that steam mitigation projects should
have wooded buffers 50 feet wide in coastal plains and 30 feet in the mountains (U.S. Army Core of
Engineers et al., 2003). The U.S. regulatory bodies do not give any credit (relating to management
subsidies) for buffers less than 15 feet (5 m), as their effectiveness below this width is highly uncertain
(Wenger, 1999). For high status waters, it would seem prudent to have vegetated buffer strips of at
least 20 m, which can also act as functioning riparian woodlands, as suggested by Schoumans et al.,
2011). Harvesting of riparian vegetation can account loss rates of 4-15 kg P ha-1 yr-1 (Hoffmann et al.,
2009).
The overall conclusion from the work on buffer strips is that it can be part of the management of
riparian zones to mitigate nutrient run-off, but the recommended widths may typically remove 10 m or
more of agricultural margin. For low intensity catchments such as present in the catchments of high
status waters, this sort of buffer may not impact much on farm income and could form part of costeffective agri-environment schemes.
There is also a need for greater attention to potential critical source areas of nutrient mobility. High
status water bodies should have at least the same level of protection as protected areas. Article 17
(2) (e) provides for a 15m buffer of organic waste from exposed cavernous or karstified limestone
features (such as swallow-holes and collapse features). These sites and dolmines (landscape
depressions that provide access to the groundwater) provide for a ready conduit to groundwater fed
turloughs, lakes and rivers.
Under Article 18 (6) of the Nitrates Regulations in “Extreme Vulnerability Areas on karst Limestone
Aquifers”, soiled water shall not be applied to land in quantities which exceed in any period of 42 days
a total quantity of 25,000 litres per hectare, unless the land has a consistent minimum thickness of 1m
of soil and subsoil combined. Extremely vulnerable karst areas include potential impact on
groundwater, turloughs or, through underground conduits, to rivers. The 25,000 litre limit appears
entirely arbitrary, and note estimation of potential P dispersion from inorganic phosphate and slurry
given above (this Section). No soiled water on these areas should be spread unless demonstrated to
have no potential to contaminate high status sites, ground waters and turlough cSACs, and risks
failure of cross-compliance with WFD and Habitats Directive.
The EU Common Agricultural Policy (CAP) was for 40 years about maximising production with little
attention to minimizing environmental impact. Traditionally, guidance for P application to grasslands in
Ireland was higher than in the UK, and led quickly to excessive amount of inputs of fertilisers (Tunney
et al., 1997). This philosophy to maximise production, without sufficient thought to the impact on the
environment, prevails, as evidenced by the aspirations for the future of Agriculture under the Harvest
2020 vision (DAFF, 2010). CAP reform and agri-environmental schemes, based on shifts in the CAP
subsidies programme, have done little to redress this position, although inorganic P use has declined
30
markedly from about 70,000 tonnes p.a in the early 1970s to about 20,000 tonnes p.a currently.
Nevertheless the legacy of enrichment, continued use of inorganic P fertiliser, and the problems
associated with disposal of animal manure provide on-going problems. The Rural Environmental
Protection Scheme (REPS) was never properly evaluated for environmental benefit, with no
coordinated system of monitoring (Finn et al., 2009; Feehan, 2002). The failure to use agricultural
environmental protections subsidies effectively is criticized by the OECD (2010) environmental
performance review of Ireland. The unfortunate, but regrettable, overall conclusion is that the delays
in implementing the Nitrates Action Programme, the allowance of high rates, and more extensive
application, of P loads to fields, and the reluctance to introduce adequately monitored agrienvironment schemes, is that the response of the Irish Government to protect surface and
groundwaters has been insufficient. Similar criticisms apply to the implementation of the Birds and
Habitats Directive. There are other more effective and imaginative ways to address the economic
concerns of agriculture, but this requires a much more holistic approach to farming, land management
and environmental protection. These and wider issues relating to the E.U Common Agriculture Policy
(CAP), including the use of Strategic Environmental Assessment are considered further in Section 4.
3.3.2. Testing and implementing the Nitrates Regulations
Notwithstanding the critique given above, an Irish Agricultural Catchments Programme (ACP) has
been established to investigate the effectiveness of the Nitrates Regulations (Fealy et al., 2010; Wall
et al., 2011). This will run till 2015, covering six catchments. It is a requirement of the Nitrates Action
Programme (European Commission, 2007; Collins and McGonigle, 2008). It needs to cover
catchments with a range of catchment activities, including those with derogation farms, and requires a
“robust monitoring and evaluation experiment be established and implemented to assess the
environmental consequences of the policies” (Fealy et al., 2010). Five river catchments were selected
that could assess maximal agriculture intensity, excluding confounding landuses such as urban (or
housing density), forestry and peatland. Additionally, a catchment with important groundwater
transport was added to include a ‘catchment’ with a low residence time aquifer and good existing
hydrometric data. The ACP is the most intensive medium-large scale investigation of agricultural
derived nutrients conducted in the Republic of Ireland. It is designed to provide high spatial and
temporal resolution data that will inform the National Nitrates Action Programme. Similar
investigations are occurring in other EU member states. These programmes are designed to link
nutrient sources and pathways to receptors (Haygarth et al., 2005). While this programme is not
designed to investigate effects on high status waters, or includes any such waters in the programme,
it does include, in the Cregduff catchment in south Mayo, a karst region that contains waterdependent cSACs. This, however, is coincidental as the criteria for the ACP does not included the
presence of high status waters, and it is not within the remit of the ACP to identify measures relating
to high status water bodies (P.Jordan, University of Ulster, pers comm). It is difficult to see the
feasibility of extrapolating ACP results for protection of high status sites, and no methods have been
suggested of how to do this.
31
It is not known if the time scale of the Irish ACP will be sufficient to establish proposed improvements
commensurate with good ecological status, owing to time lags between reduced nutrients in soil and
water quality improvements (Smith et al., 2003; Herlihy et al. 2004; Watson et al., 2007; Cherry et al.,
2008; Schulte et al., 2010) or loads and ecological response (Bowes et al., 2010; Neal et al., 2010).
Jeppesen et al. (2005), investigating a range of Danish lakes, suggest that reduced external
phosphorus loading resulted in a new equilibrium for total phosphorus within 10 to 15 years, but this
could be much longer for many biological variables. Recovery times in enclosed coastal and estuarine
systems can as also be long and, generally, greater than 5 years (Kauppila et al., 2005; Borja et al.,
2006, 2009cd; Uriarte and Borja, 2009). Hering et al. (2010) consider that “we cannot expect
European aquatic ecosystems to fully recover within 15 or even 30 years from over a century of
degradation”. The time scale for recovery is likely to be related to magnitude of impact, but even for
modestly impacted sites is likely to be several years (Donohue et al., 2010). Phosphorus
concentrations, and associated algal productivity, in the River Frome(U.K)
declined following
introduction of phosphorus stripping at sewage treatment plant (Bowes et al., 2011).
Although
significant, the decline in mean summer soluble reactive phosphorus from a maximum of 190 µg l-1 in
1989 to an average mean for 2007-2009 of 49 µg l-1 is still at a concentration that would not be
commensurate with high ecological status, and load apportionment models suggest about a 70%
contribution in the period 2006-2009 of phosphorus from the landscape (estimated from Fig 6, Bowes
et al., 2011).
The recovery of a water body in response to reduced loading at the field scale depends on an
interaction between initial P concentrations, soil properties and pathways. For the protection of high
status waters, it is quite disconcerting that Schulte et al. (2010) maintain the view that care should be
taken not to reduce field P to below what they consider optimum index 3 (up to 8 mg P l -1), using an
example from Donegal. This philosophy is simply not appropriate for the protection of high status
water bodies, and is driven by the premise that maximising productivity takes precedence over other
views.
For high status waters a quite different paradigm is required. The test of the Nitrates Action
Programme for compatibility with good status water is also fundamentally dependent on 100%
compliance of farmers in the study catchments.
This links with a high quality of farm inspections and adequate reporting and data management,
although nationally only approximately 1% of farms are to be inspected by DAFM on behalf of Local
Authorities in compliance with the Nitrates Regulations (although further inspections are conducted by
DAFM, and some by Inland Fisheries Ireland), and about 5% in catchments with high status sites
(Source: Agricultural Investigation Working Group). There are currently some shortfalls in crosscompliance reporting, training of local authority staff involved in the farm inspections, and data
availability across agencies, although these issues are currently being addressed. It is not, however,
encouraging that out of 1599 farm inspections finalised by DAFM on behalf of the Local Authorities up
to mid-May 2011, 35.5% received a penalty for non-compliance, mostly (98%) related to the farmyard
e.g. collection, management and storage of manures and organic fertilisers, and failure to minimise
soiled water (DAFM, 2011).
32
For high status waters there is a need to minimise potential nutrient diffusion from agriculture, and
particular attention and safeguards need to be in place. The mechanisms to protect high status waters
are unlikely to be found in the current policies contained in the Nitrates Regulations (DEHLG/DAFF,
2010), and will require additional measures, including further improvements of best practice,
educational programmes and fiscal incentives. The EU funded COST Action 869 on the mitigation of
nutrient movements to water provides a valuable library of information, freely available at
www.cost869.alterra.nl, including a summary of options (Schoumans et al., 2011). A review of similar
mitigation strategies employed in the U.K. is provided by Haygarth et al. (2009) and Cuttle et al.
(2007), and in New Zealand by Monaghan et al. (2008) and Quinn et al. (2009). A comprehensive list
and review of mitigation measures to reduce nutrient loss from agriculture is given by Cherry et al.
(2008). Most cost-effective for intercepting nutrients and sediment in farm effluents appear to be
wetlands, sedimentation ponds, and buffer strips (see above, Section 3.3.1).
Fundamentally, there is a requirement for maintaining reduced catchment inputs and low-level of
potentially impacting activities to protect high status water bodies. This requires low-intensity landuse.
Mitigation of impact plays its role, but minimising source of the impact is the most effective measure.
Precedent for achieving this in a high value landscape can be found in the Burren Life Programme
(www.burrenlife.com; see Section 6.2). A much greater opportunity lies, however, at the heart of the
driver of Irish agriculture, the E.U. Common Agriculture Policy and its ongoing reform (discussed in
Section 4.5) and lessons from Northern Ireland where a series of mechanisms, including fiscal and
stakeholder dialogue have resulted in significant reduction of phosphorus over a 15 year period
(1994-2009) in 245 monitored rivers in Northern Ireland, with 75% showing a significant decrease and
only 0.8% showing an increase (AFBI, 2009).
In general, there is increasing awareness of the risks of nutrients, and other pollutants, from land
management to water resources. There is also greater emphasis on the use of farm advisory services
to support reduced impact from farming on water at both national and EU level (Berglund and
Dworak, 2010). For high status and other sensitive water bodies, the current thresholds of risk in the
underlying policies and associated Regulations would appear not to be sufficiently stringent, or the
general view sufficiently cautious. It is also noteworthy that Government expectations of the ACP is to
support the expansion of intensive agricultures. A recent (14th September 2011) statement from the
Minister for Agriculture, Food and the Marine, Simon Coveney, was that “the new scientific knowledge
being generated by the Agricultural Catchments Programme will be critical to the sustainable
expansion of Irish milk and meat production from grass and therefore will be critical to achieving the
ambitious growth targets of Food Harvest 2020”.
33
Key points: The Nitrates Action Programme
The Nitrates Action Programme in Ireland has been effected through several iterations of the Nitrates
Regulations, most recently S.I. No 610 of 2010. These are designed to ensure that agriculture is
compatible with attainment of good ecological and chemical status under the WFD. Whether it is
condusive to the WFD environmental objective of good status is debatable, as it is an agreed position
between Ireland and the E.C. The 2010 Regulations are not designed to meet the more strigent
objectives for high status water bodies, so additional measures are required.
The Nitrates
Regulations may even lead directly to further pressure on high status waters as the landscape
becomes more homogeneous towards a general target of Soil P index 3. Water bodies in peatland
catchments, or those in very dynamic hydrological setting, such as turloughs, may be particularly
impacted unless additional safeguards are forthcoming. A range of mitigating measures, such as
those found on the COST website www.cost869.alterra.nl, may reduce impact, but essentially a
fundamental paradigm shift is required to link the protection of high status water bodies with
agricultural practice. Development of agri-environmental fiscal measures provide a useful mechanism
for protection. In catchments containing water dependent SACs, a closer partnership among the
relevant agencies wthin the umbrella of River Basin Management is essential.
3.4 The integrated Pollution Prevention Control Directive (96/61/EC).
Directive 96/61/EC on Integrated Pollution Prevention and Control (IPPC), transposed into Irish law in
the Protection of the Environment Act 2003, has the objective to prevent significant impact on the
environment through a process of licensing of industrial, including some agricultural, operations such
as intensive pig units. Licensing is done by the EPA, and takes account of the particular
circumstances and location of each application. While licensing has clearly improved environmental
performance of industry and, overall, reduced potentially damaging emissions there are still concerns
of under-reporting and a lack of production data in some sectors (Styles and Jones, 2010).
Nevertheless, there is an extensive obligatory process for licensed facilities to produce Annual
Environmental Reports (AERs) on their performance. Licensing requires consideration of potential
impact on cSACs and other protected sites, but there is no special provision for considering impact on
high status sites outside that network other than that discharges must comply with legislative
requirements. For discharges to rivers, therefore, emissions must comply with the Surface Water
Regulations S.I. No 272 of 2009, and allowable concentrations of nutrients to high status waters are
lower concentrations than for good status waters. For discharges that may impact high status sites,
there should be a responsibility that emission limits shall not compromise environmental objectives.
Under the current Surface Water Regulations “A public authority that authorises a discharge to waters
shall .....[in setting emission limits] aim to achieve the environmental objectives”. There is a current
process of the EPA licensing local authority waste water treatment plants. Periodic review of licences
34
should account for combined effects from any other discharges to the water body that may impact on
water quality. This would also be in keeping with the EPA view that all high status sites are at risk of
failing to meet their environmental objectives. Licensing of discharges to high status waters would
usefully be subject to Appropriate Assessment, as required for discharges to water dependent SACs
under Article 6(3) and 6(4) of the Habitats Directive.
Key point: IPPC
All IPPC or waste licensing that may impact on high status waters must set Emission Lmit Vaues that
do compromise the water body status, whether this license is providedfor industrial or waste water
treatment operations. The EPA might consider the merit of introducing an Appropriate Assessment to
assess combined effects of discharges to high status waters, as is done for discharges to cSACs
under the Habitats Directive.
3.5 Urban Waste Water Treatment Directive (91/271/EEC)
Local authorities have built over 90% of the infrastructure needed to comply with the Urban
Wastewater Treatment Directive (Directive 91/271/EEC). Significant resources continue to be
allocated to address this issue (€2.3 billion was invested in wastewater treatment in the period 20002006, and €2.5 billion allocated for 2007-2013). Compliance on secondary wastewater treatment
facilities has increased to >90%, compared with 25% at the start of 2000 (DEHLG, 2010). Investment
in rural areas is supplemented by block grants (€85 million for 2010) under the Rural Water
Programme. Licensing of local authority waste water treatment plants (WWTP) has been carried out
since 2007 by the EPA under the Waste Water Discharge (Authorisation) Regulations (SI No 684 of
2007). Licensing has been prioritised based on risk assessment, addressing the larger municipal
discharges first, but will include licensing of moderate to small municipal discharges, to help meet
obligations of the WFD. The 2007 Regulations require licensing for all discharges serving >500
population equivalents (p.e). Smaller populations require to be certified by the EPA.
The EPA
estimate, given current resources that it will take up to 5 years before this licensing is completed.
There are 21 WWTPs that serve up to 3,500 p.e within high status waterbody catchments (RBMPs,
2010). All have been either licenced or have a certificate of authorisation. Some lakes where water
dependent Annex 2 listed species occur, and some coastal lagoons, have been assessed as being at
unfavourable conservation status with small scale and larger WWTP noted as a contributory factor
(DEHLG, 2008).
At-risk waters as defined by the Water Framework Directive are prioritised for
protection under the Urban Waste Water Treatment Regulations. This includes those with limited
assimilative capacity; waters that contain sensitive species or habitats; and waters used for water
abstractions, fisheries, shellfish production or recreation.
While targets for discharges may be set by Emission Limit Values (ELVs) compatible with high status
waters, a precautionary approach would minimise loads as far as possible. For this, tertiary nutrient-
35
reduction treatment would be necessary. The Combined Approach in the Waste Water Discharge
(Authorisation) Regulations 2007 require water services authorities to comply with emission limits for
the discharge of wastewaters to water bodies arising from the stricter of either the Urban Waste Water
Regulations (S.I. No. 254 of 2001) or emission limits based on achieving the environmental quality
standards. While these are currently defined by the Surface Water Regulations S.I. No 272 of 2009
and the European Communities Environmental Objectives (Freshwater Pearl Mussel) Regulations
2009 (S.I. 296 of 2009), Article 5 of S.I. No. 254 of 2001 allows for more stringent requirements than
those specified in the Regulation, where this is required to ensure that the receiving waters satisfy any
other relevant Community Directives. This could be applied to high status sites, but requires a
reassessment of the current standards, or possible need to address each high status site on an
individual basis rather than the current approach based on river typology.
Under Annex II of the Urban Waste Water Treatment Directive,a Sensitive area (requiring priority
attention) must be designated if it is, inter alia, a natural freshwater lake, other freshwater body,
estuary, or coastal water which is eutrophic or which in the near future may become eutrophic if
protective action is not taken. A number of such water bodies are identified in the Republic of Ireland
by the Urban Waste Water Treatment Regulations (2001) and its amendment of 2010, based on
traditional criteria of “eutrophic” following OECD (1982) criteria. Eutrophic, however, only relates to
a body of water which is impacted through nutrient enrichment, and this term applies to waters that lie
along a continuum from low to high nutrients. Low nutrient water bodies can be impacted from nutrient
enrichment, as well illustrated by the shallow and internationally important Lough Carra (see
www.loughcarra.org). There seems nothing to prevent high status water bodies in receipt of waste
water discharges from been classified as “sensitive” under the Urban Waste Water Treatment
Regulations in order to support sound management.
While the majority of phosphorus emitted from waste water treatments plants, and domestic systems
is from human origin (humans excrete about 0.6 kg of phosphorus per annum (Wolgast, 1993), there
is also about a 4-6% contribution from P in washing detergent in Northern Europe (de Madariaga et
al., 2009). Council Regulation (EC) No 648/2004 on detergents requires labelling of detergents and
the biodegradability of the surfactants they contain, but there is now an intention to amend Regulation
(EC) No 648/2004 by introducing limits on phosphorus compounds in household laundry detergents.
This will reduce loads to freshwaters by about 5%. The E.U. Environmental Committee has proposed
that severe restriction on household laundry detergents due to be implemented from 1 st January 2013
is extended to household dishwasher detergents by 2015.
Key points: WWTP discharges
As for IPPC or waste licensing, WWTPs that may impact on high status waters must set Emmssion
Lmit Vaues that do not compromise the water body status. All plants whose discharges have the
potential to impact high status sites should be licensed, irrespective of capacity. Phosphorus
emissions can be reduced through use of non-P detergents. Further liaison with the Soaps and
Detergents Industry for development on zero or low P detergents is recommended.
36
This could
include public education, especially in parts of the country where high status sites are concentrated.
High status water bodies should be designated as sensitive waters under the Urban Waste Water
Treatment Regulations (S.I. No 684 of 2007).
3.6 The Sewerage Sludge Directive (86/278/EEC)
The Urban Waste Water Treatment Directive (91/271/EEC) banned the marine disposal of waterwater
sludge. The waste managment (use of sewerage sludge in agriculture) Regulations (S.I. No. 148 of
1998) transposed the Sewerage Sludge Directive (86/278/EEC) under the Waste Management Act,
1996. S.I. No. 148 of 1998, and its amendement of 2001, limits the heavy metal content of sludge
spread on land and needs to have regard to pH and nutrient content of receiving soils. In this way it
links with the Nitrates Regulations of 2010; although Article 8 (1) and (2) (b), (c) and (d) of S.I. No.
148 does not apply to sludge from septic tanks or from sewage treatment plants with a treatment
capacity below 300 kg BOD 5 per day, corresponding to a population equivalent of 5,000 persons.
Nevertheless, the local authorities have developed sludge operational management plans for the
managment of sludge arising from all souces including WWTPs, septic tanks, industry and agriculture.
The allowed
phosphorus is as specified in the Nitrates Regulations 2010; hence they offer no
additional protection for high status water bodies.
Furthermore, because where sludge is used
regularly in agriculture, soil shall be analysed at a minimum frequency of once in ten years, it
effectively allows for a build up of phosphorus. It is not known how effective sludge management is,
but the location of spreading
from agriculture slurry or domestic tanks is often not known, or
recorded. Spreading of slurry and domestic waste comprises a significant risk to ecological quality
and public health. It would appear self-evident that spreading of slurry needs to be well regulated,
with volume and location of spread-lands notified to DAFM and the local authorities, as part of an
operatior’s licensing and contract conditions. Similiar arrangements should be put in place for
spreading of waste from small or domestic waste treatment facilities.
Key points:Sewerage Sludge
The licensing condition, and location of speadlands should be clearly idenified and available to
DAFM, local authorities and the EPA.
37
3.7 Environmental Impact Assessment Directive (85/778/EEC) as amended by Directive
98/83/EC.
The implementation of the Environmental Impact Assessment (EIA) Directive is inextricably linked
with Planning and IPPC licensing (see Case C-50/09 Commission v Ireland, Judgment of the Court of
Justice 3 March 2011). While an EIA must take account of water, and flora and fauna including regard
to protected areas, there is no specific legislation that requires assessment of potential impact on
high status sites. There is a general recognition in the RBMPs of the need for strengthening the
statutory basis for integration of water quality objectives with the planning system. The Planning and
Development (Amendment) Act (2010)
goes some way towards that, and strenthens the link
between EIA and the Habitats and Birds Directives, but will require close liaison of local authority
planning sections with local authority water management as well as bodies such as the EPA. This will
necessitate that further consideration is given to screening of proposed projects that might impact
high status sites and, where deemed approriate, that an EIA is carried out.
The RBMPs (2010) refer to a proposal for new legislation that will provide for prior consideration of
the nature, location and cumulative effects of certain agri-development projects to ensure that the
obligations under the EIA Directive are fully met. This is timely, and in response to the November
2008 Court of Justice (Case C‑66/06) ruling that Ireland was over reliant on size thresholds to
determine whether an EIA is required in relation to certain agri-developments. A follow up ruling of the
European Commission took Ireland back to the ECJ for failing to implement the 2008 ruling, which
included water management projects for irrigation or land drainage. The Court ruled that Ireland did
not consider sufficiently sensitive countryside features such as loss of wetlands in determination of
whether an EIA was required. The ECJ ruled that small projects can have significant impacts on
important nature sites, and that Ireland uses very high thresholds to select projects for assessment.
Failure of Ireland to respond to that judgement, prompted the Commission to refer the case back to
the ECJ, asking for fines to be imposed (IP/10/313).
This has relevance for the protection of high
status sites, some of which may be impacted by drainage or land reclamation. In general it provides
for greater “wider countryside“ protection of wetlands that lie outside the cSAC network.
Key points: EIA
EIA needs to incorporate the possibility of impact on a high status water body. Location of high status
sites need to be contained within the local authority GIS layers, and the EPA should be a statutory
consultee, unless provision is provided to ensure appropriate local authority expertise, so that there is
an authoritive and competent opinion provided on all proposed developments that may impact those
sites.
38
3.8 The Bathing Water Directive (76/160/EEC), as amended by Directive 2006/7/EC.
The Bathing Water Quality Regulations (SI 79 of 2008), which transposed the new Bathing Waters
Directive (2006/7/EC) establishes a new classification system for bathing water quality, requiring
monitoring and management plans to preserve, protect and improve the quality of bathing waters.
These relate mainly to human health and bacterial contamination. The main relevance to high status
sites is that bacterial contamination reveals a pathway for a contamination source and, therefore, can
be concurrent with other pollutants, particularly nutrients.
The Bathing Waters Regulations allow for new bathing areas to be added annually. Cork County
Council has taken the proactive step to only accept a nomination for a bathing water if the catchment
upstream is at good status. They will, therefore, not consider a bathing area where there are known
water quality issues.
Key point: Bathing Waters
Failure to reach a bathwater standard can indicate sources of pollution, and acts as a check for high
status designation
3.9 The Drinking Water Directive (76/160/EEC, as amended by Directive 98/83/EC),
Plant Protection Products Directive (91/414/EEC), and Major Accidents (Seveso)
Directive(96/82/EC)
There are no particular aspects of these Directives that relate specifically to high status sites. The
Drinking Water Directive (Directive 98/83/EC) sets standards for potable water supply, but which may
be treated to remove pathogens. Water extracted from high status sites would be expected to be low
in nutrients and pathogens, therefore, requiring a low level of treatment. High faecal bacteria counts
in source water can indicate a source of pollution that may be associated with high nutrient
concentrations. The Plant Protection Products Directive (91/414/EEC) requires Member State
authorisation for plant protection products (PPPs) to provide a safeguard for human health and the
environment. Harmful substances are subject to a maximum allowable concentration (MAC), listed in
Table 11 of the European Communities Environmental Objectives (Surface waters) Regulations S.I.
No 272 of 2009. The Major Accidents (Seveso) Directive (96/82/EC) requires provision to be made
for emergencies, including unplanned emmissions, for major industrial facilities. In Ireland these
would tend be of potential risk more to coastal than inland sites and not in the vicinity of high status
water bodies.
39
3.10 Conclusions.
The Basic Measures identified by the WFD and listed in Annex VI fail to provide an adequate policy
framework to protect high status sites. Only one measure that relates specifically to high status sites
under the WFD has been proposed. This, for the pearl mussel, may not provide stringent enough
protection for its purpose, but does provide a useful precedent for the process needed for other
possible measures, as outlined in Annex VI, part B, of the WFD.
Section 4: Other Policies Applicable to the Protection of High
Status Sites.
4.1. Water Pollution Acts
Under the European Communities Environmental Objectives (Surface waters) Regulations S.I. No
272 of 2009, “Point source and diffuse source discharges liable to cause water pollution are prohibited
except where subject to a system of prior authorisation or registration based on general binding
rules”.
This
specifically
includes
a
number
of
controlled
substances
(cadmium,
Hexachlorocyclohexane, mercury, carbon Tetrachloride, DDT, and Pentachlorophenol) under a
number of Regulations of the Local Government (Water Pollution) Act, 1977 and 1990. These Acts
provided an initial the framework for the prevention or control of water pollution that has been largely
superseded by the WFD and, in Irish legislation, by S.I. No 272 of 2009. Local Authorities retain the
power to prosecute under the Water Pollution Acts to inter alia prosecute for water pollution offences;
issue notices or obtain a High Court injunction to effect acessation of polluting activities and remedy
impact of the pollution; and make bye-laws regulating certain agricultural activities so as to prevent or
eliminate pollution of waters, and require farmers to prepare nutrient management plans with the aim
of ensuring that nutrients applied to land from chemical fertilisers and organic farm wastes take
account of nutrients already available in the soil and are consistent with recommended application
rates, crop requirement and the need to avoid water pollution.These powers provide the opportunity to
address pollution to high status waters, which could be identified as such in local authority
development plans.
40
4.2 Planning and Development Act 2010
The Planning and Development (Amendment) Act 2010 amends the Planning Acts of 2000 – 2009. It
is particularly concerned with sustainable development. This includes a closer alignment of the
National Spatial Strategy with Regional Planning Guidelines, Development Plans and Local Area
Plans, and clarifies obligations required of Planning Authorities under the Birds and Habitats
Directives, including the need for Appropriate Assessments and more focus on the need for EIA for
potential impact on Natura 2000 sites. The new Act provides greater responsibility on the local
planning or regional authority to protect sites designated under the Habitats and Birds Directives. Of
particular relevance is the strengthening of the protection of European priority natural habitat types or
priority species. It, furthermore, better integrates WFD legislation and RBMPs into Development
Plans, which must now include objectives for compliance of land-use with the relevant River Basin
Management Plans. In particular the Act provides for “promotion of compliance with WFD
environmental standards”.
4.3. Unsewered properties
The Court of Justice ruling on 29th October 2009 against Ireland (C-188/08), highlighted “serious
shortcomings” in the way septic tanks and other private waste water treatment systems are installed
and maintained throughout the countryside, stating that Ireland had failed to fully transpose the EU
Waste Directive 75/442/EEC into Irish legislation. Inadequate management of septic tanks in Ireland
includes incorrect construction, unsuitable location, insufficient capacities, maintenance and
inspection, and insufficient enforcement by local authorities. The court expressed concern that there
was no legal requirement for the planning authorities to use the updated codes of practice (EPA,
2009b) and that the standards in Building Control Standard S.R.6 of 1991 (referred to in Technical
Guidance Document H) are not suited to the geological and soil characteristics generally found in
Ireland. The ECJ found that: 1) apart from
by-laws in County Cavan, Irish legislation does not
transpose Articles 4 and 8 of the waste directive in so far as domestic waste waters from such on-site
treatment systems are concerned; there was insufficient provision for domestic waste water from onsite systems to be recovered or disposed of without endangering human health and without using
processes that could harm the environment; a failure to provide for the prohibition of uncontrolled
disposal of such waste waters; and inadequate provision for the handling of waste by a public or
private waste collector, recovery or disposal in accordance with the provisions of the directive.
Large parts of some counties in Ireland are unlikely to provide sufficient assimilative capacity for
nutrient retention from domestic septic tanks (D.Daly, EPA, pers com.). Where this is the case, a
system of local, or domestic scale secondary treatment system could be used as an alternative, or in
addition, to a septic tank. The secondary treatment systems tend to be less effective than the septic
tanks for nutrient attenuation and removal (Gill et al., 2009). Bartley (2003) estimated that about 10%
of N applied to the surface under dairy farming reached the ground water, which Gill et al. (2009)
suggested equates to a maximum allowable density of one four-person household per 0.33 ha-1 using
a septic tank, assuming equivalence of the Nitrates Regulations limit of 170 kg N ha -1 and introduction
41
of the effluent below the root zone. The equivalent density for households using an “average” septic
tank system is 0.18 ha-1. Hence, using the latter, and lower figure, suggests an equivalence of about
5 households ha-1 provides a similar risk to groundwater N pollution as farming 1.5-2 dairy cows.
Obviously these are average figures that do not take much account of the details of local P
attenuation, which will vary depending on flow pathways and capacity for soil attenuation, but do
provide some guidance and comparison between acceptable dairy and human densities compatible
with the good chemical status, as assumed by the Nitrates Regulations. For compatibility with high
status water quality much lower densities would be recommended. Where soil attenuation capacities
are low, including on peatland soils, alternative domestic treatment systems may be effective. These
could include use of attenuation ponds and constructed wetlands, which can attenuate impacts on
freshwaters from a range of pressures include nutrient loss, storm overflows, hazardous waste, food
processing and sediment run-off (Green and Martin, 1996; Vrhovsek et al., 1996; Xue et al., 1999).
Some local authorities, such as Clare County Council are investigating effectiveness of willow
systems for small-medium scale treatment (M. Burke, Clare County Council, pers comm). Medium
scale treatments systems, using constructed wetlands dealing with agricultural run-off, have been in
use for some time for agricultural run-off in the Anne catchment in County Waterford (Harrington et
al., 2007). So far, these systems have not been well tested in Ireland, but are used in other countries
to good effect (Craggs et al., 2004). Their use for water treatment, and in combination with production
of willow biofuel is worthy of further development (Guidi et al., 2008) and they can also mitigate
effects of storm water run-off from roads and their construction. Their effectiveness for wastewater
treatment depends either on attenuation and immobilisation of pollutant transport through settlement,
biotic uptake or, for nitrogen, transformation to nitrous oxide (a greenhouse gas) or nitrogen gas. Long
term success of attenuation of P in such systems may be questionable, as assimilative capacity
becomes used up. Nevertheless, constructed wetlands are increasingly used in farm systems. A
manual on their design has been produced in collaboration with the Northern Ireland Environment
Agency and the Scottish Environment Protection Agency (Harrington Carty et al., 2008). Irish
guidance
is
available
at
http://www.environ.ie/en/Publications/Environment/Water/FileDownLoad,24931,en.pdf.Local
Authorities have the power to introduce bye laws to manage septic tank location and management,
which could be tailored to suit local soil conditions, as done in Cavan. Zoned planning through the
County Development Plans can also take the cumulative impacts of one-off housing and small
developments into account. In this way, ‘risk’ zones for high status sites can be integrated into
County Development Plans. Currently, there is uncertainty among Irish Planning Authorities of how to
calculate or estimate cumulative impacts of developments. For sites designated under the Habitats
Directive, this relates specifically to Appropriate Assessment, although further guidance on this is
required. A similar process is required for high status sites not falling within the SAC network, but
protected through the Planning and Development (Amendment) Act 2010. The Department of
Regional Development in Northern Ireland limited the number of houses than can be built in rural area
through Planning Policy Statement 14 on Sustainable Development in the Countryside 2006 (now
superseded by policy statement 22).
42
4.4. Wind farms in upland areas
As Ireland meets the challenge for reliance on external fuel supplies and the diminishing use of peat
fired power stations, alternative energy is seen increasingly as part of the solution. The target for
Ireland under Directive 2009/29/EC on the promotion of the use of energy from renewable sources is
for Ireland to obtain 16% of its energy from renewable sources by 2016. This has led in the last five
years to a proliferation of wind farms in upland areas. As of the end of 2009, there were 120 wind
farms, and 1032 turbines across 19 counties in the Republic of Ireland (Ireland Wind Energy
Association, 2009). While wind farms form a legitimate component of a diversified source of energy,
they can have negative impacts on landscape and biota, and care with location is required (US Fish
and Wildlife Service, 2011). Among possible negative impacts are those on water quality, resulting
from mobilisation of sediment and altered hydrology. Planning for wind farms, therefore, needs to
include EIAs that address potential impacts on water quality and, in the context of this review, high
status sites.
Key points: Planning
The New Planning and Development Act 2010 provides for better intergration of high status sites into
the planning process. Recent rulings means that planning needs to take greater account of loss of
small wetland areas, and use of EIA. Unsewered propertires have also been under the scrutiny of the
ECJ and better planning, management and inspection is required. There is a need for devopment of
alternative small scale treatment systems, and testing of these, to reduce nutrient emissions to water.
Identification of high status sites in spatial planning, risk assessment of impact, and better
consultative procedures are required. This should include incorporation of spatial data of inter alia
EPA, NPWS and DAFF GIS layers to ensure it is available for planning authorities, forest inspectors
and others involved in environmental impact decision making so that they can take sensitive habitats
into account.
Spatial mapping needs to be at a sufficiently high resolution to include first-order
streams and small ponds/ lakes, and ‘risk’ zones for high status sites should be integrated into County
Development Plans. GIS planning tools could be developed as a special layer including high status
sites/ catchments for planning to include buffer or no-development zones in sensitive catchments.
4.5 Common Agricultural Policy
The Common Agriculture Policy (CAP) has been a cornerstone of European policy since the Treaty of
Rome (1956), and accounts for over 40% of the budget of the European Union. On average about
€250-300 per hectare of EU funds are used to support farming, and this figure does not count
member states’ co-financing and top-up expenditures (Zahrnt, 2011). The first pillar of the CAP, the
basis of direct farm subsidies, are financed entirely by the EU.
The second pillar involves co-
financing by the member states. This accounts for about 25% of the CAP budget, covering a range of
objectives, including environmental ones.
43
There have been number of reforms of the CAP towards greater incorporation of environmental
concerns since the early 1990s. This reflects the need to mitigate the environmental damage that was
driven by the CAP (see Harvey (1997) for a critique of impacts from agriculture in the U.K.).
Decoupling of production from subsidies since 2005 removed a major driver of environmental impact,
and payments now require cross compliance with other legislation, including environmental. Whether
or not this occurs in practice is debatable. Less than 10% of the CAP budget is spent on organic or
agri-environment schemes, although high value, low intensity farming focussed on local or regional
produce has been encouraged in the U.K. (Cabinet Office, 2002), and furthered with the concept of
countryside stewardship. The U.K. Environmental Stewardship, managed through Natural England, is
an agri-environment scheme that provides funding to deliver effective environmental management,
and is open to all farmers. See (www.naturalengland.org.uk/ourwork/farming/funding/es/default.aspx)
and also Section 5.2, Case study 4.
The last round of CAP reform led to a reduction of payments to the larger farms for reinvestment into
rural development programmes. This amounts to 5% of the budget for the period 2007-2012. This
money could very easily be targeted in such a way to help protect high status water bodies, as well as
promoting rural development. The CAP is due for further reform in 2013, which is generally thought
will lead to further “greening” of the policy. This can provide for enhanced policies to support
environmental quality and protection of biodiversity, in line with European policies on protection of
biodiversity in general (European Commission, 1998; 2011), and enrichment of biodiversity in
agriculture in particular (European Commission, 2001; CEC, 2002). The opportunities to use this to
strengthen protection of high status water bodies, and conservation objectives in general, are
obvious. Protection of high status water bodies can be done to complement protection of low intensity
agriculture, which has a fundamental importance for European conservation, and which is associated
with most of Europe’s valued biotypes (Bignal and McCracken, 1996, and see below case study on
the Burren, Section 5.2).
The CAP reform provides a very important opportunity for aligning
agriculture objectives with habitat protection, and can capitalise on the availability of funds that can, if
there is sufficient political will, support delivery of WFD objectives through the CAP. It is, however,
fundamental to recognise the risk of a “greenwash” by vested interests under a superficiality of
environmental protection. While such an observation in some quarters may be considered as a little
extreme, the history of engagement by the farm lobby and DAFF provide ample evidence for such
pessimism. While there now appears a greater commitment to environmental protection within some
Sections of DAFM, the need to operate general schemes with low administrative demands makes
focussed planning for protection of specific sites (outside the cSAC network) problematic. There is an
important need for a serious engagement between DAFM and the EPA to discuss strategies that can
use the CAP, including funding for rural development, to protect high status sites. This is a matter of
politics rather than science.
The current implementation of the first pillar direct payments have continued to promote
environmental damage, as only areas which are suitable for agricultural purposes appear to qualify
for payment. This leads to increased pressure to reclaim marginal habitat, including wetlands. Those
44
who hold grazing rights for commonage areas are also entitled to receive CAP payments proportional
to the proportion of commonage which are attached to their property folio and if the commonage
areas are suitable for agricultural purpose. If they are overgrown or not in suitable condition to be
grazed or farmed (e.g. if overgrown by gorse), payments can be withdrawn. In response to this,
grazing rights holders sometimes burn areas (including within protected areas) or remove scrub or
other vegetation to ensure that their payments are not withheld (Enda Mullin, NPWS, pers com.).
Commonage farmers were ideally placed to have availed of the rural environmental protection
scheme (REPS), which was Ireland’s response to EU Regulation 2078/92 to promote agrienvironmental farming. For uplands it was a mechanism to reduce environmental degradation
through, mainly, overgrazing by sheep which also led to pollution of water bodies through sediment
and nutrient loss (Bleasdale et al. 1995; Bleasdale, 1998; Huang, 2002); itself a consequence of
European subsidy for less advantaged areas under the EEC Livestock Headage payment scheme
under EEC Directive 268/75.
A study in 2004 (Van Rensburg et al., 2009a) indicated that among
commonage farmers in Galway and Mayo, REPS had a positive effect on income and associated
enhanced awareness of environmental issues and a lower use of inorganic fertilisers. Education and
adoption of more environmentally benign farming practice is crucial for the success, and policy
justification, of agri-environmental schemes (Morris and Potter, 1995). Van Rensburg et al. (2009a)
also indicated a younger age profile among, with more farming of cattle than sheep, and less reliance
on other state income among farmers in REPS compared with those outside the scheme.
Environmental schemes to enhance biodiversity in Ireland, under REPS generally failed, and have
been criticised by the OECD (2010). There continues to be a loss of biodiversity associated with
farming, in Ireland and across Europe (Donald et al., 2006; The Pan-European Common Bird
Monitoring Scheme, 2011). One of the main problems with agri-environmental schemes has been the
lack of data to assess their effectiveness. Where data has been collected in the Republic of Ireland
(Feehan, 2002), it has shown little evidence of biodiversity enhancement. It is not clear why there was
such a paucity of data collected to assess REPS, but its collection is widely thought to have been
resisted by the farm lobbies and DAFF, because the REPs funds were seen by these organisations
as a mechanism for enhancing farm income rather than primarily for environmental enhancement.
REPS was also not targeted to achieve a high level of site-specific management goals (BurrenLIFE,
2010). As REPs participation required farm nutrient management plans, created by REPS advisors
registered by DAFF, it is often assumed that REPs has played a role in reducing nutrient emissions
from farmyards, although data to support this is also lacking. Low intensity of monitoring of the EU
agri-environmental schemes is a common feature across Europe, although there has been
considerable variation in the approaches taken by Member States (Finn et al., 2009). In an impact
assessment of rural development programmes (EC, DG Agriculture, 2004), it was concluded that
water quality was likely improved, but impacts not measured. In an analysis of all the known studies,
that examined the effectiveness of agri-environment schemes up to the early parts of the last decade,
Kleijn and Sutherland (2003) found that despite €24.3 billion invested in agri-environment schemes
(although not all schemes were for the direct benefit of biodiversity), only 62 studies across Europe
(from five EU countries plus Switzerland) examined effectiveness for biodiversity. Most were in the UK
45
and The Netherlands, and many were scientifically weak; only 58% used controls, replication and
statistics. Most also showed limited or mixed benefits for conservation, concurring with conclusions
drawn by Feehan (2002) in relation to the Irish REPS scheme. However, in a review of U.K.
Environmentally Sensitive Areas (ESA) and Countryside Stewardship Schemes (CSS), Ecoscope
Consulting (2003) concluded mixed beneficial effects for wildlife, but, overall, both ESA and CSS
provided good value for public expenditure on landscape, habitats and biodiversity. In a more recent
study, Finn et al (2009) used multi-criteria analysis and expert judgement to assess European agrienvironmental schemes for their environmental effectiveness. Many schemes were considered to
suffer from poor data collection, and lack of baseline information. The performance of an individual
farm needs to be distinguished from the success of the scheme as a whole. Key points made by Finn
et al. (2009) provide a number of recommendations for successful schemes relating to better
definition, goals and monitoring. In particular: 1. the agri-environmental schemes need to demonstrate
the additional value over and above normal environmental compliance; 2. there is a need for a more
direct connection between financial and environmental information for cost-benefit analysis; 3. poorly
designed schemes can lead to poor environmental performance that may take considerable time, and
perhaps resources, to correct; and 4. there is an explicit need for cross-compliance, which is a
requirement for EU subsidised schemes regardless of the targeted sector.
The European
Commission provides a handbook for the monitoring of all rural development interventions (DGARD,
2006). E.U. taxpayers are increasingly likely to expect a greater environmental return from the agrienvironment budget (Finn et al., 2009). The success of agri-environmental schemes is dependent on
the scale and fit-for-purpose of actions. More general prescriptions are likely to be less effect than
locally targeted ones. Richardson et al. (2008) suggests positive impacts from several federally
funded conservation programmes initiated under the U.S. Conservation Effects Project (CEAP) in
2003, that include 38 agricultural dominated catchments, including some with water quality and flow
data going back to the 1960s. Marshall
communities following
et al. (2008) demonstrated positive effects on fish
introduction of conservation programmes that included a combination of
extensive grassland management, livestock reductions, and other long-term agricultural land use
changes, although this may have taken up to three decades to be effective, as data was only
compared from the 1970s prior to the programmes with the early 2000s. Targeted local or regional
agri-environmental schemes in Ireland and elsewhere are discussed further in Sections 5 and 6.
The REPs schemes are now replaced by the new agri-environment options scheme (AEOS) under
pillar 2 of the CAP, which relates to additional payments for rural development and the environment
and is subject to different administrative rules than the single farm payment under pillar 1. The only
water - protection measure in the AEOS is the limiting of animal access to water courses, and
provision of drinking troughs. The AEOS funding (DAFF, 2011) is also prioritised for Natura habitat
and/or Non-Natura Commonage. If the number of applicants in this category exceeds the total funding
available those with designated land will be given priority followed by those with the largest
commonage areas. Further selection of applicants will be applied, if necessary, on the basis of: a)
farms in Less Favoured Areas; b) Previous participation in REPS; and c) Farm size (favouring smaller
holdings) based on the utilisable agricultural area declared in the 2010 Single Payment Scheme
46
application. The AEOS scheme reflects, therefore, two important factors relevant to farming that may
affect high status sites. First, there is a low priority for fiscal support for farms outside Natura 2000
sites, meaning that most high status sites are unlikely to benefit from the AEOS unless there is a
generally low uptake of the scheme. Second, and more importantly, it reflects the failure of the cSAC
management network to protect habitats because of insufficient direct investment into the
management of these sites. In the U.K. site protection of the Special Site of Scientific Importance
(SSSI) is supported through fiscal compensation related to estimated income loss that results from
restrictions on maximum productivity. The prioritisation of Natura 2000 sites has, if it leads to a lower
provision of agri-environment schemes outside the Natura 2000 network, the potential to increase
pressure on high status sites outside the Natura 2000 network.
Further CAP reform has the opportunity, and certain political pressure, to use more money of the
pillar 1 for enhancing environmental protection. It may also help reduce the widespread contradiction
and confusion in agricultural and environmental policy (Hendry et al., 2006; Newson, 2010). There are
strong arguments to use the CAP more for rural development and public goods rather than for direct
agricultural support. This can only be achieved, not least because of the principles of subsidiarity,
where action is taken at the national rather than EU level, through national programmes and
agreements. DAFM, therefore, has a key role in supporting the protection of high status sites and
other areas of biodiversity importance.
Their position on the Nitrates Regulations (2010), and
promotion of Food Harvest 2020 (DAFF, 2010) as the future policy for Irish agriculture is not
encouraging for environmental protection. A shift in the traditional mindset of the Department, and it
advisory service Teagasc, is required so as to align EU environmental objectives with national
agricultural policy. The objectives of Food Harvest 2020 include a 50% increase in dairy production.
This can only increase the pressure on aquatic resources across the country, with likely further impact
on high status sites. The failure of DAFF to conduct a Strategic Environmental Assessment prior to
the adoption of this policy appears to be in breach of the Strategic Environmental Directive
(2001/41/EC), although DAFF have made a commitment for each section of the policy to undergo
SEA prior to adoption. That it is acceptable for an SEA to occur in parallel to policy implementation
relies on the argument that plans not foreseen in prior legislation are formally outside the scope of the
definition of "plans and programmes". This would appear not in keeping with the spirit of both the
SEA Directive and the WFD. There is, nevertheless, the requirement for appropriate assessment
under Article 6(3) of the Habitats Directive 92/43/EEC (habitats). Ireland has not yet transposed this
requirement for plans (ECJ case C-418/04).
So far, there is little evidence that the application of agri-environment schemes funded through the
CAP has led to any significant protection of water resources in Ireland. A lack of monitoring restricted
this evaluation. Bignal (1995) observed that “despite the growing awareness and recognition of lowintensity farmland, policy makers continue to have fundamental problems in devising schemes which
have real ecological value”. This appears still largely the case. From 2013 it is likely that there will be
a reduction in the overall CAP budget. This provides another opportunity for more meaningful use of
agri-environmental schemes targeted to sustainable land use, to support the protection of high status
47
sites, the objectives of the WFD and integration with wider countryside conservation. The need to
provide for better links between EU agriculture and the objectives of the WFD is summarised by
Dworak et al (2005), in a report endorsed by the EU Water Directors. Key points that can be taken
from that report with respect to the protection high status waters in Ireland are:

Measures under pillar 2 of the CAP offer high potential to support implementation of the WFD,
but require dialogue with DAFM in order to set priorities;

A need to target areas for support, accepting that the expense to restore some areas is
disproportionate to the cost. Preventing deterioration of high quality sites is almost certainly a
more cost effective strategy than large scale restoration of grossly impacted sites. The WFD
Article 5 provides important information with which to prioritise sites; and

Owing to high competition for accessing Rural Development budgets, a strong argument is
needed if some of those funds are to be directed towards long term protection of high status
water bodies.
There are a range of existing mechanisms through CAP subsidies through less favoured areas
grants, agri-environment schemes and through the LEADER programme (Rural Development Axis
IV), which can provide a bottom up approach to WFD implementation and promotion of specific
targets (ÖIR-Managementdienste GmbH, 2004). This could apply to high status waters, with well
defined regional scope and actions at the farm level. It links well with developments and potential
under the European Landscape Convention discussed in Section 1.2.2 and provides a firm logic for
greater integration of opportunities to protect high status sites.
Any action as outlined above requires good farm advisory support, and a willingness of rural
stakeholders to participate, which is helped by clear communication of the benefits. It requires a
compensation for activities that go beyond the current view of good agricultural practice, even when
this is accompanied by high awareness of the environmental benefits. Learning from the REPS
experience, it requires effective monitoring, with feed-back to stakeholders. A more radical reform
would be using the CAP pillar 1 for environmental protection, the support of public goods and cross
compliance with Environmental Directives. There already exists precedent for this in that the Single
Farm payment is eligible for conversion of up to 10% of the eligible land to forest under the Native
Woodland Scheme (see below Section 5.2). That there are political moves across some EU Member
States to move in this direction is borne out by recent (August 2011) comments by the U.K. Secretary
of State for Environment, Food and Rural affairs, Caroline Spelman, made at the launch of the UK's
biodiversity strategy that the (U.K.) government was committed to rewarding farmers for good
practice, and would push within Europe for more emphasis on conservation in the reformed CAP.
48
There is a greater expectation across Europe of community-wide benefits from the CAP subsidies,
and promotion of demonstrable countryside stewardship. Further opportunities arise from linking
these initiatives with large scale regional projects supported by e.g EU INTERREG or LIFE funding.
There is potential to develop such initiates into regional “win-win” networks of wetland enhancement,
nutrient and flood attenuation, and rural development. The BurrenLIFE project and wetlands
development in the Anne valley in Waterford (Harrington et al., 2007) provide a useful prototype, from
which E.U. funding programmes such as LIFE can help develop into wider regional spheres (P.
Jordan, pers comm., University of Ulster). For this to be realised, however, requires a fairly radical
shift in approach, away from using subsidies as a farm income support per se, to one where farming
delivers tangible environmental goods and services.
Key points: Reform of the CAP
The reform of the CAP in 2013 provides a major opportunity to link better the objectives of farming
and rural development with those of the WFD. There needs to be immediate engagement between
EPA and DAFM in influencing these decisions and making a strong argument to align rural
development and farming objectives to support the protection of sensitive environment, including high
status sites indentified under the WFD.
4.6 Forestry
1. Current Government policy is to increase forest cover by 10,000 ha per year, continuing an
average, albeit inconsistent, afforestation rate since the mid 1980s, and one which is almost entirely
driven by government grant aid (Malone, 2008). Commercial forestry has the potential to impact on
water quality, especially when planted in upland areas, which are often prone to soil erosion, and
have a low capacity to buffer against acidification (Allott et al., 1997; Kelly-Quinn et al., 1997).
Johnson et al (2008) provide a comprehensive review of risks of acidification from forestry, and
Hutton et al. (2008) a review on impacts from sediment and nutrients; both reports commissioned by
the Western River Basin District. Impact on surface waters from sediment and nutrients can be
especially severe during planting and, even more so, during harvest. Impact from nutrients added to
forestry tend to be localised, but may also impact high status and other sites downstream. Giller and
O’Halleran (2004) considered that “although forest operations do input excess nutrients causing local
effects, it is unlikely that the resulting changes are comparable to the extensive damage caused by
intensive agriculture”. On the other hand high quantities of sediment loads can arise from catchment
disturbance associated with forestry (Robinson and Blyth, 1982; Everest et al., 1987; Scoles et al.,
1996; Swank et al., 2001). Associated with forest clearance can also be high concentrations of
solutes, as classically demonstrated in the Hubbard Brook catchment in the 1960s (Likens et al.,
1971).
49
2. In the Burrishoole catchment (Co Mayo), concentrations of mean total available dissolved
phosphorus increased from about 6 µg l−1 during pre-clearfelling to 429 µg l−1 afterwards. It took four
years for the P concentrations to return to pre-clearfelling levels, despite the adoption of good practice
and use of up to 20 m buffer strips (Rodgers et al., 2010). However, the same study-team found no
long-term impact on the suspended solid concentrations (Rodgers et al., 2010), indicating the
possibilities to mitigate sediment transport using good forestry guidelines. That these did not work well
for the dissolved phosphorus indicated different transport and mobilisation mechanisms involved, and
the mobilisation after forest clearance of soluble P. This is informative as it may be thought that tree
clearance would be more associated with mobilisation of particulate P, as occurs from many arable
systems (Schuman et al., 1973; Truman et al., 1993) or where the soil is exposed (Gillingham and
Thorrold, 2000).
Dissolved fractions are more usually associated with phosphorus export from
grasslands (Sharpley and Syers, 1979; Lennox et al., 1997; Heathwaite and Dils, 2000; Nash and
Halliwell, 2000).
3. Many Irish forests that are now mature, or approaching maturity, were planted in landscapes that
were unsuited to economically viable forest production. The increasing recognition of the impacts from
forestry on water resources has led to the development of a Code of Practice for forestry (Forest
Service, 2000a b; Purser Tarleton Russell Ltd., 2000). Nutrient enrichment from aerial fertilisation has
been curtailed to some extent through requirements for an aerial fertilisation licence from the Forest
Service under S.I. No 592 of 2006 and S.I. No 790 of 2007, European Communities (Aerial
Fertilisation) (Forestry) (Amendment) Regulations 2007. Generally forest management is based on
the Code of Practice, although a new Forestry Bill, that will replace the very out of date Forestry Act
1946, has been drafted with the opportunity to ensure that forestry management is better able to
protect sensitive habitats.
4. The Code of practice identifies well activities that pose a high risk to water quality and an extensive
range of controls are applied by the Forest Service to forest management including site-specific
conditions attached to consents, grant approvals, afforestation licenses, prohibiting new forestry on
unimproved or unenclosed land and increased use of sub-threshold EIA. Despite this, forestry in
some areas can impact surface waters from sediment, nutrients or other pollutants (unpublished data,
HYDROFOR project, www.ucd.ie/hydrofor/home.htm). The new and better awareness within the
Forest Service is, nevertheless, constrained by insufficient detail of the thresholds for activities
commensurate with good status under the WFD. Thresholds of activities to prevent impact on high
status water bodies are not developed, and may necessitate severe curtailment of current activities
and prevention of new planting because of inherent difficulties operating commercial forestry in areas
that may impact on high status waters (Eriksson et al., 2011). Many forests which are now mature, but
not conducive to commercial felling, nevertheless have the potential to impact water resources even if
left alone because of extensive windfall as the stand ages (T. McDonald, Forest Service, pers
comm.). The management of forestry operations is also addressed in the RBMPs but, as with other
aspects of the WFD, the focus is on attainment of good status.
50
Harvesting of existing sites, where
there is a risk to impact high status waters, therefore, requires particular care, and likely development
of new techniques.
5. Tailored forestry guidelines were prepared to prevent deterioration of high status water quality in
freshwater pearl mussel catchments in Ireland. These guidelines represent the most comprehensive
strategies developed for this purpose to-date and could be extended to mitigate impact on other high
status sites. Development of forestry practices such as Continuous Cover Forestry (CCF) can reduce
impact from felling. The UK Environment Agency (2006) compiled Best Practice Guidelines in 2006,
which detail several CCF case studies in the UK, with examples from elsewhere in Europe (Slovenia,
Switzerland, Austria, Bavaria, Czech Republic). Organisations such as ProSilva Ireland advocate
alternative forestry methods (Diaci, 2006), and some private forestry operations in Ireland (including
areas of Coillte forest) have adopted low-impact alternatives to conventional methods, with similar
plans prepared for certain NPWS-owned forested areas, e.g. Freshwater Pearl Mussel catchment in
Glengariff Nature Reserve (Purser et al., 2011).
Research into the value of CCF in minimising
impacts on water quality is encouraging (Reynolds, 2004). With plans to increase commercial forestry
yields in Ireland over the coming years, there is a need to further develop and test alternative forestry
techniques under a range of soil conditions.
6. For new forestry applications, it is important that the location of high status sites downstream of the
proposed area are identified. This can be done through the Forest Inventory and Planning System
(FIPS), and its successor the GIS supported iFORIS, designed to incorporate forest and site
categories with environmental information, felling control and grants administration. iFORIS is a
powerful multi-layered GIS that supports forest management. It contains water bodies as a layer, that
can be expanded to include status classes. While there is an extensive range of information contained
in iFORIS, in general, there is a need for greater common use of GIS and information sharing across
all sectors of Government and their agencies (DECLG, 2011). The new INSPIRE Directive
2007/2/EC, establishing an Infrastructure for Spatial Information in the European Community, will
provide further motivation to address this disparity. The inclusion of RBMP GIS is clearly an urgent
requirement. This also applies to the identification of the ‘aquatic zone’ used in forest management. In
the forestry and water guidelines, the forestry and biodiversity guidelines, and other Codes of
Practice, these are shown on a 6“ Ordnance Survey map. Many water courses are not visible on the
old 6” maps which have not been revised in over 50 years, and nearly 100 years in some cases.
Ignoring the protection of water courses within forest activity areas can lead to severe damage.
Water-dependent habitats such as bogs or headwaters are unlikely to be delineated clearly on old 6”
maps, and updated information from NPWS, EPA and RBMP data should be integrated into the
iFORIS system. These layers are also needed by local authorities so that effective decisions can be
made on planning applications that may impact on high status sites. While the thresholds until
recently have been below that where an EIA is mandatory, the recent ruling by the ECJ (see above,
Section 3.7) and new regulations, updating the European Communities (Environmental Impact
Assessment) (Amendment) Regulations, 2001 (S.I. No. 538 of 2001), that provide for statutory EIA
51
screening for sub-threshold forestry, provides for more attention to potential impact from small areas
of forestry.
7. Additionally, on-site survey prior to granting an application should identify sites with potential high
quality aquatic communities, either through the EIA or grant aid process. Currently, management
plans are mandatory for grant-aided plantations exceeding 10 ha and are only recommended for
smaller plantations. Furthermore, a condition of grant aid is the replanting of felled forests. This is not
necessarily in the best interests of aquatic sites that are vulnerable to pressures from forestry. There
are recent examples of inappropriate replanting on SAC catchment sites likely to impact high status
waters in the absence of the level of investigation required to fully satisfy Article 6 of the Habitat’s
Directive. At the moment the Forestry service rely on using a) their risk assessments and b)
consultation with NPWS and fisheries for licensing. Their risk assessments are based on mapping of
soil type and soil suitability. There have been instances (Bundorragha Catchment) where a license
for felling and replanting, based on risk assessment and consultation with local NPWS staff but in the
absence of a convincing Appropriate Assessment of the risk, and where the Forestry Service are
unable to control subsequent application of fertiliser, as long as it is not aerial. With respect to
rehabilitation of upper catchments, the Forest Service have no role in drainage management. A key
area of difficulty with forestry and high status sites, particularly those that need restoration of function,
is that the Forest Service grants afforestation / replanting and felling licenses without considering
these associated activities. Most viable upland forestry needs draining and fertilisation with
phosphorus. Hence, the consequences of licensing are not assessed by the Forest Service and these
sites are unlikely to be commercially viable without associated drainage, fertilisation and access
through forest roads. All of these pose risks if in the vicinity of high status sites.
8. Where replanting does occur, planting of fast-growing shrubs after clear felling can reduce nutrient
losses from soils, but under current grant-aid conditions, replanting of alder trees are covered by
grant-aid whereas trees such as mountain ash or a shrub such as Myrica gale, are not (Thomas
Cummins, pers. com, UCD soil science). Obligatory replanting following felling is, however, likely to
be repealed in the new Forestry Act.
9. Overall, there is a need for greater attention to potential impact on high status water bodies from
forestry, which could be further supported with bye laws to control forestry operations, including
cumulative impacts, and further development of a Forestry Code of Practice more closely aligned with
the objectives of the WFD. While Appropriate Assessment under the Habitats Directive can help to
control potential impact to cSACs, current assessment levels need to be radically improved with
regard to their acceptance under Article 6 of the Habitat’s Directive, including the assessment of the
level of rehabilitation needed to allow a catchment to function sustainably. It would be beneficial if this
were extended to potential impact to high status sites that lie outside the cSAC network. This type of
strategy is adopted under the US Clean Water Act 1972 to protect Outstanding Resource Waters.
This would lead to considering statutory mitigation measures to address the effects of forestry in all
catchments of special sensitivity, including those containing high status sites. This may involve a
banning of new forest plantation in such sensitive catchments, as is currently done in Ireland for areas
52
with high sensitivity to acidification. For existing catchments with forestry that may impact high status
sites, tighter controls may be needed on clear-felling, independent EIAs, and strict controls of coup
sizes harvested. The current guidelines (Forest Service, 2000a) provide for a “tactical” harvest plan
when clear-felling is not appropriate, identification of environmental issues, and consultation with the
relevant authorities. This should be a minimum requirement when harvesting might impact a high
status site. Broadly similar guidelines for forestry management occur in other European countries
(e.g. Swedish Forest Agency, 2000; Forestry Commission, 2003). The Forestry and Water Quality
Guidelines (Forest Service 2000a) for Ireland are due to be reviewed during the first cycle of the
RBMPs so as to reflect the new water quality objectives and standards. It is anticipated that there will
be an 18 month review period (K.Collins, Forestry Service, pers com.), so there is opportunity for
extensive and focussed compilation of information and international experience and best-practice.
The New Forestry Act is likely to precede that review, so it is important that provision is made to
anticipate further revision based on the findings of the review.
10. There are additional opportunities for incorporating WFD and forestry objectives. The Native
woodland scheme provides grants to support conservation of both existing semi-natural woodland
and the establishment of new areas of woodland on greenfield sites (DAFF, 2008; 2011). Element 2
of the scheme is targeted towards sites “within areas regarded as being particularly sensitive from an
environmental, landscape or amenity perspective, sites located immediately adjacent or close to
existing designated native woodland sites that create physical connectivity between existing native
woodlands and other important habitats”. This could be used to promote native woodland buffer
zones to high status waters. The Forestry-environment payments under the CAP are compliant with
EU Regulation (EC) N° 1698/2005, and can be used to support rural development, including wetlands
(Dworak et al., 2009). Under the Irish Native Woodland scheme, riparian wet woodlands may not be
eligible for grant aid owing to the criteria for “vigorous growth and sustainable long term
development”. This appears an anomaly, inconsistent with the Forest Service grant aid for water
margins, and highlights again the notion that agricultural schemes that support natural habits must
always be geared towards some measure of optional agriculture. Riparian wet woodlands have a
potential role in attenuating sediment and nutrient transfer to water bodies which can also be
beneficial for mitigating effects of flooding (Williams et al., in press). These types of initiatives can help
realise enhanced landscape connectivity of the waterscape discussed in Section 1.2.Current research
through the HYDROFOR project (www.ucd.ie/hydrofor/home.htm) supported by Ireland’s Department
of Agriculture, Fisheries and Food through COFORD and the EPA is investigating utilisation and
importance of buffer strips to mitigate environmental damage on aquatic sites. Forestry can improve
nutrient storage in soils and attenuate nutrient mobility though catchments (Schoumans et al., 2010).
Key points: Forestry
Forestry can impact water resources and many Irish forests that are now mature, or approaching
maturity, were planted in landscapes that were unsuited to economically viable forest production, with
a potential to cause impact through harvesting. The current Forestry Code of Practice, the new,,and
53
pending, Forestry Act, and greater integration with the goals of the WFD suggests greater
consideration to potential impact to water resources in the future. There are also possibilities for more
positive impacts of forestry through restructuring of grants to promote environmentally sensitive
forestry. As with agriculture in general, there is a need for better liaison with the EPA and local
authorities and more comprehensive use of shared GIS to guide planning. However, the current
approach is focussed on compatibility with achieving good status of water bodies, and more
innovative approaches are needed to support and promote protection of high status sites, many of
which are in upland areas and prone to impact from low to moderate disturbance. Consideration is
needed to the banning of new plantations that may impact high status water bodies and, for maturing
forests, harvesting limited sized coups, with strict adherence to best practice guidelines. There are
opportunities for enhancing protection of high status sites through promotion and fiscal support for
riparian buffer strips, utilising existing mechanisms allowed through the CAP, and integrating new
forest areas within the landscape with the goal to attenuate nutrient mobility.
Section 5. Stakeholder and Public Participation.
5.1. Introduction
The WFD provides for an integrated approach to catchment management that requires a broad
appreciation, though not necessarily wholehearted acceptance, or concensus across a range of views
(Clarke, 2002). It also requires a collaborative framework (Lee, 1973; Pahl-Wostl and Hare, 2004,
Watson, 2007). Environmental policy attempts to reflect current accepted understanding of pressures
and impacts, but the very nature of environmental policy is open to debate, scientific uncertainties,
and political and administrative complexities (O’ Riordan 2000; Kallis and Butler, 2001;Irvine et al.,
2002; Watson et al., 2009). While WFD implementation is guided by technical knowledge, to be
effective requires acceptance, and preferably support, by stakeholders and the wider public (Jasanoff
and Wynne, 1998). It, therefore, requires both a firm technical basis and societal support. However,
information used to guide environmental management frequently comes from different disciplines, is
difficult to interpret, and usually incomplete (Malafant and Fordham, 1998). The measurement of e.g
ecological condition, nutrient transport and effects of management also contain varying degrees of
uncertainty despite the application of precise techniques of measurement (Clark, 2002). The
appreciation of environmental uncertainty by all stakeholders is a key aspect in the implementation of
environmental policy, and unrealistic demands for certainty, can hinder environmental protection
(Westervelt, 2001). It is against such a background that stakeholder involvement occurs. Article 14 of
the WFD requires public consultation, which reflects the current paradigm of environmental
management, although the gap between policy requirements and stakeholder aspiration and
compliance provides particular challenges (Glicken, 2000; Lubell, 2004). Its implementation may be
one of the more difficult aspects of the Directive. Water management in Europe has tended to be
dominated by a centralised and technocratic approach. The WFD requirement for public participation
54
(Annex VII) includes the need for RBMPs to provide a “summary of the public information and
consultation measures taken, their results and the changes to the plan made as a consequence”.
The procedures for integrating information and experience across multiple disciplines and interest
groups are inherently challenging given that multiple stakeholders will often have conflicting interests
and disproportionate power (Gregory and Keeney, 1994; Freudenburg 1998; van der Arend and
Behagel, 2011).
Stakeholder engagement is also time consuming, and promoting landuse and other activities for
environmental protection is dependent on an array of common factors (Cuff 2001; Pahl-Wostl and
Hare 2004; Mostert et al. 2005, 2007; Steyaert and Jiggins 2007; Acland 2008). Mostert (2007),
reviewing 10 River Basin case studies across Europe, identified 71 factors fostering or hindering
social learning, in particular: clarity of
institutional
attitude to
stakeholders’ role in decision-making; the political and
multi-stakeholder
engagement; opportunity for
interaction
amongst
stakeholders; motivation and skills of leaders and facilitators; openness and transparency;
representation of stakeholders views; and resource availability. A lack of clarity regarding the role of
stakeholders was identified as the factor which posed the greatest impediment to success.
Acland (2008) provides a summary of guiding principles for public participation. These include:
inclusiveness; transparency, openness and clarity; commitment; accessibility; accountability;
responsiveness; willingness to learn; and productivity and reflect the ‘Core Values’ of the International
Public Participation Association (http://www.iap2.org) and the principles inherent in a Learning
Alliance (www.irc.nl/la). In a review of these principles applied to the River Basin Management public
consultation process in Ireland, and the wider governance of stakeholder participation through the
River Basin Advisory Councils, Irvine and O’Brien (2009) concluded that while there was a public
participatory process, there was no analysis of its effectiveness. This is a common phenomenon
(Rowe and Frewer, 2000), but paying attention to some simple principles can enhance the success of
public participation. Considering the requirement of the WFD, Özerol and Newig (2008) identify five
key factors affecting it success: 1.the scope of the participants; 2. communication; 3. Capacity
building; 4. timing; and 5.financing of participation.
The public consultation process for the RBMPs in Ireland focused on the production of the document
Water Matters. While produced for each RBD, it was generally a generic document overseen by the
then Department of Environment Heritage and Local Government (DEHLG). As a process, the
document and associated public meetings, many of which were poorly attended, fell short of the
recommendations of Acland (2008). Somewhat similar experiences of the RBMP process in the U.K.
are reported by Watson and Howe (2006). This matches many of the experiences of RBMP
consultation across Europe. Watson et al. (2009) report that “it quickly became clear the EA [UK
Environment Agency] staff involved had a very fixed idea of what they wanted to achieve and were
unwilling to allow the participatory process to evolve or deviate from the original aims”.
In the
Netherlands, which has a tradition of stakeholder engagement in water management dating back to
55
the 1200s, the public participatory process for the WFD RBMPs led to a powerful reaction from the
farming and business interest groups that led the Dutch Government to adopt a “pragmatic approach”
to implementation (van der Arend and Behagel, 2011). This resulted in a discourse that focused on
where proposed measures were ‘feasible and affordable, which restricted exploration of more
ambitions options that might be considered too costly or involve changes in land-use. One outcome of
this approach was to accentuate differing values across stakeholders and polarise the debate.
Centrally organised participatory activities then came into conflict with actual participatory practices,
leading to disengagement or informal participation by a number of stakeholders (van der Arend and
Behegal, 2011). The Dutch experience likely reflects a common position across many of the
European Member States where “participation hardly coincide with the official and academic accounts
that are mostly given of them”, and that the “organised participatory processes do not and cannot
stand alone outside the “old” democratic institutional settings, without risking undermining participants’
power base (van der Arend and Behegal, 2011).
The early experiences reported by Watson and Howe (2006), Irvine and O’Brien (2009) and van der
Arend and Behegal (2011) indicate some of the difficulties with national RBMP consultation. This
process was required by the WFD so risks either being done to be seen to be compliant or, and
associated with that, to lack credibility across the wider stakeholder community. The response to
consultees comments in Ireland suggests that the process was there to support, rather than
challenge, central policy (Irvine and O’Brien, 2009). It is, however, easy to be critical after the event
and there is nothing to suggest that those involved in administrating the consultation were in any way
deficient, given limited budgets and time. The RMBP consultation was also set up to introduce and
address a very wide range of issues at effectively a national scale, while most public interest and
engagement is at the local scale. The acute interest in waterbody protection in Ireland that occurs at
the local scale (e.g. Huxley and Irvine, 2008), requires fostering by both a local “bottom-up” process
and “top-down” support. It could also be argued that the Advisory Councils, created under European
Communities (Water Policy) Regulations (S.I. No. 722 of 2003, amended by S.I. No 413 of 2005)
were limited in scope and influence, lacking in a clarity of their roles and capacity to influence
decisions. Anecdotes of poor attendance by members, or members attending meeting for only a very
short time, despite substantial travel to get to meetings, are common. The fact that the Advisory
Councils have not met since their dissolution because of local elections in 2009, suggests a less than
satisfactory view of their role by central government.
The public participation process for the RBMPs, as set out in Annex VII of the WFD, was In Ireland,
and in many Member States, effectively a three stage process of 1) Government dissemination of
information, 2) inviting a response from stakeholders and 3) responding in a fairly generic manner
with a summary of stakeholder views. In Ireland, the U.K., and in other countries the formal WFD
consultation process often failed to generate a real sense of serious stakeholder engagement.
56
Public participation involves a number of techniques each with their own strengths and weaknesses,
so that a variety of engagement methods is advocated (Rowe and Frewer, 2000; Ridder et al., 2005).
A focus on specific issues, capitalizing on local knowledge, is likely more successful than national or
regional consultation. Very often this necessitates a progressive and sustained process of learning
and adaptive management beyond provision of information (Lee, 1973; Folke et al., 2003; Du Toit,
2005). Stakeholder engagement can usefully use Decision Support Tools, including mathematical
modeling (reviewed in Irvine et al., 2005) and scenario testing, that explore alternative management
options, to build consensus (Schauser et al., 2003). Multicriteria decision analysis (MCDA) can be
used to assess the response of stakeholders to various management options, and social and
economic choices (e.g. Lahdelma et al. 2000; Belton and Stewart, 2002; Park et al., 2004; Prato and
Herath, 2007; Schulte et al., 2009). Cost-Effective Analysis (CEA) and Cost-Benefit Analysis (CBA)
can be built into such decision support tools, but can suffer from a number of limitations that
undervalue social and environmental goods and services (Hobbs and Meier, 2000; RICS, 2001).
These “externalities”, which can play a key role in decision making, and are not easy to place a
monetary value on, are often estimated by techniques such as contingent valuation and tradeoff,
necessitating inherent value judgements (Joubert et al., 1996).
The limitations of CBA can be
addressed through MCDA (Gowdy and Erickson, 2005; Prato and Herath, 2007), which is designed to
accommodate complex and conflicting information and values (Belton and Stewart, 2002).
A number of market based instruments can be used to improve delivery of environmental goods and
services (Stavins, 2000), and can be used to support the WFD (Cockerill and Hutchinson, 2009). For
agricultural land use in Europe, which operates within a subsidized economy, market instruments are
likely better than fixed subsidies, which lack incentive for developing new and innovative approaches,
and have limited sustainability beyond the contract period (Latacz –Lohmann and Schilizzi, 2005). An
exploration of the feasibility for land-use auctions was explored for use in the Lough Melvin catchment
(see below, Section 5.2, Case Study 2). Typically, farmers submit bids to the Regulator, with those
offering most benefit accepted for contact. It also reveals true cost of participation and provides a
cost-efficient allocation of funding. Auctions have been shown to be an efficient means to alter land
use for environmental benefit (Latacz-Lohmann and van der Hamsvoort, 1997) and their use in the
U.S. Natural Resource Conservation Program have reduced land use pressures on water resources
(Huang et al., 1990; Young and Osborne, 1990). Trials of auctions in Australia has also reaped
environmental benefits (Stoneham et al., 2003), but at high administrative cost (Bryan et al., 2007).
An option for reducing nutrient run-off to sensitive catchments is nutrient trading (USEPA, 2007). This,
like C trading, can be used where low nutrient loads (or low intensity land use) can be used to provide
salable credits for more intensive land use. This could be targeted for high status sites, and used to
balance impact from intensive farming elsewhere in the county.
Although it may be difficult to
administrate, not least because of the uncertainty of causal links, there is also much uncertainty in
effectiveness of current C sequesation policies, which are an important element used to combat
global C emissions. A similar trade in nutrient emissions could operate with legally binding Emission
Limit Values, such as operated by the US TMDLs system (See Section 2.2).
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While, under the WFD, there are some technical deliverables that are needed irrespective of
stakeholder views, meaningful stakeholder engagement provides benefits. Stakeholder workshops
that capitalize on existing knowledge, and propose realistic outcomes, can mitigate conflict between
decision makers and end-users, and build confidence in workable solutions (Watson and Wadsworth,
1996; Hofmann and Mitchell, 1998; Huxley and Irvine, 2008). Sophisticated stakeholder participation
using integrated decision support tools are not well established for natural resource management in
Ireland, although there are some examples from the Burren and the Lough Melvin catchment which
have explored and developed this. Similar experiences of stakeholder participation for water resource
management have occurred in some river basins in the U.K. The next Section provides examples
from, respectively the Burren, Lough Melvin, and the U.K.
5.2 Case studies of public participation in Ireland and U.K.
Case Study 1.The Burren Life project (www.burrenlife.com).
The BurrenLife project (BurrenLife, 2010a) provides an example of a mechanism of how farming and
environmental protection can be supported through EU and Irish government funds, involving
collaboration among a number of state agencies and farmer organisations. In particular BurrenLIFE
fostered collaboration among stakeholders who represent quite different interests and often hold
conflicting views. The project, which ran from 2005-2010, aimed to develop a model for sustainable
agriculture which supports conservation through a range of practical farming measures, and provided
the basis of a new “Burren Farming for Conservation Programme” under REPS IV. Utilising CAP
funds,it extended the project into a second phase, with a target of 100 participating farmers. This is
reported to be “Ireland’s first evidence-based, area-specific agri-environmental scheme” (BurrenLIFE,
2010). It commenced in 2010, with up to €1 million available to eligible participants in each of the
years - 2010, 2011 and 2012.
The Burren, of distinct national and international conservation importance, has a species rich
landscape dependent on low-intensity grazing, but where traditional farming is of marginal economic
viability. Mainstream agricultural policies have contributed to a loss of high quality habitat, either
through scrub encroachment because of a lack of grazing through land abandonment, or because of
more intensive farming. The BurrenLIFE project concentrated on 20 farms, covering 3097 ha, of
which 80% was in cSACs. Farm size ranged from 40-448 ha and stocking rates from 0.19-0.81 LU
ha-1. EU LIFE funding of 75% of total costs, with the remainder coming from the NPWS, Teagasc and
the Burren Irish Farmers Association, enabled financial support for specific conservation criteria, and
without the general constraints of open-availability of financial support that is a usual feature of usual
DAFF support under pillar 2 of the CAP. Participating farms had a management agreement with the
project which, although having no legal basis, was sufficient to ensure farmer cooperation and
management focussed on the threats to priority habitats (BurrenLIFE, 2010). By providing additional
58
financial and technical support, BurrenLIFE enabled a more focussed conservation approach than
had been done previously, despite cSAC designation, or participation in REPS. Development of
improved knowledge and management relating to stocking regimes, supplementary feeds, and
economic viability of farming for conservation in the region should facilitate future sustainability of
conservation orientated farming in the region. The results so far indicate an improvement in grazing
regimes and widespread scrub clearance (BurrenLIFE, 2010). Although grassland monitoring station
were set up, there appears to be no published data from these yet. However, like nutrient
management, lag phases between action and conservation results are likely.
For the protection of wetland habitats the project developed a risk-assessment model for nutrient
export from agriculture to surface waters, groundwaters and groundwater-dependent terrestrial
ecosystems (Bartley et al. 2009).
This was based on Ireland’s groundwater risk assessment
methods given in DEHLG, EPA and GSI (1999) and EPA (2007). At the field scale, farming practices
and hydrogeological characteristics vary.
Pressures and pathways were examined to identify
agricultural pressure ‘strength’ and hydrogeological ‘pathway sensitivity’. A scoring system assigned
risk of nutrient export to receiving waters. A nutrient budget using these data showed that most fields
in the high nature value conservation area showed low pressure strength scores, with point sources
posing the highest risk. Such sources included animal watering points, feeders, and/or dung and
urine hotspots where animals congregate. The model was developed to work at the field scale, and
the methodology could be adapted to other farming systems.
The project also tried to develop marketing and branding to support local products and low intensity
farming. While the conclusions from this study was that extra income was likely to be marginal to
existing income, the LEADER scheme (see also Section 4.5 above) was identified as a possible
mechanism to promote markets and rural development in the area. An application for funding to the
scheme was unsuccessful because it was “too agricultural in nature” (its basis being farming for
conservation) and went against the principle of cohesion which sought to amalgamate existing rural
groups.”. This was unfortunate, given that the economic study (Van Rensburg et al, 2009b, 2010) of
BurrenLIFE provides evidence that farming for conservation in the Burren has significant externality
benefits for the wider rural economy. The new Burren Farming for Conservation Programme is
expected to have significant tourism spin-off. There is also potential for more local market
involvementto promote high quality local meat production. The BurrenLIFE project benefited from a
payment to farmers to promote conservation that went beyond the limits of the normal pillar 1 and 2
payments under the CAP.
The subsidies, therefore, enabled good farming practice, that was not
implemented under REPS scheme, but which were vulnerable to policy change (Van Rensburg et al,
2009b). They do, nevertheless, provide an example of the type of fiscal mechanisms which can
support regional environmentally sensitive farming.
Stakeholder consultation
59
Farm liaison was a core activity of BurrenLIFE, with regular interaction among participating farms, the
project team and associated partners. While there were a number of formal events, and discussion
groups, there was also continuous intercourse among the range of players. This helped to develop a
strong sense of ownership among many of the participants, invigorating a sense of conservation
awareness and pride in the tradition of the landscape. The benefit of this myriad of interactions
resulted in a very favourable view of the project from local farming families. In a survey of 245
respondents, 32% considered that BurrenLIFE best represented the local farmers, compared with
24% who thought that the IFA (also a project partner) did. The close link with participants of
BurrenLIFE, general awareness raising and activities of the project was further supported with farm
visits to assess planned works under the scheme and formal monitoring of the project objectives. This
involved not just scientific monitoring but gathering farmer feedback to the project.
In summary, BurrenLIFE provided a demonstration, including use of farms to show-case to the
general public its achievements, that agri-environment schemes can be supported by local
stakeholders. Over the life of the project €68864 was paid to participating farmers (BurrenLIFE, 2010).
For participating in the project farmers received payments ranging from €111 to €3176. A total of
€6088 was paid to cover participants to attend workshops and other events. These sums are modest
compared with some single farm payments under the pillar 1 of the CAP. The clear lesson is that
some fiscal encouragement to support the maintenance of an internationally important and highly
valued landscape has cost-benefit, and provided a local sense of pride in farming to support local
heritage. While the BurrenLIFE project was not focused on high status aquatic habitats, although
some were included, the parallels with an approach to protect important habitats, within or outside the
cSAC network, are self-evident.
Case Study 2 . Lough Melvin catchment management
Lough Melvin, on the border between the Republic and Northern Ireland, is an important lake for its
wild fish populations and is designated an cSAC under the Habitats Directive. In recent years it has
undergone deteriorating water quality, with increased phosphorus (P) loads and water colour
(Campbell and Foy, 2008; Girvan and Foy, 2003; 2006). Agriculturally derived nutrients are estimated
to account for more than 50% of estimated P inputs. Farming in the catchment would be considered
low intensity, with extensive sheep and cattle grazing and average stocking density of about 0.5
livestock units ha-1, but nevertheless about 22% of fields surveyed in the catchment contained high
soil P concentrations (Schulte et al., 2009).
There is some forestry, some small villages and a
scattering of single houses in the catchment. Soils are dominated by peats (47%) and gleys (47%)
(Schulte et al., 2009). Both of these soils are associated, for different reasons, with high P transport
rates. Owing to the decline in water quality of Lough Melvin, it is possible that additional measures
will, in due course, be introduced to protect the lake. The Lough Melvin Catchment Management
Plan, and associated Nutrient Reduction Programme was set up as a cross-border initiative to
investigate both the causes and solutions to the nutrient pressures on the lake (Campbell and Foy,
2008). It involved a participatory approach with the stakeholders (Doody et al., 2009a). Based on a
60
risk assessment of nutrient loading (Magette et al. 2007), and discussions with farmers, a number of
potential measures were outlined. These were assessed by stakeholders at two workshops that
included considerations of their cost-effectiveness (Schulte et al., 2009; Byrne et al., 2008).
The Lough Melvin farmers expressed a preference of measures that included feeding low-P
concentrates to livestock, soil analysis and nutrient management planning, installing sediment traps in
drainage ditches, installing hedgerows across slopes and putting hardcore gravel around slopes,
although the latter of these was considered one of the least cost-effective measures. Not surprisingly,
in general the farmers preferred measures that involved little change to current farm practices and or
labour input (Schulte et al., 2009). Nutrient budgeting was an important aspect of the project. This has
been shown in a number of studies (Oenema and Roest, 1998; Goodlass et al., 2003; Cherry et al.,
2008) to be an effective management strategy and one which seems entirely obvious. It is most
effective with regular technical input from farm advisory services (Goodlass et al., 2003).
Although there was an assumption that the Melvin catchment was low intensity, widespread soil
testing as part of the project indicated a very patchy distribution of nutrient status. This highlights one
of the major problems with the current Nitrates Regulations (DEHLG/DAFF, 2010) as operational in
the Republic of Ireland, in that there is an assumption that a soil has a soil index 3, unless shown
otherwise. This also highlights: 1. that small patches of high nutrients if connected by hydrological
pathways can provide disproportionate impact on a receiving water, and is a target for management
(Haygarth et al., 2005; Newsome, 2010); and 2. the inappropriateness of aiming for Morgan’s soil
index 3 in vulnerable catchments (See Section 3.3.1), even though this is advocated by Schulte et al.
(2009).
Catchments containing high status waterbodies and, often, associated habitats of
conservation value, are typified by low input-low output nutrient regimes (Bignal and McCracken,
1996). In Northern Ireland chemical P can only be added after there is demonstration for its need
following consideration of P applied from manures. For catchments containing high status water
bodies, however, and as discussed in Section 3.3.1, there is need to move away from the mantra of
optimal production and the assumption that a soil P index of 3 provides that.
The Lough Melvin decision support process was designed to integrate farmers into the decision
making process, and build consensus. It also provided a mechanism for increased understanding
across all participants. This included those who set up the process, helping to dispel feeling of
mistrust between stakeholders and regulatory agencies. Traditionally this mistrust is fostered through
real or perceived imposition of policy options, although it also raises dilemma in the Regulatory arena.
European member states face sanction for non compliance of Directives. In the case of the Nitrates
Regulations, there is also the feeling among many farming stakeholders that the Regulations are
already too stringent (Schulte et al., 2009), so additional measures are difficult to implement without
stakeholder suspicion or opposition. For the implementation of the WFD, and maybe more so for the
protection of high status water bodies, stakeholder support is crucial. It also sets a context that may
be more challenging than participatory approaches that are less legally constrained. In the Melvin
61
case the participatory framework was drawn from a number of sources (DETR, 2000; Booth and
Richardson, 2001; Wolfe et al., 2001; Bickerstaff et al., 2002; Becker et al., 2003; Vantanen and
Marttunen, 2005; Doody et al. 2009b) outlined in Doody et al., 2009a. The key criteria considered,
reproduced from Doody et al. (2009a), were that:

aims for the involvement of stakeholders are clearly defined;

participants are representatives of the target public;

barriers to stakeholder participation are overcome;

an effective method for interaction with stakeholders is developed;

the outcomes of the process reflect the inputs of the participants; and

feedback is provided to the participants.
These are very similar to the criteria recommended by Acland (2008). Participants were selected on
the basis of farm type, farm size, famer age, gender, and location. Owing to resource limitations,
which will always be a factor, the first phase in deciding the suites of measures involved 25 farms
only. A second phase was to bring these measures to a workshop of agency and regulatory persons.
The third phase was the ranking of measures (reported on by Schulte et al., 2009) with an
independent group of farmers. The process of engagement with the first group of farmers included
individual discussions and walking field sites to both gain a mutual trust, but also for mutual
information dissemination. In an evaluation of the process, Doody et al. (2009a) highlighted the value
of a locally focused process. This can be compared with the more general approach of the national
RBMP consultation process in the Republic of Ireland. Success for locally focused participation was
also highlighted by Wilson et al. (1995) in a catchment in Germany with many similarities to the Melvin
catchment. This provides important insight to the process that might be useful in catchments in
Ireland where the goal is to implement strategies to protect high status sites, and which would be
usefully linked with the Water Management Units (WMU) identified in the RBMPs.
The Lough Melvin catchment work suggested a high potential for success in developing awareness
and catchment management with stakeholder participation. The initial efforts have not been followed
up, and this risks a loss of momentum. This has also meant that there was insufficient time to
evaluate effects of the management on environmental quality (but see comments below). The key
outcome though was the demonstration that a concerted effort with local stakeholders can establish
awareness and a shift in management. For successful participatory management, the process itself
needs to be sustainable. Doody et al. (2009a) makes some very important concluding remarks:
“Integrating the outcomes from this participatory process is complicated by the existing “top-down”
centralized system of decision making around which local and national authorities are structured”;
noting that Ireland’s centralised structure to agri-environment schemes, compared with many other
EU states, is not conducive to “bottom up” participation, and that the Melvin initiative occurred outside
the existing governance structures.
A similar point for the need for appropriate institutional
frameworks for integrated catchment management is made by Haslam and Newsome (1995) and
Mainstone and Holmes (2010) with respect to the U.K. This administrative framework reduces
62
controlled feedback into the decision making process. Mitigating mechanisms may, therefore, be
important not only for potential environmental impact, but also for the RBD structure itself. While
Ireland is subject to the demands of Article 14 of the WFD on public participation and Directive
2003/35/EC, it is the only EU state not to have ratified the Aarhus Convention on Access to
Information, Public Participation in Decision-making and Access to Justice in Environmental Matters
(www.unece.org/env/pp/). Nevertheless, the experiences of the Melvin participatory “experiment” are
encouraging. Doody et al. (2009a) suggest two final points. The first is that the process moved the
level of participation up one notch along the hierarchy proposed by the International Association of
Public Participation (www.iap2.org): 1) Inform, 2) Consult, 3) Involve, 4) Collaborate and 5) Empower,
to “Involve”. This is one level higher than the WFD’s requirement to “Consult”. The second that there
needs much greater infrastructure of civic deliberation to facilitate this at local and national levels
(Abelson et al., 2003).
Like the BurrenLIFE programme, the Lough Melvin catchment programme focused on the process of
catchment management and working with stakeholders. Neither programme was implemented for a
sufficient period to demonstrate environmental improvement. This, however, is common in landscape
management programmes. While focussing management on critical source areas where high nutrient
inputs coincide with surface runoff is well founded (Heathwaite et al., 2005; Gascuel-Odoux et al.,
2009; Trevisan et al., 2010) quantifying a benefit is difficult owing not only to limited timeframes, but
also working within a complex matrix of multiple habitat and practices. Schoumans et al (2011) in
their review of mitigation options refer to the value of catchment modeling to understand nutrient
mobility (Jetten et al., 1996; Sorrano et al., 1996; Cerdan et al., 2002; Kersebaum et al. 2003; Wolf et
al., 2003; Durand et al., 2004; Wade et al., 2005; Aurousseau et al., 2009) as the effectiveness of
such options can take many years (Foy et al., 1995; Smith et al., 2003; Herlihy et al. 2004; Watson et
al., 2007; Schulte et al., 2010). For high status waters this is even more complex, as the objective is
to prevent impact, so there is only the possibility of drawing on existing knowledge and catchment
models, with a likely need to focus management at local scales on a case by case basis (Dillaha and
Inamdar, 1997).
Case Study 3. Rivers Trusts
The Rivers Trusts in the U.K., and River and Fisheries Trusts in Scotland, originally instigated
primarily by angling interests, have developed a wider remit for habitat and water quality protection.
They represent well-organized stakeholder interest groups, with charitable status, which were largely
set up in response to the failure of regulatory authorities to maintain and/ or rehabilitate water quality
in many areas of the UK. They work with landowners, farmers and the wider community through
education, and grant aid. The Trusts operate through catchment planning and stakeholder
involvement, attempting to identify critical sources of pollution and effecting management measures in
both the catchment and riparian zones. Liaison with landowners include development of site-specific
management plans, focusing on cost-effective conservation management. The West Country Rivers
Trusts (www.wrt.org.uk), for example, estimated average annual savings per farm in excess of £1369
63
as part of the Cornwall Rivers Project (2002-2006), funded by the U.K. Government and the E.U. to
tackle diffuse pollution IInman (2005).On the urban fringe on London, the Blackwater Valley Country
Side Trust, dates back to 1971, with the aim to promote conservation and amenity of the Blackwater
Valley, coordinating a range of activities, and commenting on local planning issues (www.blackwatervalley.org.uk). Specific actions are contained in the Blackwater Valley CountrysideStrategy 2011-15
(Blackwater Valley Countryside Partnership, 2010). The strategy highlights management for specific
areas of the valley generally small-scale habitat, landscape, and access improvements through liaison
with landowners. Valley wide projects include accessing funds and wildlife survey. The U.K. River
Trusts provide a useful link between policy and local interest and activities. Activities vary across
amenity and conservation interests, often with the goal of linking the two.
In the Republic of Ireland, organized local involvement in habitats management is generally less
pronounced, although present. For example the mission statement of the Nore Suir River Trust is to
“support research, conservation and public education initiatives that will benefit the freshwater fish
resources, associated fisheries and environment of the Nore and Suir Rivers, and to preserve for
future generations a valuable part of Ireland's natural heritage." (www.noresuirrivertrust.org). The
Trust was active in the campaign to ban salmon drift netting in Irish coastal waters, and this has
spurred further activities in habitat restoration and survey, often working with the local authorities and
Inland Fisheries Ireland. It also has a published strategy for protection, preservation and community
involvement in the management of the river. The strategy includes the aspiration to “build
relationships with stakeholders, volunteer groups and individuals to promote involvement in
education, awareness, utilisation and action plans for catchment improvements”. Similar initiatives are
happening e.g. on the Slaney River (www.slaneyrivertrust.ie), the
Carra Mask Corrib Water
Protection Group (http://www.cmcwpg.ie) and local groups that campaign for protection of the Swilley,
Lough Derg and Lough Ree (see www.swanireland.ie) for more details. Some groups are better
mobilized than others, or have a longer history. All of these groups, both in the U.K. and Ireland are
motivated by a sense that state authorities do not provides adequate protection for a particular river or
lake. While this can provide tension between the individual voluntary and state bodies, these largely
voluntary groups provide the evidence of local interest which could be linked to the protection of high
status water bodies, through liaison with, and support from, public bodies. Local interest groups
provide an obvious conduit to local awareness and have the potential to facilitate good habitat
management.
Case Study 4. The U.K. Environmentally Sensitive Areas, Countryside Management and
Environmental Stewardship scheme.
The U.K. has provided funding for agri-environment schemes in designated sensitive areas since the
late 1980s. The requirements for the initial Environmentally Sensitive Areas (ESA) and Countryside
Management (CMA) schemes were similar, and developed into a mechanism designed to exceed the
requirement of Council Regulation 1782/2003 for cross-compliance of protecting the environment in
order to receive direct agricultural support. Participation was a voluntary agreement to farm in an
64
environmentally sensitive manner, so as to reduce agricultural impact on biodiversity and water
quality. It was open to all farmers in the designated landscape areas, of which (for example) there
were 5 in Northern Ireland and 22 in the England. Over this period there were also a number of
iterations of these countryside schemes.
The CAP reforms in 1999 allowed an increase in
expenditure on rural development measures under the new EU Rural Development Regulation (RDR
-Council Regulation (EC) No. 1257/1999). The current most applicable scheme for wider countryside
protection of high value landscape and its environment schemes is the U.K. Environmental
Stewardship scheme, managed by Natural England.
Environmental Stewardship schemes cover three tiers. Tier 3, the High Level Stewardship (HLS),
aims for significant environmental benefits in priority areas, implemented through agreements that last
10 years. It is governed by EU Council Regulation 1698/2005 and Commission Regulations
1974/2006 and 1975/2006 (as amended). The focused nature of the HLS on targeted management,,
that is supported with professional advisors, makes it applicable to the type of local management
objectives that would benefit high status water bodies in the Republic of Ireland. The U.K. scheme
also allows additional funding for complementary works from other schemes such as the Heritage
Lottery Fund and the Hill Farm Allowance. Ina similar way development of such an approach in
Ireland could provide a framework for cooperative funding, and even provide incentives for targeted
areas containing high status water bodies, or the restoration of water bodies of national significance
that could be restored to high status. All that is really required is a more imaginative and focused
approach to use of CAP funds for environmental protection and enhancement (see Section 4.5).
Conclusions from the Irish/U.K. case studies
Building understanding of catchment processes and management to reduce, mitigate impact from, or
prevent, nutrient loading to surface waters is complex and time-consuming.
One of the main
difficulties in stakeholder participation is the need for certainty that management approaches will
deliver a better environment with minimal costs in time and money, and especially within the
timeframe of most studies. Hydrological complexity, a landscape mosaic and varied approaches to
management often makes it difficult to quantify the success of measures, and mitigation almost
always requires a range of interconnected measures (Cherry et al., 2008).
Even in the reduction of
point source pollution this has been challenging (e.g. Neal et.,al , 2011). Nevertheless, demonstration
of reduced nutrient loads from diffuse pollution as a consequence of management is required by
some U.S States as a condition of grant-aid for protection of some high quality sites (see below,
Section 6.2).
The expectations of stakeholders can be very different from that of the technical experts charged with
the implementation of an environmental policy.
There is, therefore, a need for considerable
interaction with stakeholders in order to build trust for implementing potentially successful
management. A key aspect is the willingness for stakeholders to participate and for mutual learning
from all actors. Progress towards mutual understanding and problem solving can be supported with a
65
range of techniques and models, and use of multi-criteria assessment or exploration of scenarios and
mathematical decision support models. This can be a powerful aid to decision making and consensus
building in order to optimise the use of knowledge, propose realistic outcomes, and provide for more
effective decision-making (Watson and Wadsworth, 1996; Schauser, et al, 2003). Coupled with this is
the need for stakeholders to appreciate uncertainty in model predictions and catchment management.
The early experiences of RBMP participatory approach both in Ireland (Irvine and O’Brien, 2009), the
UK (Watson and Howe, 2006; Petts, 2007; Blackstock and Richards, 2008), and in other EU Member
States, indicates a process that is still developing and varied in quality. Programmes to reduce
loading of nutrients or other pollutants, by their very nature, are efforts to restore rather than prevent
impact. Restoration programmes targeting diffuse pollution are subject to lag times, and complex
scalar effects. With some exceptions such as the protecting the source of New York’s drinking water
(Dell et al., 2009-see below) and of Vital mineral water (Gras and Benoît, 1998), programmes to
maintain high quality waters through management are scarce. Generally, the need for prevention,
although cost-effective, is overshadowed by the need to restore. This provides a crucial dimension
for protection of high status waters as it is impossible to demonstrate the success of preventative
measures other than the maintenance of no impact. There is, however, over a century of research
and knowledge that demonstrates the impact of pollutants on water bodies and human health.
Regulation by itself, as done for example through the Nitrates Regulation (see Section 3.3.1), can
also only go so far in achieving environmental objectives, even if the underlying principles are sound.
For the protection of high status sites a need for more local and incentive based schemes are
essential. It is only through local engagement that the appropriateness of local management options,
such as buffer strips, fencing, stock holdings, hot-spots of sediment and nutrient emissions, and
attenuating water movement through local wetland creation can be assessed. Newson and Chalk
(2004), however, warn of the dangers of stakeholder opportunisms unless there is adequate provision
of information to support management (and financial) decisions.
While awareness among stakeholders of an environmental problem is necessary to promote good
practice, it is seldom sufficient by itself without supporting mechanism such as market incentives or
targeted grants (Curry, 1997; Curtis and Delacy, 1998; Rhodes et al., 2002). Newson (2010) puts it as
“Securing catchment-relevant patterns of enterprise management is perhaps the most secure way for
catchment controls to survive”, and, like Doody et al. (2009) advocates use of the market place, albeit
a subsidised one.
Public authority awareness campaigns themselves can be undermined by
conflicting information or denial through vested interests, as has been apparent in the past through
lobbying by large farming organizations.
Change to management occurs when landowners are
aware of a problem, and have the knowledge and resources to address it (Wilson, 1997; Rhodes et
al., 2002; BurrenLIFE, 2010). Awareness campaigns, involving not just technical knowledge but also
promoting available fiscal schemes are most effective when a variety of techniques are used, and
which integrate technical knowledge, and support mechanisms, with public values (Bohneblust and
Slovic, 1998).
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5.3 WFD RBMP and public participation in Europe
The WFD requires that all EU Member Sates prepare RBMPs and, as a step towards that, assess
risks of failing to meet environmental objectives through their Article 5 reports. Article 14 of the WFD
requires a public consultation process on the RBMPs. This process has provided challenges across
Europe, although some countries were better set up for these than others. Hering et al. (2010) provide
a review of this across Europe, and the achievement made after the first decade of implementation.
In some cases this appears extensive and open.
Hering et al. (2010) identify North Rhine-
Westphalia, Germany (http://www.flussgebietenrw.de/Mitwirkung/index.jsp), the Basque country
(http://www.uragentzia.euskadi.net/u810003/es/contenidos/informe_estudio/planificacion_dma/es_doc/indice.htm), Finland and Sweden as
examples where extensive and open consultation occurred with stakeholders. An overview of
European RBMPs river basin management plans can be found on http://cdr.eionet.europa.eu/ and on
http://ec.europa.eu/environment/water/participation/map_mc/map.htm..
A series of summary country reports provided by participants at a meeting in September 2010 in
Luneberg, Germany to discuss “Public participation and RBM in the implementation of the WFDtaking stock and looking forward” identified a large range of intensity of public engagement, from
minimal
to
serious
attempts
to
involve
stakeholders
and
the
public
(www.waterscale.info/index.php). The meeting represented 18 countries, and 16 EU member
states,who had experience in the WFD public participation process. There were many similarities in
the view of public participation in the preparation of the RBMPs. In addition to the examples provided
by Hering et al (2010), some German Lander (e.g Schleswig-Holstein and Lower Saxony, Kastens
and Newig, 2008) also appear to be good examples of an active public participatory process, although
others, often from the former GDR, have had a more minimal consultation. In Spain, water
management historically has been dominated by powerful economic interests, such as irrigators,
hydro-electric power generation and urban suppliers. Since the late 1990s, however, this balance of
power has been shifting in some of the States. In Catalonia, a regional “Sustainable Water Use
Council” was created in 1998, with representatives from a broad spectrum of interested stakeholders.
The WFD has strengthened this process, with what appears to be a genuine interest in public
engagement, and political commitment to provide more effective water management, taking account
of citizens views.
This has involved a significant shift in the traditional culture away from the
dominance of a heretofore technocratic view.
For many countries (Austria, stands out as a notable example), however, the process appears to be a
minimal effort to meet the requirements of Article 14, and that water protection is considered to be
“seen merely as a technical and scientific problem by the authorities”, who adopt a paternalistic and
technocratic approach to WFD implementation. As such implementation of Article 14 is used more as
a mechanism to educate, inform and prevent conflicts (Fritsch and Feichtinger, 2009). A similar
67
perspective would seem to apply to a large number of agencies across Europe, with the public
participation for the WFD primarily there to increase the legitimacy of the process rather than to try
and understand and address stakeholder values. This often relates to a view to limit investment to
keep costs down, while being WFD compliant. The challenge for the future is “how to change a
practice that is faithful to the letter of European Directives, but indifferent to its spirit” (S. van der
Arend, Delft University of Technology, pers com.).
For many EU states or regions, there is not a strong history of proactive engagement with
stakeholders or the public on issues of natural resource management, apart from engagement with
e.g. the industrial representatives of fisheries and agriculture. Other E.U. Member States such as
Finland, Sweden and Denmark have has long traditions of public engagement. The RBMP
consultation in Denmark elicited nearly 2500 suggestions from a range of stakeholders (O. Fritsch,
University of Cophenhagen, pers com.), but was countered somewhat by the Danish government’s
approach to keep costs as low as possible. In Denmark, the Ministry of Finance plays a central role in
implementation of the WFD. This is new for a European environmental Directive in Denmark. For
many of the former Easter bloc states, public participation in environmental decision making is
relatively novel.
A strong public and stakeholder participatory process provides a foundation for a structured approach
to site protection. This is extremely relevant to the protection of high status sites, because of the
inadequacies of the process to protect them within the Irish legislative framework, as discussed in
Section 3. This lack of protection for high status sites appears to be Europe wide. Several
investigative email shots and personal contacts with European workers suggests that the focus of the
WFD on achieving good status for water bodies has left a significant gap in the legal or administrative
framework for the protection of high status sites. This is, therefore, of enormous importance not only
for Ireland but across the E.U. It is not clear if there has been a similar decline of these sites across
Europe as has occurred in Ireland. Of itself the requirement under the WFD for no deterioration of
sites appears a weak provision to safeguard the best quality aquatic sites that lie outside the Natura
2000 network. As already discussed in Section 3, the Natura 2000 network does not, in any case,
ensure that aquatic sites within the network are protected sufficiently.
WFD implementation within the EU is in its infancy, and still within its first cycle. The WFD was seen
by many as a radical document that perhaps has not lived up to immediate expectations (Moss,
2008), despite a concerted and wide ranging technical input (Hering et al., 2010). Implementation is
hindered not only by technical challenges, but by political and administrative structures. In Ireland a
deep rooted fragmentation of water policy is clear. As reviewed in Section 3, existing POMs as listed
in Annex VI of the WFD do not provide an effective basis for the protection of high status sites, and
alternative mechanisms (referred in the WFD as Supplementary Measures) are required. Some of
these may be legislative, but there is also scope to employ and develop a range of other approaches
such as market based mechanisms and locally focused stakeholder involvement.
68
The socio
economic setting and a largely centralized policy arena, has diminished a perceived relevance of a
participatory management approach in Ireland, and elsewhere in Europe. Yet, such an approach is
increasingly promoted in the less developed world and in the U.S. (see next Section) as a mechanism
for sustainable management and promotion of environmental stewardship (Burger, 2002; UNDP,
2011).
Section 6. Experiences outside of the EU useful for the
protection of high quality habitats.
6.1 Introduction
The final Section of this review summarises some strategies used outside the E.U. that can inform
adaptation and development of policy for protection of Ireland’s high status water bodies. These
strategies can also be applicable to the restoration of targeted sites back to high status. The
protection of Irish heritage should not be wholly dependent on the threat of EU sanction. The review
scanned a wide range of countries to explore potential mechanisms applicable to the protection of
Ireland’s high status water bodies. These inevitably require landscape management. The best
examples were found from the U.S. and New Zealand.
6.2 The U.S. Clean Water Act 1974: A Tiered system of protection for water bodies.
In the U.S., the federal Clean Water Act (CWA) provides for a tiered system of protection, with the
most protective designation offered being tier 3 for the highest quality waters, designated as
‘Outstanding National Resource Waters’ (ONRW). U.S. States are required to develop water quality
standards (WQS), subject to US EPA approval and which identify designated use, water quality
criteria, and a three-tiered anti-degradation system. While all States must designate tier 1 and 2
waters, there is no requirement for states to designate tier 3 waters, and where this has occurred, the
level of protection can vary considerably. Twenty-seven states have designated tier 3 waters, with 17
having designated tier 2.5 waters as a compromise between tiers 2 and 3 (Edmonds, 2010). Where
the ONRW designation is utilised to its full potential (e.g. North Carolina) it can provide effective
protection. Several elements of the CWA are worth noting as they go beyond the provisions of the
WFD, including provision for stakeholder participation in both designation and management. While the
WFD designated quality status according to an Environmental Quality Ratio (EQR, see Section 2), in
the U.S. stakeholders can nominate waters for designation as ONRWs.
Under the CWA, maintenance of water quality centres around designated use as a public resource.
Once a designated use is established for a water body, the State develops the water quality criteria
necessary to assure and maintain this use, based on ‘sound scientific rationale’. The WQS consider
69
a water’s assimilative capacity for different levels of pollution, dependent on its designated use,
defined as total maximum daily loads (TMDLs) under Section 303(d) of the CWA.
These allow
States to prioritise waters of higher concern, and assist regulators in establishing permit limits for
discharges to ensure discharge commensurate with WQS. As opposed to tier 2 designation, where
pollutant increases and any lowering of water quality must be approved by the EPA, no permanent
degradation is allowed in tier 3 waters or their tributaries, although certain activities which only result
in temporary and short-term changes in water quality are allowed if they do not affect the beneficial
use.
A noteworthy provision is that an entire landscape areacan be designated as an ONRW,
providing it with tier 3 protection.
The most comprehensive policy cited by Edmonds (2010) is that of North Carolina, the key features of
which include:

Site-specific management strategies, specific stormwater management and stringent
regulations relating to dredging and infilling;

A buffer credit system, whereby a minimal buffer area is specified, and additional credits can
be awarded for buffer zones which are larger than the required width. The minimum buffer
width for sensitive stream mitigation projects is 50ft (ca 15 m), with 30ft (ac 8 m) required in
the mountainous areas; no buffer credits are available for buffers less than 15ft (ca 4.5 m).
Buffers for certain high risk areas are required to be designed for optimal nutrient removal.
For example, in the Tar-Pamlico and Neuse Rivers within Pitt County of N.Carolina there is a
“Riparian Buffer Protection Ordinance” that provides for a mandatory buffer strip of 50 ft
adjacent to surface waters in order to stabilise stream banks and reduce diffuse nutrient
pollution. Buffer strips comprise undisturbed vegetation for the first 30 ft by the water and an
additional 20 ft of managed vegetation.

Comprehensive nutrient control strategies are developed for each site by local advisory
committees. Nutrient removal functions that were provided by trees prior to their harvest shall
be replaced by other measures that are implemented by the owner of the land from which the
trees are harvested.

Standard best management (BMP) practices for various pollution sources, e.g. agriculture,
must achieve a minimum of 30% reduction in nutrient loading; other BMP options can be
imposed for sensitive areas and BMPs are reviewed every five years and upgraded if they are
not effective.

More targeted ‘High quality’ BMPs have been applied in some states.
Any person can petition to classify a surface water of the state as an ORW. A public hearing
is mandatory for any proposed permits to discharge to waters classified as ONRW.
70
‘Basin
Oversight Committees’ comprise various authorities and stakeholder groups who reviewing
and recommend updating management strategies.

Section 401 certification [CWA] requires analysis of cumulative impacts to downstream water
quality.
In 1987, amendments to the CWA established Section 319 Nonpoint Source Management
Programmes, which offers finance for water management and restoration. Under Section 319, states,
territories and tribes may receive grant money which supports a variety of activities to assess the
success of specific nonpoint source implementation.
North Carolina’s Agricultural Incentive
Programme draws Section 319 grant funding (from the US Farm Bill) to encourage environmentallyfriendly farming.
Other policies in the U.S. that have reduced non-point pollution, include an agriculture Cost Share
Program for Nonpoint Source Pollution, Conservation Reserve Enhancement Program, Conservation
Reserve Program, environmental Quality Incentives Program, a wetlands Reserve Program, and a
Wildlife
Habitat
Incentives
Program.
The
Wetlands
Reserve
Program
(http://www.nrcs.usda.gov/programs/wrp/) provides funds for conservation programmes and, flood
mitigation programme for wetland enhancement. See Williams et al, (in press), for a review of the link
between flood mitigation and wetland quality relevant to Ireland.
6. 3. California Integrated Regional Management Plans
The California State Integrated Regional Water Management (IRWM) Planning supports sustainable
water uses and environmental stewardship, achieved through stakeholder partnerships, and financed
through the California Department of Water Resources and California State Water Resources Control
Boardthrough competitive grants. The programme developed from the CWA, to provide a more
flexible and encompassing approach to integrated management. IRWM projects will typically involve a
range of stakeholders who will work towards prioritising water management objectives. This has led
to partnerships among conservation organizations and public agencies, environmental justice groups,
and small water companies. IRWM appears better able to address non-point pollution than the CWA.
As the regulatory framework can fail to protect high quality sites, civic partnerships present an
opportunity for more effective protection.
6.4 Source water protection for New York City
(http://www.epa.gov/r02earth/water/nycshed/filtad.htm)
The U.S. federal Safe Drinking Water Act requires filtration of drinking water which comes from
surface water sources (i.e., lakes, streams, reservoirs). This requirement, which is meant to protect
against waterborne disease, can be waived if a water system provides safe water, and if its catchment
is protected to ensure that safety in the future.
71
New York City’s water supply originates from three
upstate catchments; the Catskill, Delaware and Croton systems, encompassing approx. 5,100 km 2. It
is the largest unfiltered water supply in the United States, serving 9 million people daily.
The
programme is estimated to cost approximately 3 million US dollars annually, with a saving of an
estimated 1 billion dollars cost required to add filtration to the city’s water supply (Dell et al., 2009). To
avoid that expenditure, the city embarked an integrated water resources management approach to
protect the Catskill/Delaware catchment. Cost–benefit ratio has been estimated as up to 1:10 (TEEB,
2009). Stakeholders involved in the development and implementation of the programme included
foresters, landowners, farmers, government official, technical agencies and business. Buffer strips
form part of a suite of measures that involved management agreements with farmers for low intensity
farming (Bryant et al., 2008). There were strict legal requirements for waiver of the need to provide as
filtration system, and the scheme was agreed only after intense negotiation over a period of about 18
months, and involved a number of Federal, State and local bodies. Costs are born by the New York
City taxpayers. With their input, an integrated “Watershed Development Programme” was effected,
balancing economic growth with drinking water protection for NYC. The programme comprises:

Land Acquisition and Stewardship Programmes involving purchase of water quality sensitive,
undeveloped

lands,
especially
near
reservoirs,
wetlands
and
watercourses.
Partnership Programmes: With support from NYC, foresters, landowners, environmentalists,
loggers, farmers, government officials, technical agencies, and businesses to develop
participatory “Watershed Agricultural Program” and the “Watershed Forestry Program”.

Wastewater Management: The NYC Department of Environmental Protection has
collaborated to construct centralized sewage systems, and to inspect and rehabilitate septic
systems.

Stormwater management: Funding from NYC is available for implementing measures to
prevent contamination of stormwater, e.g. stream corridor protection projects and improved
storage of sand, salt, and other de-icing materials.
A Watershed Memorandum of Agreement in 1997 established a partnership between the US EPA,
New York State, New York City, the many counties, towns and villages located throughout the
catchment, and environmental NGOs. This agreement created a framework for catchment protection
programmes.
Key elements to the success of this were: participatory stakeholder involvement,
guided by local leadership; early buy-in from farmers and other stakeholders who traditionally mistrust
regulators; and connecting policies for economic development with those for sustainable
management.
6.5 New Zealand Landscape Assessment
72
The New Zealand Resource Management Act (1991) (RMA) provides a legal requirement to protect
outstanding landscapes.
To fulfil this requirement, a landscape assessment establishes the relative
importance that should be attached to different types of landscape and their individual components,
and likely effects of different types of development. For each landscape unit, an overall sensitivity
rating is assigned, ranging from 1 (low sensitivity) to 7 (extreme sensitivity). The Landscape
Assessment aims to: assign landscape categories and identify their dominant characteristics;
determine landscape vulnerability; and prepare maps and policies to identify landscape management.
A biodiversity strategy, published in 1999 recognised the importance of extending conservation
protection and restoration beyond the boundaries of protected reserves, to cover private lands. This
led to the development of existing policies and provision of new ones, including fiscal incentives, to
promote a wider countryside approach to habitat protection.
New Zealand has some striking similarities to Ireland in that its agriculture is based on
grassland
farming and it promotes itself as a high quality environment for tourism, with often a focus on
freshwaters. Like Ireland there is concern about the effects of intensive grassland farming on water
quality, and about 50% of dairy farms have soil P above agriculturally optima (Wheeler et al., 2004;
Quinn et al., 2009).
A major difference is that New Zealand abolished farming subsidies in 1984.
Nevertheless, the farming economy improved, but with associated water quality problems. (Monaghan
et al., 2008; Wilcock et al., 2009).
The RMA provides the basic provision for good agriculture practice that includes a series of
regulations and sanctions. The use of non-regulatory mechanism are, however, increasing promoted
by the Regional Councils in order to manage diffuse pollution and enhance biodiversity on pastoral
land (Quinn et al., 2009). These initiatives are supported, and made more effective through fiscal
incentives for schemes such as fencing and buffer strips (Rhodes et al., 2002; Monaghan et al., 2008;
Wilcock et al., 2009). Riparian management incentives have had positive impacts of water resources,
and there are now a variety of schemes within New Zealand that are designed to improve the natural
environment (Quinn et al., 2009). One industry-led initiative is the “The Dairying and Clean Streams
Accord”, which arose through public concern of the effect of dairy farming on water quality.
Reporting is based on self-assessment, and while this has received criticism as not involving
independent monitoring (Deans and Hackwell, 2008), Quinn et al. (2009) consider it a major step
forward in changing farmers’ behaviour, and reducing environmental impact.
Other initiatives used in New Zealand to manage diffuse nutrient run-off include the use of attenuation
ponds, nutrient management computer-aided decision support tools, and issuing of licences for
farming in sensitive areas (Quinn et al., 2009). The potential success of many of the mitigation
options available to New Zealand farmers in experimental systems are summarised by Quinn et al.
(2009) from work in experimental systems. While such success may not be as prevalent in real
farming situations, good management has been shown to lead to reductions of >70% of both N and P
loads from such activities as reduced or better management of inputs to land of farm effluent
73
(Houlbrook et al., 2004; 2006), and fencing riparian zones (McDowell, 2008). In the generally subsidyfree environment of New Zealand farming, farmers will only adopt nutrient management if it is not a
net cost (Bewsell et al., 2007). Strategies have been developed in New Zealand to mitigate effects
on water quality arising from dairy farming across a number of catchments in order to try and
reconcile the sometimes conflicting objectives of intense agriculture and protection of a high quality
environment (Monaghan et al, 2008). While this is applicable in general for agricultural management
in Ireland, the scale of risk, and mitigation, for high states waters requires a more preventative, rather
than restorative measures.
74
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