Management Strategies for the Protection of


Feb 22, 2014 (7 years and 10 months ago)


EPA STRIVE Programme 2007

Management Strategies for the Protection of
High Status Water Bodies. A Literature Review



University of Dublin, Trinity


Kenneth Irvine and
Ní Chuanigh

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: Website:



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 a
nd promoting environmental research.


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


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.


Published by the Environmental
Protection Agency, Ireland 4


Details of Project Partners

Kennenth Irvine

School of Natural Sciences, Trinity College, Dublin 2


01 8961366



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
. 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 simil
ar 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
reference state, it provides a baseline against which monitored sites can be compared. Within the 5
point classification scale of the WFD, reference sites represent the upper end of the scale of
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 equa
te 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 num
ber 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 i
ntensification 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 inc
reased 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

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 (9
2/43/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, an
d 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

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) r
ely 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 w
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 6
10 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 Nitrat
es Action Programme is designed so that modern farming is
compatible with WFD compliance for
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 protectio
n 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 exporte
d 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


Directives, nine others are listed in Annex VI Part A which are to
de Basic Measures for implementing the WFD. The

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 wi
der 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

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

(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 li
nks 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 licens
ing condition, and location of spreadlands should be clearly identified and available to local
authorities and the EPA.


Environmental Impact Assessment

Directive (85/778/EEC) as amended by Directive 98/83/EC,
links with the new Planning Act 2010. The
re 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 Ma
nagement 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 i
n 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 b
acteria 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 hea
lth 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)
ctive (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 ar
e 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 Regulat
ions 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 aquat
ic 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
vident 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, t
o address
the highly fragmented structure of the country's water governance.


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 opportu
nity 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 int
ernational 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. A
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 No
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. Th
e 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 prote
ction 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 confusio
n 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 s
tatus 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 integr
ation 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 pr
otection 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. 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 r
iparian 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.

ithin 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 site
s 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 stakeho
lders 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 re
sources 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, inclu
ding 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 th
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 par
ticipating 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 include
d 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
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


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 so
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 loca
l 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 (C
WA)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 differ
ent 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 phos
phorus 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 i
s 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 bu
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 devel
opment 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 w
ithin 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 t
he 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 obligation
s under the Ramsar

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



“Protection against losses needs to be seen to have a value in
the same way as
tation of goods”

(Newson, 2010).

onitoring and determination of high status sites



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” inde
x used in central Europe from early in the 20

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
orerunner 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
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

(Champ et al., 2
009). 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 person
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

etwork 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 decl
ine 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
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 ce
nt of the total sampled in the 1987
1990 period to less than 17 per cent in 2006

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

lakes (Taylor et al., 2006). The decl
ine in
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
y of our lakes has disimproved dramatically in a short number of years”

(Downey and Ní Uid,

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 go
od 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 moni
tor 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
typology, salinity can m
imic a response to pollution

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 draina
ge (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

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
, 2003
) is that these sites which are depen
dent on groundwater
bodies or are protected areas (all turloughs are within the former, and
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)
consultative procedures that
support protection of high quality sites, including case studies from outside Ireland. Objectives 1 and
2 depe
nd 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 20
15. The WFD environmental
objective that prohibits decline of class of a water body has received far less attention

therefore, review what is meant by high status and how that is determined, before moving on to the
review of how national legislati
on and policies protect high status sites, including cross
with other policies.



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 b
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

distinction between
“reference condition” and high status (McGarrigle and Lucey
2009); Pado et al.,

This allows setti
ng an anchor point for reference states from which

a departure from this

estimated as an Ecological Quality Ratio (

can be estimated, following the logic that all
sites lie along a continuum of quality irrespective of the status class in whic
h they are classified.
eference sites are

de facto

high status, but not all high status sites will
be in


Strict screening of proposed reference sites
across Europe
is now

in terms of land use and
water chemistry
. The EQRs for these sites
normalise the
estimation of metrics, and has been
successful in aligning E
Rs across
widely different

body types. While distinguishing between
reference and other high status sites has led to a more robust classification p
rocess, 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 magnitud
e anthropogenic pressures.

Identifying reference conditions

, however,

difficulties that

can influence
status classification (Kelly
Quinn et al., 2009). While the EU
wide Iintercalibration process has
attempted to provide
a consis
tent approach, the setting of reference conditions remains

a significant
challenge (
Erba et al., 2009;

et al, 2009).
There are also important, and largely unresolved

discussions of what is “natural”
Boon et al.

) or the effect that
alien species

(Stokes et
al., 2004; Maguire et al., 2005)


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

3.2 below.

The lack of suitable reference sites in some regions (e.g Bennion et al., 2004; Borja et al., 2007

led to

selection of ‘least impacted’ or ‘best available’
as reference
sites. This is not the same as
erence state (Irvine et al., 2006
although initial
ly a preferred

option by many Member States

Lucey, 2009),
and one that
also influenced the decision making process in the
Republic of Ireland (Kelly Quinn et al., 2005; 2009; Dodkins et
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 report
s prepared
by the local aut
horities 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

was based on sites attaining the highest quality of the EPA Q
selection of
reference sites in Ireland was based on expert opinion and existing data


than a pressure
analysis as used, for example,
in the EU intercalibration for stream invertebrates (Erba et al., 2009)
There is a risk that

fication of “minimally impacted” (
as required by the WFD; European
Commission, 2003a)

based on the biological communities
rather than

an assessment of minimal

risks a circular, or self
fulfilling, premise based on long
standing perceptions rather

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

While the density of
monitoring points in Ireland is one of the highest in Europe (M.M
Garrigle, 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 o
n 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


still under development
, but are likely to be based on amalgam
ation of unmonitored sites in water
management units (WMUs).
For lakes
verification of putative reference sites was done through

(Leira et al., 2006)
, but requires much more work to provide a comprehensive
coverage. This will be done throu
gh 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.
compliant classification systems and models which incorp
multiple, and potentially interacting pressures compounds the problem
, and require further
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” (

), “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.
“historical condition” (Young and Sanzone 2002; Nijboer et al., 2004)
Setting a particular
date for such impact is, h
owever, ill advised (
European Commission,

; Taylor et al
, 2006).
Verification of putative reference lakes,

using diatom records in sediment cores, showed that 11 out of
35 candidate


reference lakes were
verified as being
in reference condition
(Liera et al., 2006).
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
with in
creased 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).


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
, 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 acr
oss identified reference sites and within water body types, are
nested along spatial (Little, 2008), biogeographical

(Moog et al., 2004; Borja et al., 2009
), 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 f
urther 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 c
lass 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
t of ongoing work (

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 quit
e 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 acro
ss regions

(Borja et al.,
Noges et al., 2009
). In general

the development of classification techniques has
focused on metrics based on the composition of a variety of biological groups, rather than attributes of
ecological function



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., 200

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.

to account for
temporal or spatial

variability (Irvine, 2004), or
there are

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

in dose
response relationships

each waterbody responds
to pressures

(Yarrow and Marin,
2007). Variation in natural background conditions can also mimic anthropogenic disturbance (Fabris

et al., 2009; Hartnett et al., 201
). Site
specific assessment requires long
term monitoring, or robust
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
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 an
thropogenic impact is self
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 i
n Irish rivers demonstrated by long
term monitoring (EPA,
), 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
Generic land

se mode
ls have limited v
lue 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

, impacts is a feasible option

and li
ely much more cost

than large scale restoration

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 bi
odiversity, 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
atus through implementation of the WFD is realised (Meyer et al., 2007
; see also

Habitat variability, or patchiness, at local scales tends to be greatest at low levels of impact
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 an
d global

ambitions to halt biodiversity decline (
of the Convention of
Biological Diversity, 2001; EC, 2011). Globally, aquatic ecosystems are the most impacted habit
by human activities and continue to decline at an alarming rate (Groombri
dge 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 encourage
d drainage and
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
he 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

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; Rams
ar Convention Bureau, 1996), and by 22 November 2001, 130
States had become Contracting Parties
. As

of 10

June 2011, 1933 sites
have been declared
Wetlands of International Importance.

While the Ramsar convention does not feature prominently in
sion 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., 199
9; 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 (Lu

tz 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).

Ireland and elsewhere many high quality sites
found in the
headwaters of rivers
ing ecological communities

of fundamental biod
The same principle applies to small standing water
heir protection is well justified

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,

3.7) highlights the importance of small wetlands. In 2007,
high qualit
y 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 ). T
his 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

education; the improvement of damaged landscapes; and the integration of landscape in all relevant
There is certainly scope for a greater integration of landscape protection (Heritage

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

. The revised Planning
Development (Amendment) Act, 20
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.
ples of particular interest for
Ireland may be the approach of the French Regional Parks (; Guihéneuf

2009), The Catalan Landscape Observatory (
) and the Canadian
Heritage Rivers System (www.c

The French Regional Parks provide

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

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 governm
ent to develop small agricultural areas for biodiversity conservation.
PNR must
adhere to principles


sustainable development
PNRs aim to make agricultural and forestry
practices compatible with the conservation of natural environments by establishing agri
contracts. This approach
has led to

the establishment of the 46 Regional National Park
. Similar
are available to Irish Local Authorities under

204 of the Planning
Act 2000 to establish Landscape Conservation Areas but has been under

The first
Landscape Conservation Area may be established in the Tara/ Skyrne Valley
in Co. Meath

The Catalan Landscape Observatory aims to increase the knowledge of landscapes among Catalan
society and to support implementation of the European Landscape Convention by
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 preservatio
n, and propose the quality objectives to be met”. The Landscape Catalogues
integrate landscape into planning
, promoting

the diversity and value of landscapes.

The Canadian Heritage Rivers System is a non
statutory model that works through re
cooperation to supp
rt community involv
ment in the protection of rivers, designated on the basis of

importance for local her
tage and recreation. While largely set up to pr
vide a cooper
k to support protection of large rivers,

ent of indigenous communities, the

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
tes landscapes. This is likely a key factor in securing stakeholder participation in protecting
high status water bodies


. The designation of ‘high status’ waters according to purely
scientific criteria under the WFD may not inspire a cultu
ral 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 open
conference o
n 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 c
ould facilitate
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

), a
Landscape Characterisation Assessment

been developed
by the U.K. Department of
Environment, Food and R
ural 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 th
country (Fig

). 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 ref
ion 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.
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

3.3 and

). There is only moderate, and chance,
overlap between high status sites and

candidate Special Areas of Conservation (
. 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


the glob
al biodiversity crisis (Abell, 2002)
requires a new paradigm of thinking.


Figure 1.

High status surface water bodies nationally as identified in the RBMPs 2009


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 Obser
vation Networking improving Europe's environment
), a network of the European Environment Agency could


possibly linked to

Term Ecosystem Research sites (LTER:


(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
is difficult to
distinguish from
reference state
, and s
hould be considered synonymous
with it for management.

regional connected network of
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

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
on, fundamental to the implementation of the WFD, is further supported through other
legislation such as provided by Environmental Impact Assessment.


2. L
egislation and Policies for the Protection of
High Status


The W
FD 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

impacts. Similarly, while the very essence of the WFD is ba
sed 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

favourable conservation status (
Withrington, 2005) and where this, by default, achieves

high status under the WFD (see below,

3.2). Under the WFD, the mechanism to effect
environmental objectives are the Programmes of Measures (POMs), outlined in Article 11 of the
Directive. I
n 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
by the IMPR
ESS 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 ( Detailed ass
essment 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);


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 Protec
tion 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

the minimum measures required to meet the environmental objectives for good status. In
previous Gover
ment 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,

referred to
in Annex VI, part B of the WFD.

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
to protect aquatic system
over and above those provided by the
legislation and the 11 associated Directives

an option.

This is of crucial importance f
r the
protection of high status. It is also important
that a) there is no fundamental conflict between the
bjectives of those Directives

listed in Annex VI part A
, and any subsequent amendments, and those
of the WFD and b) that sufficient attention

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

cross compliant

with the WFD. This

reviews these questions at three scales:





legislation in Ireland and if this is fit
purpose for protection of high
status sites;


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


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

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 bee
n sufficiently
effective for water governance in Ireland which,

in common with environmental protection policies in

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 resource
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 o
f 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 le
ss 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 authori
ties. 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
in harmonising the WFD with the
Habitas Di
rective, 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 n
POMs to

water managment problems in areas protected for nature conservation.
This is discussed


3.2 below.

There is also no

in the European Communities Good Agricultural Practice

for Protection of Waters S.I. 610
of 2010

of strategies to protect

sites designated under the

The Surface Water Regulations


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, af
fecting both surface and
groundwaters. Waste Water Treatment Plants (WWTPs), septic tanks (unsewered
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
tified. 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 o
f 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


For the mo
nitoring 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 con
tained in Standard
Operating Procedures used by the agencies charged with these estimates.
Techniques for
developing EQRs for different elements

water bod
es 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 st
for integrated
monitoring and protection.

The relevance of an integrated assessment process can be illustrated by lo
oking 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

of unfiltered molybdate

reactive phosphorus (MRP) as

a median value or < 0.045 mg l

as a 95 percentile. These values are
equivalent to 25 and 45 µg l

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 nu
trient 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, considere
d that “Reference or natural concentrations for soluble
reactive phosphorus in rivers are typically in the range 0
10 μg l
Moorkens (2006) used a median
value of 5 μg l
MRP, 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 1
5 μg l

This is
the target concentration
for the upper catchment lakes that supply source water for New York City (Dell et al., 2009
, and see

Using information from diatoms in sediment Foy et al (2003) estimated a reference TP in
Lough Neagh of 19 μg l
. 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
, with many between 15
20 μg l
. Anderson
(1997) estimated diatom
inferred TP concentration in s
ix small rural lakes in Counties Down, Armagh
and Tyrone to range from 6
58 μg l
, all from before 1900. The interpretation of diatom
inferred (i.e.
modelled) TP needs to be treated with some caution (

et al.,
, and in Ireland significant
act 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 concen
trations (EPA data) < 20 μg l
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

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

are likely
<10 µg TP l

(Hobbs et al., 2005; Donohue et al., 2010; Peirera, 201
). Kilroy (2001),
using national

data from 1995 to 1997, found a median value of 17 μg P

of unfiltered molybdate
reactive P in groundwaters, but that one quarter of the data were higher than 30 μg MRP l

, 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 la
kes was modelled from their residence time (

PLake = 1.55 [Pinput/1+√

from the equation originally formulated by OECD (1982), and used widely in lake
(Note a modification of the OECD equation estimated by Foy (1982) for Irish lakes, and used by Irvine

et al., 2000, provides slightly higher estimates of in
lake TP). The model was run using both the high
good (median of 25 µg l

of MRP) an
d good
moderate boundary (35 µg l


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
(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 an
d 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, e
stimated 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
ion 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).

Fig 2

Modelled mean concentrations of TP (µg l
) 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

(diamonds, lower)
and 35 µg l

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 t
han are in the current Regulations S.I.272 of 2009.
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


(Donnelly, 2001). There is


a strong need to reassess the
rationale for P mang
ment entering high quality lakes.
A similar mismatch may apply to
nutrient loads
Tw (years)
TP in lake




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., 199
8; Jennings et al., 2002; Bartley and Johnston, 2005), with an association

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

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 qualit
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

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 an
d diffuse loads.

Under the WFD
, m
any small water bodies (lakes


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 da
ta (Ní Chatháin, 2006).

Owing to current lack of method development


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
rinciples of
nutrient parsimony, naturalness, food web structure, hydrological connectivity and
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

elements. The current protocols developed for the ecological assessment is not, therefore, the same
as a focus on ecol
ogy, 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
requirement in order to identify impact of, and mitigation from, pressures. Addit