ICES CM 2010/P:13
Not to be cited without prior reference to the author.
Reversing the burden of proof in fisheries management in the context of
an integrated ecosystem approach.
Fitzpatrick, M.1, Graham, N.2, Rihan, D.J.3 , Reid, D.G.2, Sutton, G.1 and MacMullen,
P.4.
1. Coastal & Marine Resources Centre, University College Cork, Glucksman Marine
Facility, Naval Base, Haulbowline, Cobh, Cork, Ireland.
2. Marine Institute, Rinville, Oranmore, Co. Galway, Ireland.
3. Bord Iascaigh Mhara, PO Box 12, Crofton, Road, Dun Laoghaire, Co. Dublin, Ireland.
4. Sea Fish Industry Authority, Origin Way, Europarc, Grimsby, N E Lincs, DN37 9TZ, United
Kingdom.
The implementation of an ecosystem-based approach to the management of marine
resources will create additional layers of uncertainty and complexity which may be
incompatible with current management approaches. Shifting to a results-based
management approach while also derogating greater responsibility in decisionmaking will require a reversal of the burden of proof. Since an ecosystem approach
implicitly requires integration across sectors and areas, it follows that sectorally and
spatially integrated approaches to reversing the burden of proof will also be required.
This paper explores how reversing the burden of proof has been approached in some
fisheries globally and also in the management and regulation of another marine
resource sector - aggregate extraction. Comparisons are made to the current
approach in European fisheries which operate under a system where the burden of
proof rests largely with government institutions.
Through examples, we show how multi-sector and societal demands can result in
marine use conflict and increasingly how there is a burden of proof requirement to
provide information which demonstrates the legitimacy of an activity. We explore
how some sectors of the fishing industry have responded and how the provision of
higher resolution data has provided evidence revealing that perceived conflicts may
in practice be unfounded.
Finally we examine the implications of an integrated ecosystem approach for results
based and devolved or self-management as possible outcomes of reform of the
Common Fisheries Policy.
1. Introduction
Implementation of an ecosystem approach to marine management will produce more
complex monitoring requirements and give rise to greater uncertainty than has been
the case in previous single sector approaches (Cochrane 1999; Degnbol 2002; Garcia
and Cochrane 2005). The micromanagement scenario evident in current European
fisheries management is likely to be even less effective in the context of an
integrated cross-sectoral management regime as envisaged by the Integrated
Maritime Policy and the Marine Strategy Framework Directive (Parliament 2008).
Results Based Management (RBM) or self-management approaches have been cited
as being more robust and resilient in the face of uncertainty than traditional topdown management regimes (Charles 2002). A critical aspect of RBM approaches is
the Reversal of the Burden of Proof (RBP), which involves industry proving that its
activities are not a risk in relation to resource sustainability. It can be can be seen as
a necessary trade-off in return for an enhanced role in management and a reprieve
from top-down micro-management.
The Marine Strategy Framework Directive (Parliament 2008) will require EU Member
States to collaborate in the development of regional strategies to deliver Good
Environmental Status by 2020. At the same time a key element of the Integrated
Maritime Policy is the streamlining of management and economic development of
different industrial sectors.
Reversal of the Burden of Proof (RBP) in a fisheries management context is linked to
the broader issue of Results Based Management (RBM) or self-management. RBM in
turn can be considered to be part of the broad suite of ecosystem approaches in that
it is in part aimed at dealing with broader ecosystem attributes beyond single target
stock considerations.
The debate about where the burden of proof should be placed is not new and as far
back as 1919 W.F. Thompson wrote, “Proof that seeks to change the way of
commerce… must be overwhelming” (quoted in Fluharty et al, 1998). Before the
widespread application of the precautionary approach the burden of proof in the
majority of fisheries lay with managers who were required to demonstrate that
fishing levels were demonstrably creating a resource problem. The desire of
fisheries managers among others to reverse the burden of proof has been expressed
in a number of papers by Sissenwine 1987, Mangel et al. 1996, Dayton 1998,
(Gerrodette, Dayton et al. 2002). The precautionary approach may have shifted the
burden of proof somewhat and a commonly expressed industry feeling now is that it
places a burden on them to provide evidence that in any case is often rejected
(Charles 2002). However a full reversal of the burden of proof, similar to that which
exists in the medical and pharmaceutical industries whereby no activity can take
place until the prospective practitioner demonstrates that defined environmental
limits will not be breached, has not occurred to a wide degree in fisheries
management to date.
The current debate in European fisheries management about RBP can be linked to
the 2009 Green Paper on Reform of the Common Fisheries Policy (European
Commission, 2009) that proposed co-management or self-management
arrangements as a potential solution to the problem of increasingly complex and
costly micro-management of fisheries. The fishing industry has in turn responded
positively to this proposal in submissions made to the CFP reform consultation
process (European Commission, 2010). The possibility of industry groups developing
fishing plans with periodic audits have been proposed by a number of industry
representative bodies. However the proposals to date have not contained sufficient
or indeed any detail on what constitutes proof or how audits could verify compliance.
Clearly a more thorough debate is required on the details of how this process could
work.
2. Current global approaches to RBP in fisheries management.
2.1 Spencer Gulf Prawn Fishery, Australia.
In Australia’s Spencer Gulf Prawn fishery 39 demersal trawlers are licensed to fish.
The fisheries success in maintaining good catch rates has been ascribed to the
decision to limit entry from the onset of the fishery in 1968 and also to the
collaborative co-management arrangement that exists between the fishermen’s
association, government managers and scientists (Townsend et al 2008). Although
the Minister retains ultimate control management of the fishery has been delegated
to a Fisheries Management Committee since 1995. The management plan for the
fishery (Government of South Australia, 2007) lists a number of goals, objectives and
strategies (with associated indicators and limit reference points) across a wide range
of factors including fisheries ecosystem impacts. The success of impact mitigation
strategies are assessed through stock assessment surveys which are conducted on
industry vessels three times yearly, risk assessments, dedicated fishery-independent
bycatch surveys and observer programmes. The Committee at Sea, made up of
skippers and licence owners are also required to report regularly on compliance with
harvest control rules.
Under a new Fisheries Amendment Bill, which is currently being debated in Australian
parliament, there will be further delegation of management responsibility to fishing
industry groups. These changes will strengthen the ability of industry groups to
commission and design their own research programs and refocus the Australian
Fisheries Management Agencies role on its auditing function.
2.2 Fully Documented Fishery trial, Denmark.
A recent year long trial project on board Danish fishing vessels is a good example of
how a comprehensive approach to reversing the burden of proof may be taken.
Based on the assumption that total catches (landings + discards) would be deducted
from fishing quotas in the future vessels could have the opportunity to carry cameras
onboard to document discard levels and in return would receive improved quotas
(Dalskov and Kindt-Larsen, 2009). The feasibility of this approach was tested and
the research found that discard and catch levels could be determined accurately by
viewing the electronic monitoring records onshore. The survey also found that the
cost of verifying discards in this way was significantly lower than through the use of
onboard observers and that fishermen involved were incentivised to actively avoid
areas of high discarding, particularly for juvenile Cod.
2.3 Pacific Coast Groundfish, Canada.
On the Canadian Pacific coast the fishery for Sablefish is co-managed using an ITQ
system by the Department of Fisheries and Oceans and the Canadian Sablefish
Association. This co-management arrangement is formalised through a legally
binding Joint project Agreement which requires the fishermen to fund and organise
research, management, monitoring and enforcement.
Due to concerns over the stock status of a number of species a pilot Commercial
Groundfish Integration Plan (CGIPP) was initiated in 2006 based on a management
plan that has a number of guiding principles that are relevant to this paper (DFO,
2009). These are that all groundfish catches, across a complex mix of 60 species of
which sablefish is one, must be accounted for. Also the plan stated that both at sea
and dockside monitoring arrangements must be revised which resulted in a 100% atsea and dockside monitoring programme which is jointly paid for by DFO and
industry. The at-sea programme allows vessels to choose between carrying an
observer or a video monitoring system. This is backed up with a dockside
monitoring programme which combines video footage of all landings of which
approximately 10% are checked against vessel declarations with random inspections.
Discrepancies between the declared catch and the dockside video footage can trigger
100% dockside monitoring the cost of which is charged to the fisherman.
Following an evaluation of the pilot plan it was agreed to extend the arrangement
and 2010/2011 is the first year of a more permanent integrated plan (DFO, 2010).
Although the pilot plan evaluation indicated that conservation objectives were being
achieved there are some concerns about the sustainability of funding for what is an
expensive monitoring programme.
3. Current approach to Reversing the Burden of Proof in Marine
Aggregates (MA’s).
The bulk of countries in the North Atlantic areas that are members of ICES where
marine aggregates are extracted from the seabed (this includes all the major
extractors in terms of volume) have regulatory regimes in place to manage these
activities that are based on the ICES guidelines for the management of marine
sediment extraction (ICES, 2003; ICES 2009). These guidelines cover issues such as:
• The necessity of a strong regulatory and strategic framework to manage
applications and resolve spatial conflicts
• Balancing conservation with demand
• Minimising adverse environmental effects and use of an ecosystem approach
• Account for interests such as fisheries
A key recommendation contained in the guidelines is for an environmental impact
assessment which should include information on physical impact assessment,
biological impact assessment, interference with other legitimate uses of the sea,
evaluation of impacts, mitigation measures, authorisation issues, and monitoring
compliance with the conditions attached to an authorisation. Screening is usually
carried out as the first step in the planning process in order to determine if the
proposed activities, due to scale or other circumstances, are exempt from the
environmental assessment process. In Ireland, as set out in the European
Communities (Environmental Impact Assessment) Regulations, 1999 (S.I. No 93 of
1999) an Environmental Impact Statement must be provided in cases involving:
“Extraction of stone, gravel, sand or clay by marine dredging (other than
maintenance dredging) where the area involved would be greater than 5 hectares”.
In certain cases, the Minister may require a sub-threshold EIS to be prepared.
In relation to monitoring compliance with authorisations for marine aggregate
extraction, although not mandatory, the use of black box monitoring systems
onboard aggregate dredging vessels is now common practice amongst those ICES
Member Countries who are the principal producers of marine aggregate, including
Belgium, the Netherlands, Spain, Germany and the UK. For example the UK has
more than 10 years of detailed records of dredging activity covering more than
300,000 hours and 30 million individual dredging records. In Denmark, dredging
activities have been reported in detail since 1990. This information in turn has
provided unparalleled levels of information on the scale, extent and intensity of
dredging operations, providing benefits to both regulators and operators.
In England, the licensing process is governed by a set of regulations which are
accompanied by procedural guidance in “Marine Minerals Guidance Note 2” which
supplement the existing “Marine Minerals Guidance Note 1”. These documents
contain procedural guidance explaining the application process for marine minerals
extraction in British waters together with guidance on environmental assessment,
mitigation and monitoring criteria, based in part on the 2003 ICES WGEXT
Guidelines.
In the UK, the advent of electronic monitoring data, and in particular the annual
summaries of activity, has allowed the industry and landlord (The Crown Estate) to
produce annual reports detailing the area of seabed licensed and dredged. Analysis
of electronic monitoring data allows the annual extent and intensity of dredging
activity based on dredging hours recorded in individual 50 m x 50 m grid cells (fig.
3.1.1).
Figure 3.1.1 UK analysis of dredging activity from Electronic Monitoring
System data (ICES, 2010).
This information in turn has become a guide to the industry’s overall environmental
performance. While information on the extent of dredging activities is reported
annually, by combining this information, it is possible to consider the cumulative
footprint (the total extent of dredging activity) over a period of time. The UK has
begun this exercise and has found that over a 5 year period the total area dredged
by the marine aggregate industry totalled 380 km2, compared to annual totals
ranging from 220 to 149 km2. This information is particularly relevant to research
and monitoring when attempting to relate observed environmental impact or
recovery to the timing and intensity of actual dredging operations.
However, the move towards more spatially restricted and therefore more intensive
levels of dredging activity does raise an additional issue which requires investigation.
While the total spatial footprint of the impact is reduced, the increased levels of
intensity can affect the timescale for the recovery of the environment. The
availability of detailed black box data to assess historical dredging activity will allow
this issue to be examined in more detail.
The evolution of management of marine aggregate dredging activities has seen
some significant advances over the past 10 years. While the pressures of
environmental regulation and control have continued to increase – particularly as a
result of the European Environmental Impact Assessment and Habitats Directives,
some of the greatest changes in management and control of dredging operations
have come from the industry themselves. These are not only linked to improving
resource management, but also reducing spatial conflicts with other marine users.
This has obvious links to the development of wider marine spatial planning initiatives.
(See http://www.bmapa.org/issues_other01.php for industry codes of practice re
minimising areas involved, archaeological heritage, sustainablility etc.)
4. Case Study: Response of inshore UK fishermen to RBP requirement
In England fishermen are rapidly having to learn to live with new marine
management arrangements resulting from the Natura 2000 network and the
provisions of the two UK Marine Acts. Both of these involve the introduction of
marine protected areas (MPAs). The extent to which different fishing operations may
be excluded or restricted will depend upon the evidence base available – relating to
habitat sensitivity and the known or perceived impacts of fishing – the degree to
which zoning and buffering can be negotiated between the various parties involved
and the level of compliance that can be demonstrated consistently by the catching
sector.
Lyme Bay, on the south coast of England, contains good scallop (Pectin maximus)
grounds as well as sandstone reef habitats that contain species such as the bryozoan
Ross corals (Pentapora foliacea) and sea fans (eg Eunicella verrucosa). These
features have sustained some damage over the years from beam trawling, otter
trawling and scallop dredging but local sources claim that improved
navigation/position fixing have improved the situation quite substantially (J. Portus,
pers comm). When features in this area were not designated in the first tranche of
N2k sites local conservation organisations launched a very high profile anti-fishing
campaign and demanded full N2k protection.
In parallel with this initiative fishermen in SW England formed two new associations,
the South West Inshore Fishermen’s Association and the Channel and West
Sustainable Trawling Group. These had the aims (inter alia) of agreeing codes of
good operating practice (ie respecting conservation interests), demonstrating
compliance with local access arrangements such as were being set up in Lyme Bay
and applying for MSC certification. A zoning exercise was carried out through
discussions between the industry and statutory conservation advisors, and closed
areas were defined and agreed. Vessels working in Lyme Bay committed to
compliance and merchants buying from the area agreed to buy only from compliant
vessels.
Funding was accessed from Seafish and an inshore VMS project was initiated. The
aims of this work included looking at a vessel monitoring system for scallop vessels
working in Lyme Bay in order to verify that they were in compliance with the access
agreements. Unfortunately, while this monitoring project was being set up the
fisheries department imposed an emergency closure on a 60 square mile (~155 sq
km) area of Lyme Bay.
Despite this setback the industry side remains committed to the VMS pilot. One very
strong, underlying reason is that the MPA designation process in the UK is subject to
an extremely tight timescale; about two years to completion. The statutory
conservation advisors and putative site managers are hard pressed to map fishing
activity comprehensively and appropriately, conduct sensible impact and risk
assessments, negotiate with the industry, and produce consensus management
plans, all in the time available. This puts the onus on industry to demonstrate that it
can operate responsibly, contribute to marine environmental conservation and be an
effective steward of the environment.
The inshore VMS system features potentially high ‘pinging’ frequency (4 seconds
maximum), ‘geofencing’ of sensitive areas, and the identification of individual units of
fishing gear through RFID tags read by a component within the VMS unit. Their
deployment should permit protection boundaries to be drawn much more tightly
around seabed features, biotopes, etc, than has been the case to date. Fishing gear
types can be identified so that access may be allowed preferentially for low impact
gears, and effort levels can be monitored.
Such initiatives by industry should do much to counter what often seems to be an
irrational rejection of all fishing activities, allow management plans to be agreed that
are sensitive to local operating practices, and help to defuse the problems attendant
upon the MPAs being introduced within a very short timescale.
5. Cod Long Term Management Plan
The practical application of burden of proof until recently, has not been applied in the
context of European Fisheries. However, new EC regulations aimed at reducing
fishing mortality on cod has introduced the concept into the regulatory framework.
Many cod stocks exploited by EU fisheries are well outside desirable biomass levels
and exploitation rates and have been for the past decade at least. Despite earlier
attempts and the introduction of a management plan which aimed to rebuild stocks
by 30% annually (EC regulation 423/2004), many stocks have continued to decline.
More recently, a new Long Term Management Plan for Cod (EC regulation
1342/2008) has been introduced and this has shifted the focus from rebuilding
biomass to setting a target fishing mortality. Incorporated within the plan are annual
reductions in both TAC and effort, the extent of which are dependant on stock status
and fishing mortality rate. For cod stocks in VIa and VIIa, this has meant an annual
reduction of 25% in both TAC and effort allocation. However, the plan offers
individual countries to implement alternative measures to the effort reductions
provided they can demonstrate that equivalent reductions in cod catches are being
achieved (article 13) or in cases where annual catches at an individual vessel level
do not exceed 1.5% of the total catch, then these vessels can apply to be exempted
from the effort control element of the regulation. It is important to emphasise that the
regulation is based on catches and not declared landings, and as such it is
necessary that the catch rates are adequately monitored either by onboard observers
or remote sensing equipment i.e. video cameras. These elements of the cod plan
have stimulated individual member states and fishermen to adopt measures that
minimise cod catches through technical modifications to gears or the application of
closed areas. In either case, individual countries must now supply adequate data on
an annual basis that demonstrates that cod catches are below specified levels and
how these levels are being achieved if they are to avoid effort reductions or in the
case of vessels that are exempted, to remain outside the effort control scheme. This
approach is in sharp contrast to previous management plans whereby, the onus is
now on the individual country or fishermen to demonstrate that adequate measures
are in place in order to avoid a default reduction in fishing opportunities.
There are two important distinctions between article 11 and article 13 in terms of
burden of proof and the data and analytical requirements necessary. Under article
11, if a group of vessels can demonstrate that cod catches are below 1.5% on an
annual basis then they may apply to be excepted from the regulation. A number of
EC member states (Sweden, Ireland, UK, France and Spain) have had some metiers
excepted, through the provision of adequate catch data from at sea observers. The
data required is easily collected from at sea observers and if achieved through
technical means (technical decoupling) is justifiable in the sense that a positive action
has been taken to avoid catching cod in the first instance. One flaw with the
regulation is that if cod catches are below the desired level (1.5%) without technical
modifications, it is difficult to ascertain whether the cod catches are low because the
fishing activity is being conducted in an area outside the normal distribution of cod
(spatial decoupling) or because of the depleted nature of the stock (depletion
decoupling).
Burden of proof is more problematic under article 13(c) which allows member states
to introduce cod avoidance or discard reduction plans which reduces mortality of cod
by at least as much as the effort reductions that would be introduced in the absence
of such a plan. It is simply not sufficient to contrast one years catch data with the
years prior to implementation due to the dynamic nature of the stocks, but rather to
quantify the partial fishing mortality associated with the participating vessels in the
year prior to and during the period of the plan. In theory, this is of course is possible
provided that there is a robust and reasonably precise full analytical assessment of
the cod stock in question, this is not the case certainly for cod stocks in VIa and VIIa.
Alternative approaches, such as contrasting CPUE levels between participating and
non-participating vessels is used and spatially linked through VMS. What is clear
however, the data requirements to assess the proportional reduction in mortality is
administratively difficult and resource demanding. If cod mortality levels fail to reduce
quickly enough and effort allocations continue to diminish, it is likely that more
member states will seek alternative methods to reduce cod catches increasing the
burden of proof requirements. Demonstrating that a particular action has resulted in
a quantifiable reduction in fishing mortality, is very difficult from a scientific
perspective.
6. Possible impacts indicator for use across sectors
In the 2010 Quality Status Report (http://www.ospar.org) OSPAR identified eight
major categories of human activities that may impact on the marine environment in
general. These include; mariculture, shipping, tourism, wind and wave power,
cables, mineral extraction, and dumping, as well as fishing. In terms of geographical
extent and definition one could divide these into two categories. The first would be
where the activity and it’s associated pressures were local and specific in area, e.g.
mariculture, wind and wave power, cables, mineral extraction, and possibly dumping.
The second category would be where the activity is much more widespread and
indeterminate in area, e.g. shipping, fishing, and possibly tourism.
For the first category, there is usually either an installation or a specified
extraction/dumping site and these would usually have a small and well defined
footprint. Evaluation of the ecological impact of such activities is fairly well developed
(Patin 1999; CEFAS 2004; OSPAR 2008). Ecological impact assessment would be
expected to occur and would be relatively straightforward. Specific indicators,
monitoring, and the burden of proof could be established and would usually be
expected to lie with the operator.
For the second category, the issues are more complex. In the case of fisheries, the
spatial footprint, and it’s definition would be much less easy to define. In Europe,
fishing would be expected to take place over much of the continental shelf, and as it
exploits a mobile resource, the area of operations would be very wide and to some
extent unpredictable, unlike the immobile resources exploited by oil, gas, or gravel
extraction operations. The impacts of fisheries are also more difficult to evaluate,
and hence to determine appropriate indicators, monitoring and appropriate remedial
action. This also complicates the issues of burden-of-proof. This is probably also true
for shipping and possibly tourism, where the location of the activity and potential
impacts would be dispersed and variable.
On this basis alone, it is arguable that the scope for common indicators across
sectors is minimal. Another element to consider is the objectives that we would have
for fisheries in contrast to the other sectors, and the problems of developing relevant
indicators for these, even without attempting a cross-sector approach.
The principle environmental impacts from fishing were identified in the OSPAR QSR
process as being the removal of species (commercial and other), and habitat damage
and loss. A similar evaluation was made for Australian fisheries (Pascoe, Proctor et
al. 2009), where the key environmental pressures were bycatch and habitat damage.
Sustainability of commercial stocks was identified separately. Pascoe et al also
defined an objective to “minimize environmental impacts”. Essentially, these
pressures can be summarized as; removal of commercial fish and shellfish; removal
of non-commercial fish and shellfish (discards and bycatch); and habitat damage and
loss. The objectives, possible indicators, and cross sectoral relevance of these are
considered individually below.
6.1 Removal of commercial fish and shellfish
In it’s current form, within the EU, fisheries management focuses on the removal of
species, and commercial species in particular. The current TAC based management
approach used in Europe, can be seen to address the “removal of commercial fish
and shellfish” aspect. The objective would be to have sustainably fished stocks. The
indicators are likely to remain as biomass (B) and fishing mortality (F), based on the
precautionary approach or a sustainable yield approach (MSY). These would be exsitu indicators. Essentially, we currently have a results based management approach,
if the indicator passes a certain threshold, we reduce or increase TAC accordingly.
The scope for results based management and co-management lies in the way the
TACs might be taken, the fishing plans (Lassen et al 2008). In situ indicators could
be based on on-board monitoring of catch and catch rates. The reversal of burden of
proof would require a monitoring and auditing scheme that was robust to
manipulation. Both B and F indicators, and any on-board indicators can essentially be
seen as intrinsic to fishing, and would have no real relevance for any other sectors.
One exception would be tourist recreational fishing, but this is best considered as
another fishing métier.
6.2 Removal of non-commercial fish and shellfish
Essentially this pressure describes those organisms that are caught in the fishing
process and then discarded. Anything landed would be assumed to have commercial
value. Discards of undersized or over quota commercial fish would be included
above. For a few species it may be possible to estimate B and F. The problem
though with these indicators is that they are also subject to a wide range of other
factors, e.g. recruitment, growth, natural mortality (including predation) and disease.
So fishing would not be the only factor causing change in B and F. In any case B & F
are unlikely to be known for most non-commercial species. Again, it seems likely that
only in situ indicators, e.g. on vessel discards rates and weights, would be
appropriate for fishery management. Ex-situ indicators such as fish community
structure and biodiversity are, like B & F, subject to many other natural processes as
well as fishing, and additionally may take many years to respond to changes in
fishing pressure. The burden of proof for in situ indicators would again lie with
robust and accepted on board monitoring of discarding as well as landed catches.
Again, these indicators would have little or no cross sector relevance.
6.3 Habitat damage and loss
Pascoe et al (2009) showed that habitat damage had a broadly similar weighting to
bycatch across a wide range of stakeholder perceptions. The key problem for the
development of objectives, indicators and management lies in the difficulty in
quantifying the impact of fishing on seabed habitats. There have been many local,
and usually short duration, studies quantifying the impact of fishing, particularly
bottom fishing on habitats (Kaiser et al 2006). Based on these studies, it would be
theoretically feasible to develop a range of habitat indicators that could be monitored
to evaluate fishing impacts e.g. on functional epibenthos groupings (de Juan et al
2009). There are several difficulties in using these as indicators for managing fishing
activity. Firstly, as with B & F, there will be other factors that can change the
indicator levels. Secondly, there may, again, be a considerable delay between the
pressure and the indicator response. But, most critically, it would almost certainly be
prohibitively expensive to routinely monitor such indicators reliably, say using
research vessels. It may be more feasible to quantify habit loss, but by it’s nature, it
is difficult to manage to protect something once it has been permanently lots!
An alternative solution would be to use the spatial and temporal pattern of fishing
activity as our indicator of habitat damage. If our main objective is to reduce and
minimize habitat damage, we may not actually need to know the absolute state of
these habitats. We could develop an approach that manages fishing activity to
reduce the pressure, and assume that this would translate to reducing impact.
Fishing effort data is available via VMS (Lee et al 2010; Gerritsen & Lordan in press),
using simple speed windows to determine if a vessel is fishing. Model based
approaches linking fishing activity to the benthos (e.g Hiddink et al 2006) could then
allow us to calculate the impact of fishing activity on benthos biomass and
production. This could also include assessments of sensitivity of particular habitat
types to fishing activity (Hiddink et al 2007). Potentially this could also be partitioned
according to gear, say ranging from dredges, through beam trawls and otter trawls,
to seine and passive gears.
The core indicator for this approach would be VMS based fishing activity in space and
time, and partitioned by gear etc. This indicator would be relatively simple to use for
management purposes. It is collected remotely and analysis approaches are already
available (Lee et al 2010; Gerritsen & Lordan in press). It is also available
immediately, with no time delay. Linking this to the type of models developed by
Hiddink et al (2006,2007) would allow us to interpret that fishing activity in terms of
impacts on various types of habitat, and to develop management measures to
achieve particular objectives, e.g. reduction in impact. There would be considerable
scope for co-management, as there would often be a range of possible routes to a
particular objective.
To illustrate, we could imagine we have two habitat types (A & B), A is more
sensitive than B (i.e. more easily damaged and slower to recover). We could have
two gear types (X&Y), X is more likely to cause benthic damage than Y (e.g. a beam
trawl and a Danish seine). Effort is currently evenly split between gear and habitat. If
our objective is to reduce habitat damage we would have many possible
management options; shift activity from Habitat A to B, shift from Gear X to Gear Y,
use only Gear Y in Habitat A, reduce overall activity or some combination of all of
these. We could then use the VMS activity and our models to predict what each
option might yield in terms of our objectives.
The approach offers considerable options for self-management, or at least, industry
led approaches to achieve mutually agreed objectives. The objectives could be
expressed in terms of, say, benthic biomass and production, which would allow direct
comparison to other pressures on marine habitats, e.g. wind farms or shipping. But
they could be managed in terms of spatial and temporal allocation of fishing effort
i.e. VMS records. The burden of proof would be for the industry to develop fishing
plans (space, time, gear etc.) that could deliver the objective, and they could
demonstrate compliance using VMS. Compliance could also be demonstrated at an
individual vessel level.
The key to this approach probably lies in all parties’ willingness to trust the models.
But even this may not be essential, as long as they were willing to agree on
interpretations. For instance, it would probably be relatively easy to agree that
deploying beam trawls in lophelia beds is probably more environmentally damaging
than crab pots on sand. The models and other empirical studies might be used to
develop a mutually agreed impact level for a matrix of gears Vs. habitats, and then
use this to manage spatial allocation of effort.
7. Conclusions
• A coherent approach to multi-sector integrated marine management will have
to encompass approaches to reversing the burden of proof.
• There are currently major spatial and sectoral differences in how RBOP is
handled. These differences have the potential, notwithstanding the
implementation of integrated marine management under the umbrella of the
MSFD and the IMP, to create inter- and intra-sectoral conflicts. Differences in
the approach to RBP are manifest between fisheries and marine aggregate
extraction for instance at the earliest stages of exploitation of the resource.
For marine aggregate extraction it is generally the responsibility of the
developer to undertake a succession of surveys which are designed to
identify, and demonstrate the viability of specific resource areas. Even at this
preliminary stage activity is normally restricted to the terms of exploration
licenses, the granting of which in many countries requires the developer to
have undertaken some form of assessment of the environmental impact of
the prospecting activities.
• Fisheries activities, by virtue of their existence prior to the advent of
environmental risk assessment type legislation, such as the Environmental
Impact Assessment Directive (85/337/EEC), have occupied a privileged
position relative to some other marine industry sectors in relation to RBP
approaches.
• The development of cross-sectoral indicators and cumulative impact
assessments has the potential to address the vacuum in terms of prelicensing impact assessment.
• A range of legislative and market-led drivers are in any case requiring
fisheries to conduct a posteriori risk assessments and to assume the burden
of proof to varying degrees.
• For the fisheries sector, the tensions which could be produced as a result of
inconsistent approaches to RBP, have the potential for a positive outcome in
accelerating the adoption of self-management or results-based management
approaches.
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