24
cultural management have been based on sector-specifi c laws, policies and
institutions (e.g. Sielke and Dreyer in 2015 ; Mee 2005 ; Mee et al. 2008 ), and their
relationship to other sectors has seldom been considered.
However, eutrophication is a complex phenomenon. Apart from the intricate
array of primary and secondary ecological impacts shown in Fig. 2.1 , nutrient
sources are diverse and stem from numerous natural and anthropogenic sources in
several sectors, both in the drainage basin in question and on a wider international
scale through atmospheric and riverine transport (Lundberg 2005 , 2014 ). The relationships between these socio-economic pressures and the marine ecological state
are usually non-linear (Mee 2005 ) and prone to quick and fundamental shifts when
thresholds are passed (e.g. Österblom et al. 2010 ). Furthermore, it may take a long
time before any reductions in nutrient input from land may allow for an improved
situation (Elofsson 2010 ). The degradation of organic matter can, for example, be
inhibited by negative effects of existing hypoxia on infauna (sediment living animals) (Conley et al. 2007 ) or by so-called internal loading if buried phosphorus
leaks from anoxic sediments (Vahtera et al. 2007 ; Zillén et al. 2008 ). In addition,
other major disturbances, such as overfi shing , and introduction of invasive species
may also infl uence the recovery of an ecosystem from eutrophication. Variability in
climate-related factors, such as storms and water temperature and stratifi cation, may
additionally cause unexpected responses. The interwoven links between marine
eutrophication and all these natural as well as human-induced, biotic as well as
abiotic, processes, systems and feedback mechanisms (Caddy 1993 ; Cloern 2001 ;
McQuatters-Gollop et al. 2009 ) make science-based reductionist governance
models highly insuffi cient. As Elliot ( 2002 ) has pointed out, the management of
marine ecosystems needs to consider this full complexity .
In response to these shortcomings, the ecosystem approach to management
( EAM ) has emerged during the last decades as a central component of environmental governance (Atkins et al. 2011 ; Curtin and Prellezo 2010 ; Trush and Dayton
2010 ). In the words of Browman and Stergiou ( 2004 ), EAM is based on the insight
that the whole complex ecosystem (including its capacity to deliver important
ecosystem services ) is greater than the sum of its parts. Problems and risks need to
be managed in a holistic manner, not independent of each other (Hammer 2015 ),
and both ecological and social dimensions need to be considered. Moreover, EAM
takes into account existing knowledge as well as uncertainties and other forms of
complexity . Besides for being different because it is based on a multiple factor
approach, EAM also differs from traditional management in its application to a
specifi c geographic scale (Curtin and Prellezo 2010 ).
EAM is defi ned in the Convention for Biological Diversity (CBD 1998 , 2004 ) as
well as in conventions on regional seas, for example, those concerning the
governance of the North Sea and the Baltic Sea (HELCOM and OSPAR 2003 ). It is
also included in several policies and laws, for example, the EU Water Framework
Directive (WFD) (EC 2000 ), the Marine Strategy Framework Directive ( MSFD )
(EC 2008 ) and the Helsinki Commission’s (HELCOM) Baltic Sea Action Plan
( BSAP ) (HELCOM 2007a ).
M. Karlsson et al.
cultural management have been based on sector-specifi c laws, policies and
institutions (e.g. Sielke and Dreyer in 2015 ; Mee 2005 ; Mee et al. 2008 ), and their
relationship to other sectors has seldom been considered.
However, eutrophication is a complex phenomenon. Apart from the intricate
array of primary and secondary ecological impacts shown in Fig. 2.1 , nutrient
sources are diverse and stem from numerous natural and anthropogenic sources in
several sectors, both in the drainage basin in question and on a wider international
scale through atmospheric and riverine transport (Lundberg 2005 , 2014 ). The relationships between these socio-economic pressures and the marine ecological state
are usually non-linear (Mee 2005 ) and prone to quick and fundamental shifts when
thresholds are passed (e.g. Österblom et al. 2010 ). Furthermore, it may take a long
time before any reductions in nutrient input from land may allow for an improved
situation (Elofsson 2010 ). The degradation of organic matter can, for example, be
inhibited by negative effects of existing hypoxia on infauna (sediment living animals) (Conley et al. 2007 ) or by so-called internal loading if buried phosphorus
leaks from anoxic sediments (Vahtera et al. 2007 ; Zillén et al. 2008 ). In addition,
other major disturbances, such as overfi shing , and introduction of invasive species
may also infl uence the recovery of an ecosystem from eutrophication. Variability in
climate-related factors, such as storms and water temperature and stratifi cation, may
additionally cause unexpected responses. The interwoven links between marine
eutrophication and all these natural as well as human-induced, biotic as well as
abiotic, processes, systems and feedback mechanisms (Caddy 1993 ; Cloern 2001 ;
McQuatters-Gollop et al. 2009 ) make science-based reductionist governance
models highly insuffi cient. As Elliot ( 2002 ) has pointed out, the management of
marine ecosystems needs to consider this full complexity .
In response to these shortcomings, the ecosystem approach to management
( EAM ) has emerged during the last decades as a central component of environmental governance (Atkins et al. 2011 ; Curtin and Prellezo 2010 ; Trush and Dayton
2010 ). In the words of Browman and Stergiou ( 2004 ), EAM is based on the insight
that the whole complex ecosystem (including its capacity to deliver important
ecosystem services ) is greater than the sum of its parts. Problems and risks need to
be managed in a holistic manner, not independent of each other (Hammer 2015 ),
and both ecological and social dimensions need to be considered. Moreover, EAM
takes into account existing knowledge as well as uncertainties and other forms of
complexity . Besides for being different because it is based on a multiple factor
approach, EAM also differs from traditional management in its application to a
specifi c geographic scale (Curtin and Prellezo 2010 ).
EAM is defi ned in the Convention for Biological Diversity (CBD 1998 , 2004 ) as
well as in conventions on regional seas, for example, those concerning the
governance of the North Sea and the Baltic Sea (HELCOM and OSPAR 2003 ). It is
also included in several policies and laws, for example, the EU Water Framework
Directive (WFD) (EC 2000 ), the Marine Strategy Framework Directive ( MSFD )
(EC 2008 ) and the Helsinki Commission’s (HELCOM) Baltic Sea Action Plan
( BSAP ) (HELCOM 2007a ).
M. Karlsson et al.