PMNQ%20Deliverable%201%20scientific%20principles

A Review of The scientific principles

underlying impact of aquaculture on the

environment Deliverable 1 for PHILMINAQ

Mitigating impact from aquaculture in the Philippines Project number: FP6-2004-INCO-DEV-SSA- 031640 6th Framework Programme

Duration: 08/2006 to 02/2008 (18 months) Specific Support Action

The scientific principles underlying impact of aquaculture on the environment

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Table of contents 1 Abstract ..................................................................................................................3 2 Environmental impacts from Aquaculture .............................................................3

2.1 Environmental awareness and sustainability .................................................3 2.2 Types of waste and effluent deposition..........................................................4

3 Environmental impacts...........................................................................................5 3.1 Soluble wastes ............................................................................................5

3.2 Solid wastes....................................................................................................9 3.2.1 Organic matter......................................................................................12

3.3 Chemicals and other additives .....................................................................13 3.4 Effects on wild fish populations...................................................................14

3.4.1 Contamination with waste....................................................................14 3.4.2 Parasites and diseases...........................................................................15 3.4.3 Escapees and genetic transfer...............................................................15

3.5 Direct impacts on wildlife ............................................................................16 Implications for biodiversity ................................................................................16

4 Scientific rationale behind monitoring methods ..................................................17 4.1 Hydrographic/ water quality analyses and other assessments......................17 4.2 Chemical analyses ........................................................................................18 4.3 Benthic faunal analyses................................................................................18

5 Aquaculture Regulation .......................................................................................19 6 Conclusions ..........................................................................................................23


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1 Abstract As aquaculture expands, regulation to prevent environmental damage is an essential requirement for sustainability. In this review, the scientific principles underlying impact on the environment, monitoring programmes and aquaculture regulation pertaining to 1) protection of other resource users, 2) protection of ecosystem structure (conservation) and 3) protection of ecosystem function (recycling) are discussed. Some of the approaches taken to regulation of aquaculture in several countries are presented emphasising the need for these to be based firmly in a good scientific understanding of the ecosystem and the processes by which it interacts with aquaculture.

2 Environmental impacts from Aquaculture The aquaculture industry has experienced considerable expansion over the past decades, with attention being paid subsequently to the environmental effects of its activities. The potential impacts of aquaculture are wide-ranging, from æsthetic aspects to direct pollution problems. Marine aquaculture operations and their associated infrastructure can disturb scenic rural areas. Fish production may generate considerable amounts of effluent (e.g. waste feed and faeces, together with associated by-products such as medication and pesticides) which can have unacceptable impacts on the environment. There may also be some unwanted effects on wild fish populations, such as genetic disturbance and disease transfer by escapees or ingestion of contaminated waste. Direct outputs from aquaculture operations to the aquatic environment may be summarised into three broad categories: aquaculture production (continuous discharges); farm activities (periodic discharges); and medication (periodic discharges). The behaviour of any type of waste released into the water column will largely depend on the current conditions and bottom topography of the area in question. Indeed, the selection of sites with good water exchange, together with good management practices that minimise food waste, will reduce the environmental impacts of aquaculture and optimise fish health and growth. Environmental awareness and sustainability The aquaculture industries have experienced considerable expansion over past decades, with attention being paid subsequently to the environmental effects of such activities. It is not possible to generalise and distinguish between the actual and potential impacts of aquaculture given that a multitude of approaches is in place. However, in general, the potential impacts of aquaculture are wide-ranging, from aesthetic aspects to direct pollution problems (O'Sullivan 1992). Marine aquaculture operations and the associated infrastructure can, for example, impact on scenic rural areas. Fish production generates considerable amounts of effluent (e.g. nutrients, waste feed and faeces, together with associated by-products such as medication and pesticides) which can have undesirable impacts on the environment (Gowen and Bradbury 1987; Ackefors and Enell 1994; Wu 1995; Axler et al. 1996; Kelly et al. 1996). There may also be unwanted effects on wild populations, such as genetic disturbance (Crozier 2000; Fleming et al. 2000), and disease transfer by escapees or ingestion of contaminated waste (Heggberget et al. 1993), and effects on the wider ecosystem. Table 1 summarises current environmental concerns arising from marine aquaculture operations. Other coastal activities can also have unacceptable impacts on aquaculture operations (e.g. effluent discharges and physical conflicts with other users) (Anderson 1995; Shereif and Mancy 1995; Phyne 1996).


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Table 1. Current environmental concerns arising from marine aquaculture operations Potential Direct Impacts Potential Consequences Management Actions • Organic enrichment • Nutrient enrichment • Chemicals release • Spread of diseases

• Impact on wildlife/ habitats

• Trigger of toxic blooms

• Demise/quality of wild stocks

• Locational guidelines • Biomass maximum • Maximum feed limit • Restricted use of

chemicals • Management

Guidelines (inc. Codes of Practice/Conduct)

• Escapees • Genetic dilution • Demise of wild stocks

• Improved cage design • Management

Guidelines • Interaction with other

coastal activities

• visual impacts and conflict with e.g. tourism, recreational fishing, maritime transport

• Locational guidelines • Derive regional/local

Coastal Plans and integrate with national Coastal Management Plan

Types of waste and effluent deposition Discharges from aquaculture to the aquatic environment may be summarised into three broad categories: • Aquaculture production (continuous discharges): mainly deposition and spreading of

organic material in the form of faeces, waste food and accidental food spillage. The various organic compounds include proteins, carbohydrates, vitamins and pigments. Some inorganic excretory products are also released, mainly ammonium and species-dependent trace quantities of bicarbonate, phosphate and urea.

• Farm activities (periodic discharges): processing waste and eventual regulated dumping of mortalities, usually in silage form. Mainly inorganic discharges including detergents, effluent from net washing (antifoulants, heavy metals), and eventual particle deposition from, for example, waste incineration. Most of these discharges are usually released from the non-ongrowing part of the aquaculture farm.

• Chemicals (periodic discharges): various chemical compounds may also be released from the production sites, mainly antibiotics and pesticides.

The behaviour of any type of waste released into the water column depends on the hydrographic conditions, bottom topography and geography of the area in question. Dissolved products include ammonia, phosphorus, dissolved organic carbon (which includes dissolved organic nitrogen and dissolved organic phosphorus) and lipids released from the diet, which may form a film on the water surface (Black 2001). The environmental impact of these dissolved products depends on the rate at which those nutrients are diluted before being assimilated by the pelagic ecosystem. In restricted exchange environments, there is a risk of high levels of nutrients accumulating in one area (hypernutrification) and potentially creating undesirable effects (Midlen and Redding 1998). In shallow waters, with weak currents, solid waste products from aquaculture installations will settle to the bottom close to the discharge point. In this case, continued production can


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give rise to a rapid local accumulation of waste material on the sea floor. On the other hand, effluent released into deeper waters, or where the bottom is well swept by strong currents, will be dispersed over a larger area. The vast majority of the chemicals used by the aquaculture industry are for the treatment or prevention of disease, although cleaning agents are also used (Costello et al. 2001). The environmental concerns over the use of chemicals in the aquatic environment relate to the following (Midlen and Redding 1998): (i) the direct toxicity of the compounds; (ii) the development of resistance to compounds by pathogenic organisms; (iii) the prophylactic use of therapeutants; and (iv) the length of time they remain active in the environment. In addition to selecting sites with good water exchange, management practices that minimise food waste and chemical usage will also reduce the environmental impacts of aquaculture and optimise fish health and growth (Cho et al. 1994; Ervik et al. 1994).

3 Environmental impacts Aquaculture waste released into the water column can be in both dissolved and particulate forms. In the marine environment the effects of dissolved nutrients from aquaculture waste are generally rapidly dissipated. However, particularly in brackish areas or where there is exceptionally poor water exchange, dissolved forms of phosphorus and nitrogen can stimulate the production of phytoplankton. The resulting algal blooms can harm both farmed and wild fish. It should also be noted that different water bodies have different limiting nutrients; the Atlantic is thought to be nitrogen limited whereas the Mediterranean is thought to be phosphate limited. Neither bicarbonate nor phosphate are considered to be important factors in marine pollution but the latter can contribute to eutrophication in brackish areas. Wastes from fish farming are usually considered in 3 overlapping categories: 1. Soluble wastes, being the products of fish excretion and including reactive nitrogen

species such as ammonia 2. Solid wastes, being mainly uneaten feed material and faeces 3. Chemical wastes, being medicines - such as antibiotics and anti-parasitcs, disinfectants,

and antifoulants – such as tin or copper compounds formulated into coatings or paints. 3.1 Soluble wastes Cage farming results in considerable nutrient release into the water column. For farmed salmon and trout mass balance models have been developed for nitrogen (Gowen & Bradbury, 1987; Hall et al., 1992), phosphorus (Holby & Hall, 1991) and silicon (Holby & Hall, 1994). According to these models 50% of the nitrogen and 28% of the phosphorus supplied with the food is wasted in dissolved form (mainly NH4 and PO4). Seasonal variability in food supply determines the seasonal variability in environmental loss of carbon, nitrogen and phosphorus towards the seabed and the water column. However the resulting seasonal pattern in nutrient availability in the water column is anticipated to differ from the “natural” seasonal pattern. In the case of non-impacted marine ecosystems, nutrients are abundant during winter and early spring and are gradually depleted in the surface waters during the warm season, whereas in mariculture-impacted ecosystems most of the nutrient loss during feeding (and consequently most of the nutrient enrichment in the water column)


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occurs during the warm period (i.e. summer and early autumn). Furthermore, fish farming wastes provide dissolved nitrogen and phosphorus but not silica and therefore such wastes could be expected to favour the growth of certain phytoplankton groups such as flagellates or cyanobacteria (Parsons et al., 1978; Doering et al., 1989), though not the Si-limited diatoms. This modification of the structure of the phytoplankton community could in turn further affect zooplankton and subsequent trophic links since the quality of phytoplankton biomass as a food resource varies greatly among different phytoplankton groups (Bianchi et al., 1993). There is limited information on such impacts; however investigation into phyto- and microzooplankton communities in Mediterranean fish farms revealed no significant changes in terms of species composition or diversity between cages and control sites (Pitta et al., 1997). Soluble nutrients from aquaculture may constitute a risk of eutrophication. The EU definition of eutrophication is: "the enrichment of water by nutrients especially compounds of nitrogen and phosphorus, causing an accelerated growth of algae and higher forms of plant life to produce an undesirable disturbance to the balance of organisms and the quality of the water concerned". The undesirable consequences of eutrophication include(Black et al., 2002):

• increased abundance of micro-algae, perhaps sufficient to discolour the sea and be

recognised as a bloom or “Red Tide”; • foaming of seawater; • killing of free-living or farmed fish, or sea-bed animals; • poisoning of shellfish; • changes in marine food chains; • removal of oxygen from deep water and sediments as a consequence of the sinking

and decay of blooming algae . A nutrient budget for Atlantic salmon shows that the majority of the nutrients that are input to the system are subsequently lost either directly as soluble wastes or through losses from solid particulate wastes (figure 2). One method of determining the risk of elevated nutrient concentrations is called the Equilibrium Concentration Enhancement (ECE) model. The ECE model is a box model of dissolved ammoniacal nitrogen arising from farmed fish occurring within an enclosed body of water of known dimensions that is being exchanged at a steady rate (Gillibrand and Turrell, 1997).


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Figure 2. A budget for the flow of nutrients from oceanic wild caught fish to the coastal

environment for a harvest of 1 kg of farmed salmon assuming no substitution with vegetable protein or oil and a ratio of fish feed to product of 1.2:1 (Black, 2001)

Model input data include the flushing time and volume of the system, rate of excretion of ammonia by fish and annual production. Using this model, systems can be ranked in terms of the potential contribution of aquaculture to raise nutrient levels and assigned a Nutrient Enhancement Index (NUI). In Scotland this index was proposed in a government document

Wild fish 17% protein, 7-10% oil, 75% water

2.8 kg for protein

Plus 0.8 – 2.3 kg extra for oil

Fish Feed 40% Protein 30% oil 9% water 1200g: N 96g P 18g

Harvest Fish 1 kg N 26g P 3 2

Mortalities and escapes N 1.9g P 0.4g C 13g

Particulate wastes N 22g P 9.5g

Soluble wastes N 46g


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called the Locational Guidelines for Fish Farming (Gillibrand et al., 2002). Table 1 shows the predicted ECE for Scottish sea loch systems, their NUI and the distribution of lochs by NUI.

Table 1. ECE values for nitrogen for Scottish sea lochs and classification using the Nutrient Enhancement Index (Gillibrand et al., 2002). For an ECE of 0 (i.e. where no emissions from aquaculture exist, a value of 0 is assigned).

ECE (µmol l-1) Number of Scottish sea lochs Nutrient Enhancement Index > 10.0 5 (4.5 %) 5 3.0 – 10.0 15 (13.5 %) 4 1.0 – 3.0 23 (20.7 %) 3 0.3 – 1.0 22 (19.8 %) 2 < 0.3 46 (41.4 %) 1 Total 111 More complex models exist, where the biological response from nutrient additions is predicted in terms of phytoplankton biomass or chlorophyll. Once such model, currently being adapted for an aquaculture (Laurent et al., 2006) was initially developed for predictions of the eutrophication potential of nutrient discharges from sewerage marine outfalls. This model, originally called the CSST model, has been described by Tett et al. (2003). Such models are capable of being applied at a range of scales and in non-enclosed water bodies, provided appropriate boundary conditions are known, can be measured or can be predicted from larger scale models. Absolute standards for what constitutes eutrophication are not available. What constitutes a harmful disturbance to the ecosystem will vary according to the normal seasonal cycle which may be very different in temperate, sub-tropical and tropical regions. Thus to assess the acceptable level of phytoplankton biomass will require some knowledge of annual nutrient cycles and the primary production response for different environments and such measurements have often not been made. An alternative is to look for a deleterious ecosystem response such as the hypoxia caused by enhanced carbon inputs to deeper waters and sediments. Hypoxia may constitute a threat not only to the ecosystem but to the farmed fish themselves. Poor water quality, resulting from self effects of the fish farm, tends to be a very localised problem, mainly affecting the cultured fish and local animal and plant populations that cannot move away from the affected water body. The main problems are H2S/ methane in the water, arising from outgassing of severely affected sediments, as well as local oxygen depletion, generally due to poor water exchange and/or failure to prevent clogging of the nets by fouling organisms.. These issues need to be considered primarily as a local, management-related problem but should not be disregarded when considering broader environmental interests. In addition, dissolved products including ammonia, phosphorus and dissolved organic carbon may lead to possible hypernutrification and the potential for the development of eutrophic problems in the vicinity of fish farms (as suggested in MacGarvin (2000)). The nature of this cause and effect relationship and its theoretical connection with the occurrence of blooms of undesirable algal species in coastal waters is far from clear and should be addressed in the context of multiple coastal use. Whilst some research is on-going in this area (OAERRE 2001) much more attention and finances should be directed into this field.


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Solid wastes Particulate material from aquaculture effluent in the water column is generally a relatively localised problem, although there is a risk of affecting wild fish populations and other aquatic animals. Both particulate and dissolved waste can be of consequence in areas where the ecosystem is particularly fragile, such as conservation areas (e.g. coral reefs, seagrass beds), or where the water is used for other activities, such as bathing beaches. Perhaps the most conspicuous impact of mariculture is the sedimentation of wasted food and faecal material under the farm cages. The accumulation of such material on the seabed results in a loose and flocculent black sediment under fish cages (Hall et al., 1990; Angel et al., 1995), commonly named “fish farm sediment” (Holmer, 1992). This sediment is characterized by low values of redox potential (Hargrave et al., 1993), high content of organic material (Hall et al., 1990; Holmer, 1992) and accumulation of nitrogenous and phosphorous compounds (Hall et al., 1992; Holby & Hall, 1991; Karakassis et al., 1998). These changes in the physical and chemical characteristics of the seabed induce conspicuous changes in the structure of the benthic communities (O’Connor et al., 1989; Weston, 1990; Pocklington et al., 1994). Kupka-Hansen et al. (1991) have reported a farm sediment thickness up to 30 cm (with 60-90% SWC) in W. Norway indicating that no fauna larger than 5 mm was found when thickness exceeded 20 cm. An azoic zone has also been reported from the Mediterranean with farm sediment thickness less than 6 cm (Karakassis et al., 1998). The succession pattern along the nutrient enrichment gradient is similar to that described by Pearson & Rosenberg (1978). Although significant impacts have been reported at distances as far as 100 m from the cages (Weston 1990) in general it seems that this impact is a highly localized phenomenon not exceeding 20-50 m around the cages (Beveridge, 1996). A partial recovery process of the sediment quality during the winter (Figure 2) has been reported from a Mediterranean fish farm (Karakassis et al., 1998). The major solid wastes from fish farming are from uneaten feed and faecal material. The quantity of both these components will vary by species, food conversion ratio, food digestibility and the skill of the operator in matching feed availability to demand. Benthic macrofaunal communities in sediments receiving normal detrital inputs derived from planktonic production in the overlying water column are species rich, have a relatively low total abundance/species richness ratio and include a wide range of higher taxa, body sizes and functional types, i.e. they are highly diverse communities (Pearson, 1992). The total productivity of the system is dependent on the availability of food - organic matter, and its quality. Animals have evolved to maximise the utilisation of the available resource by virtue of a wide range of feeding modes and some species can vary their mode of feeding depending on environmental factors. Benthic types include filter feeders that gather detrital material from the water column above the sediment, surface deposit feeders that feed on material deposited on the sediment surface, sub-surface deposit feeders that consume buried organic material by burrowing, and carnivores that prey on other macrofauna. Microbes degrade organic material and are themselves consumed by macrofauna, mediating the transfer of nutrients up the food chain. A variety of terminal electron acceptors are used by different bacterial communities in marine sediments. The oxygen concentration at any point in the sediment is dependent on the rate of its uptake, either to fuel aerobic metabolism, or to re-oxidise reduced products released from deeper in the sediment. When the oxygen demand caused by input of organic matter exceeds the oxygen diffusion rate from overlying waters, sediments become anoxic


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and anaerobic processes dominate. As sediments become more reducing with increasing distance from the water column interface, a range of microbiological processes become successively dominant in the order: • aerobic respiration, ammonium oxidation (to nitrite) and nitrite oxidation (to nitrate).

These aerobic nitrifying processes are inhibited by sulphide and are, therefore, of limited importance in sediments beneath marine fish farms;

• denitrification (producing dinitrogen from nitrate); • nitrate reduction (producing ammonium from nitrate) and manganese reduction; • iron reduction; • sulphate reduction (producing hydrogen sulphide) • and lastly, under the most reducing conditions, methanogenesis (producing methane). To some extent, these processes may overlap spatially. Oxygen in sediment porewaters is rapidly depleted and sulphides are generated by sulphate reduction, which is the dominant anaerobic process in coastal sediments (Holmer and Kristensen, 1992). These effects on sediment biogeochemical processes have profound consequences for the seafloor fauna that becomes dominated by a few small, opportunistic species, often at very high abundances, and confined to the upper few centimetres of the sediment (Brooks and Mahnken, 2003a; Brooks et al., 2003a; Brooks et al., 2003b; Hargrave et al., 1997; Heilskov and Holmer, 2001; Holmer et al., 2005; Karakassis et al., 1999; Pearson and Black, 2001; Pearson and Rosenberg, 1978; Weston, 1990). Away from the farm, as organic material flux and oxygen demand decreases, animal communities return to background conditions typified by high species diversity and functionality (Gowen and Bradbury, 1987; Nickell et al., 2003; Pereira et al., 2004). The redox potential (Eh) profile measured down the sediment column to a depth of 10-15 cm gives a useful guide to the relative degree of carbon enrichment in the sediments (Pearson and Stanley, 1979). Positive Eh values are indicative of aerobic conditions whereas negative values are associated with anaerobic microbial processes. Under normal rates of detrital carbon input to sediments the redox discontinuity level (RDL), i.e. the point at which anaerobic processes become predominant, lies some centimetres below the surface. As carbon inputs increase the RDL approaches ever closer to the surface as the BOD (Biological Oxygen Demand) within the sediments increases. Eventually, under very high detrital inputs, the RDL coincides with the sediment/water interface, where, under low flow conditions, it might even rise into the water column. It is important to emphasise that highly organically enriched sediments can occur naturally from large marine or terrestrial inputs of detritus. This may be transient and localised or long-lived and wide scale. Hypoxia/anoxia in sediments and overlying water occurs when the supply of new oxygenated water is poor as may be the case, for example, in deep silled fjordic systems. In such systems, benthic communities are modified and specialist opportunist animals may dominate. The process of organic particulate material impacting the seabed and causing benthic effects is amenable to modelling (Cromey and Black, 2005; Cromey et al., 2002a; Cromey et al., 2002b; Silvert and Cromey, 2001; Silvert and Sowles, 1996). A widely used model is that of Cromey et al. (2002a) – DEPOMOD. The outputs of this model are in terms of accumulated carbon per unit area of sea bed per unit time – the word accumulation is used here as


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resuspension processes are accommodated in the model. The output is visualised as a contour plot of carbon deposition on a spatial grid (an example is shown in figure 3).

Figure 3. Example of a DEPOMOD particle tracking model output. Cages are represented

by squares; the contours show the accumulation of organic matter on the seabed beneath and around the cages.

In Scotland, as in several other countries, the regulator (Scottish Environment Protection Agency, SEPA) is required to manage the impacts of fish farming to avoid unacceptable damage of the sea bed and its fauna. SEPA have gone a little further than most regulators in giving some examples of where they think the boundary between acceptable and unacceptable seabed conditions lies. They have established Sediment Quality Criteria (SQC, table 2) as indicators of when they will take action in order to reduce impacts, e.g. by reducing the maximum allowable biomass or by entirely revoking the discharge consent. The SQC are not the only criteria used – SEPA will accept and consider all the available evidence – but as many benthic indicators co-vary, they do offer a meaningful insight into what SEPA consider to be unacceptable benthic conditions. Discharge consents have monitoring conditions specified in detail: both their level (i.e. the number of stations, types of measurement and analysis) and their frequency are matched to the perceived risk of the farm. For example, a small farm over a hard sediment with strong currents will be monitored less intensively than a large farm over a soft substrate with weak currents. This process is given in great detail, together with its underlying philosophy and science, in the regularly-updated Fish Farm Manual that can be downloaded from the SEPA website (www.sepa.org.uk). Table 2 Sediment Quality Criteria (SEPA Fish Farm Manual, Annex A) Determinand Action Level Within

Allowable Zone of effects Action Level Outside Allowable Zone of effects

Group 3

0 50 100 150 200 250 300 350 400 450 500

Easting (m)

0

50

100

150

200

250

300

350

400

450

Nor

thin

g (m

)

A3B3M

25

500

2500

5000

15000

g solids m bed yr -2 -1


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Number of taxa Less than 2 polychaete taxa present (replicates bulked)

Must be at least 50% of reference station value

Number of taxa Two or more replicates with no taxa present

Abundance Organic enrichment polychaetes present in abnormally low densities

Organic enrichment polychaetes must not exceed 200% of reference station value

Shannon -Weiner Diversity

N/A Must be at least 60 % of reference station value

Infaunal Trophic Index ( ITI )

N / A Must be at least 50% of reference station value

Beggiatoa N/A Mats present Feed Pellets Accumulations of pellets Pellets present Teflubenzuron 10.0 mg/kg dry wt/5cm core applied

as a average in the AZE 2.0 ug/kg dry wt/5 cm core

Copper* Probable Effects 270 mg/kg dry sediment Possible Effects 108 mg/kg dry sediment

34 mg/kg dry sediment

Zinc* Probable Effects 410 mg/kg dry sediment Possible Effects 270 mg/kg dry sediment

150 mg/kg dry sediment

Free Sulphide 4800 mg kg-1 (dry wt) 3200 mg kg-1 (dry wt) Organic Carbon 9% Redox potential Values lower than -150 mV (as a depth average profile)

OR Values lower than -125 mV (in surface sediments 0-3 cm)

Loss on Ignition 27% *A detailed description of the derivation of these action levels may be obtained from SEPA on request. The SQC (or Action Levels, Table A1) are levels at which SEPA may take action against the farmer i.e. reduce or remove the consent to discharge. Implicit within the approach are:

a) that the farmer is required to monitor the sediments around the farm to measure compliance or otherwise, and

b) the concept of the Allowable Zone of Effects (AZE). The AZE represents an area around the farm where some deterioration is expected and permitted. Thus for several determinands, two SQCs are proposed: one within the AZE and one at any point outside the AZE. The SQC inside the AZE is less demanding than that outside the AZE. The SQC approach thus constrains the level of ecological change while the AZE limits the spatial extent of major changes.

3.1.1 Organic matter A variety of physical, chemical and biological changes occur in sediments exposed to continual deposition of organic waste from aquaculture. Where the rate of waste deposition, mainly in the form of faeces and waste food, exceeds the natural rate of breakdown in the sediments, a layer of this waste will settle on top of the natural sediment at the sea floor. This fine-grained, and often slimy, material has a very high organic content. In the absence of breakdown of this organic material, the sediments can become very acidic, and toxic


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gasses such as H2S and methane may be produced. In extreme cases, these gasses may bubble out through the sediments to the detriment of the fish in the overlying water mass. In these cases, there are generally no macrobenthic animals remaining in the sediments. If conditions are allowed to deteriorate, the sediments may become oxygen depleted or even fully anoxic. The deposited material on the sea floor may become blackish in colour, with a noxious smell and a layer of white, chemoautotrophic bacteria (Beggiatoa sp., Achromatium sp.) usually forms at the surface (Midlen and Redding 1998). During the deterioration of healthy sediments, the community structure of benthic animals changes: the less resistant forms die out to be replaced by fewer, more tolerant forms, which then become more abundant. This predictable response forms the basis of the biological component of monitoring programmes. The depth to which the sediments are oxygenated gradually decreases, from often many centimetres deep, to a very shallow or even absent oxygenated layer. The depth of sedimentary oxygenation can be detected by measuring the redox (Eh) potential down the sediment profile (for example at 1cm intervals). The point at which the sediment switches from being well oxygenated to anoxic is seen by a marked switch in redox potential readings, known as the redox potential discontinuity (RPD) layer. This is also a useful indicator of the status of the sediments and is widely used in many monitoring schemes (Black 2001). Effluent from net-washing machinery can produce significant local sources of dissolved contaminants in the water column. This can, however, be minimised in areas where there are centralised facilities for net-washing, which use water treatment or filtering of the effluent. Chemicals and other additives Because vitamins and other nutritional additives as well as antibiotics and other medication are usually introduced via feed, some of these compounds will also remain bound to the organic matter in waste feed and faeces. The effects of these compounds on the aquatic environment and the food web are not yet well understood and potential deleterious effects on the benthic system have been suggested (Black et al. 1997; Davies et al. 1998; Grant and Briggs 1998) and need further research. Several classes of chemicals are used in fish farming. Three of the most important are antibiotics, antiparasitics and antifoulants. Since the advent of effective fish vaccines for several important bacterial diseases, the use of antibiotics in salmon culture has declined since the early 1990s (Alderman, 2002), but less is known about the use of antibiotics in other species, particularly those in the developing world (Graslund and Bengtsson, 2001; Holmstrom et al., 2003; Tacon et al., 1995). Antibiotics The main concerns relating to antibiotic use relate to the possibility of the development of bacterial resistance that can be transferred to human pathogens reducing the efficacy of antibiotics in human medicine (Cabello, 2004; Cabello, 2006). Prophylactic use of antibiotics is a particular concern, as is adequate testing of aquaculture products for antibacterial residues. Where residues persist, low levels of antibiotic intake can stimulate resistance in human pathogens but another concern is that some of the antibiotics used in aquaculture may not be approved for human medicine and carry a health risk. For example, nitrofurans (e.g. furazolidone) are a group of antimicrobials which possess either carcinogenic or mutagenic properties whose use is banned in many countries but still


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permitted in some. In general, limits have not been set for the concentrations of antibiotics permitted in the environment. It has been estimated that 50-80 tonnes of antibiotics per year were being used by the salmon industry in western Europe by the end of 1980s (Beveridge et al., 1994). The main impact of anti-microbial agents is related to the development of resistant strains of aquatic microbes which could have serious implications not only for farmed fish, but also for the wild populations or even for human health (IOE, 1992). The overuse or abuse of chemotherapeutants could also negatively affect the bacterial activity in the sediments (Beveridge et al., 1994) and therefore it could inhibit the main mechanism for recovery of heavily polluted areas under farm cages and interfere with the food chain processes both in the water column and the sediment. Antiparasitics Of the anti-parasitics, the best studied are those that treat infestations of sea lice on salmonids. These are generally highly toxic substances where the use is tightly regulated. In Scotland, for example, ecological quality criteria have been set both for both sediments and the water column www.sepa.org.uk/pdf/guidance/fish_farm_manual/annex/A.pdf and access to these medicines is strictly controlled. Medicines are only available to farmers after a multi-step regulatory process of authorisation that includes testing for efficacy, food safety and environmental effects. Antifoulants Antifouling agents are also widely used in mariculture. These slow-leaching biocides prevent the settlement of the juvenile plankton stages of fouling organisms on the cage structures. Some types of antifouling chemicals (e.g TBT) have been prohibited after having been proved to be highly toxic to molluscs (Davies et al., 1988) and they could affect farmed fish (Bruno & Ellis, 1988). The most infamous antifoulant product, now banned, is tributyl tin (TBT). TBT disrupts the endocrine system in invertebrates leading to imposex (Miller et al., 1999) and has now been replaced by products with a high concentration of copper. Copper, however, is also a potent toxin, hence it’s utility as an antifoulant, and quality criteria have been set for both the water column and sediments. A recent study at a fish farm in Scotland found copper at higher concentrations than the sediment quality criteria (Dean et al., 2007) although its toxicity in anoxic sediments may be limited by its precipitation as the sulphide (Brooks and Mahnken, 2003b). Effects on wild fish populations

3.1.2 Contamination with waste The attraction of aquaculture installations to wild fish is widespread, and seems to be a combination of the artificial reef effect (physical shelter) as well as food availability (the waste from the farmed fish). What is interesting in an international context is that aquaculture installations in Mediterranean waters seem to attract far greater populations of wild fish than is the case in Atlantic waters. Therefore, there is a greater risk of transfer of contaminants and pathogens via aquaculture waste to wild populations in the Mediterranean, or other warmer waters, than may be the case in colder Atlantic waters.


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Wild fish in the Mediterranean also seem to be responsible for removing a greater proportion of deposited waste from fish farms than is seen at Scottish and Norwegian farms. This is particularly important to bear in mind when modelling or predicting waste dispersal scenarios. 3.1.3 Parasites and diseases Currently, a major area of concern is the potential transfer of parasites (e.g. sea lice) and pathogens from farmed fish to wild fish stocks. These may be transmitted in two main ways – either by wild fish coming into contact with caged fish and their waste-products or by contact between escaped farm fish and wild individuals. Therefore, the frequency and intensity of parasite infestations in wild and cultured fish should be addressed by monitoring programmes. Any overlap between parasites and diseases in wild and cultured populations will help to identify potential vectors of pathogens. Comparing parasite densities and intensities at farm sites with those in unaffected areas will help to assess the effect of fish farms on the health of wild fish. The issue of potential infestation of wild fish populations with parasites of fish farm origin is a controversial one. Particular conflict may arise in areas where sport-fishing/angling is a major source of income and where parasite infestation threatens the popularity of a particular water course. Several national and international applied research programmes are in progress, which aim to trace the source of parasites on or in wild fish, using genetic and other methods. 3.1.4 Escapees and genetic transfer Another concern is that escaped farm fish may interbreed with wild populations, and contaminate the wild gene pool. The fear is that the resulting progeny will lose their homing ability and not spawn where expected, or that there will be a reduced ability to survive in the natural environment. These issues are generally addressed by national nature management authorities and will not be considered further in this document. Further information can be found in Youngson et al. (2001). The effect of accidental or intentional release of cultured fish into the marine environment causes widespread concern due to the risk of genetic loss, or of significant modification of important natural traits and adaptive variations of the wild populations (Hinder et al., 1991). Evidence of such interactions has been presented for the Atlantic salmon in N. Ireland (Crozzier, 1993). Exotic species, or varieties, selected for particular traits such as rapid growth or high fecundity, are commonly moved from one continent to another in an attempt to capitalize on already developed technologies and markets (Beveridge et al., 1994). Many of these species successfully colonise natural waters, after escaping from aquaculture systems. Therefore there is a potential for impact on natural communities, either through competition with natural populations, or (in the case of varieties) through modification of the gene pools of natural stocks. The introduction of exotic parasites or other disease agents along with the introduction of exotic farmed species, also causes concern for impacts on native populations which have not previously been exposed to such agents.


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Direct impacts on wildlife Disturbance of feeding and breeding of wildlife is also a potential consequence of mariculture since aquaculture activity involves increased human activity, noise, road and boat traffic and building of floating and land-based structures (Beveridge et al., 1994). Other potential impacts involve the attraction of predators and scavengers, due to increased densities of food resources, which could displace local species or increase the conflicts between human and wildlife resulting in increased (intentional or unintentional) mortality of natural populations. Some evidence of such risks was presented in respect of double-crested cormorants (Phalacrocorax auritus) which is perceived to be a significant pest of aquaculture (Glahn & Stickley, 1995; Nisbet, 1995). Anti-predator nets and shooting of potential predator species has been resorted to by fish farmers. However there is no evidence that such practices have had a significant effect on local bird population size, while even less is known with regard to the effects on mammal populations (Munday et al., 1992). Gowen (1992) noted that there is a lack of reliable data on this issue since the available information comes mainly from surveys using answers from questionnaires sent to the farmers.

Implications for biodiversity Almost all the above-mentioned types of impacts could result in decreasing diversity at various levels. However not all of them would induce such serious ecological changes as to affect biodiversity at a significant spatial scale, or to produce irreversible environmental changes. The effects on benthic communities are similar to those produced by several other sources of organic enrichment. The local decrease in diversity could be acceptable unless the following situations obtain:

(i) the damaged ecosystem constitutes the habitat of an endangered species (ii) the damaged ecosystem is a nursery ground for species affecting the ecology of a

broad marine area (iii) the damaged ecosystem is a rare and region-specific habitat (iv) the damaged ecosystem is impaired, to that an extent that its loss is irreversible on a

human time scale Marine systems harbour the highest number of phyla (Ray & Grassle, 1991) and the majority of marine phyla occur in the coastal zone, so it could be argued that the marine coastal zone is the most biologically diverse realm on the planet (Ray, 1991). Although deep-sea benthic biodiversity, in terms of species, has been estimated as extremely high (Grassle & Macioleck, 1992), it has been shown that biodiversity per unit area is higher in the coastal environment than in the deep sea (Gray et al., 1997; Karakassis & Eleftheriou, 1997). Furthermore coastal ecosystems contribute to the global carbon fixation, to an extent disproportional to the area they occupy. Estimations of productivity for the worlds' aquatic ecosystems, based on Walsh & Dieterle (1988), give a productivity 5 times higher in shelf ecosystems than that of the Ocean. Therefore it could be expected that large scale ecological impacts on the coastal zone would result in severe impairment of the environmental conditions beyond the limits of the impacted area. The destruction of seaweed communities for instance could be significant for biodiversity since they are essential habitats for various life cycles, while they also protect against shoreline erosion. However no such large scale impacts have been reported in relation to mariculture with the exception of mangrove habitats in the tropics used for the culture of milkfish and shrimp (Iwama, 1991; Beveridge et al., 1994).


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Hypernutrification and change of the trophy status of a particular area could be important only in highly enclosed areas while exposed areas with adequate water renewal are not expected to present eutrophication symptoms. Although significant increase in nutrient concentrations has been reported in the vicinity of fish farms by many authors (references in Beveridge, 1996) any significant increase in phytoplankton biomass has been rather exceptional (Wallin & Hakanson, 1991) and no large scale eutrophication has been reported in relation to aquaculture. However, such impacts may not be excluded in respect of the large-scale expansion of Aquaculture in combination with unsuitable siting and specific hydrographic conditions. Loss of biodiversity as a result of genetic interactions also depends on the state and the size of wild populations, the number of farmed individuals released into the marine environment and their ability to survive there in direct competition with native species (Munday et al., 1992). For instance the use of triploids, which are reproductively sterile, eliminates the risk of hybridization. However, in some cases, sterile fish live longer and are more robust and therefore it has been suggested that if competition or predation problems arise between stocked triploids and native species, then halting stocking would eliminate these interactions within one life cycle (Habicht et al., 1994). However in the case of large numbers of triploids, within this period of one life cycle, the natural populations could undergo a considerable decrease, resulting in "bottle neck effect" and considerable loss in genetic variability.

4 Scientific rationale behind monitoring methods Hydrographic/ water quality analyses and other assessments The dispersal of aquaculture discharges is dependent on the hydrographic conditions and water exchange mechanisms in the area. Environmental impacts on bottom substrates and the water column are minimised in areas with strong current flow. However, the degree of exposure must not exceed the physical tolerance of the production installations or the culture organisms. It is therefore important to measure a range of hydrographic variables, particularly during the planning or pre-licence phases. The most important information required is as follows: current speed; current direction; neap and spring variations; current residuals; bathymetry; winds and wave action; temperature; salinity; water stratification; other regional factors (such as ice action and super-cooled water in cold-water areas). Since hydrographic conditions are seasonally variable, measurements should be carried out at certain time intervals, to take into account different sea temperatures and seasonal stratification/upwelling of bottom water masses. In areas with significant tidal flows, current measurements should be made over a full tidal cycle, preferably at both spring and neap tides. Peaks in production activities should also be taken into consideration. Further, where possible, hydrographic measurements should be carried out at several sites within a given area to assess local variations in hydrography that could affect the aquaculture activities. For example, small variations in topographical features could cause local anomalies in current flow. Particular attention should be paid to such variability in sill-fjord, sill-loch areas, or where there are small coves within a bay or gulf. In addition, it is essential that accurate position fixing is used and that relocation of stations for time series analysis or for specific spatial monitoring, is considered.


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Sensory measurements (e.g. presence of gas bubbles, sediment colour, smell, consistence and softness) are also important in providing a quick assessment of local environmental conditions. In Norway, these are employed to assess environmental status in the local impact zone, as part of the MOM system (modelling - ongrowing fish farms – monitoring), and a score system is used to quantify the impact (Maroni 2000). In addition, redox measurements, as mentioned below, provide a semi-quantitative measure to assist with the sensory descriptors. Chemical analyses Chemical analyses of bottom sediments around aquaculture sites quantify the levels of contaminants or the chemical properties. These can be compared with natural conditions in the area as well as pre-defined levels of acceptability or consent. Some measurements can be carried out in situ using probes for which the results are immediately available. These include the following variables, all of which give information on the sediment loading: pH, redox (Eh), oxygen. Laboratory analyses most often focus on the following: organic content, hydrogen sulphide, sediment granulometry (indicative of organic deposition, bottom current speeds and also used for normalisation of levels of contaminants) and heavy metals. In addition, specific pesticides, medicinal agents and antifoulants may be analysed according to individual requirements. Benthic faunal analyses Changes in the organic loading of an area result in marked changes in the fauna present on the sea floor. Extensive work has been carried out in sedimentary areas, for which a model framework has been developed in relation to an organic enrichment gradient (Pearson and Rosenberg 1978). The normal situation in undisturbed areas is a high species diversity with few dominant species and relatively few individuals representing each species. In general, organic loading of the sediment leads to a reduction in species diversity and an increase in abundance of opportunistic species. However, if the level of pollution becomes very severe, even opportunistic species will disappear, such that there are no macro-organisms present in extreme conditions. In such cases, if sedimentation is not very heavy, a white bacterial layer may form on the sediment surface. This principle of macrofaunal responses to pollution gradients in marine communities is explained in detail in Pearson and Rosenberg (1978). In this way, the changing status of macrofaunal communities is a very sensitive indicator of sedimentary conditions, allowing the detection of environmental impacts, where other methods might not record them. Specific methodology concerning benthic faunal analysis (Holme and McIntyre 1984) would include, for example, specification of sieve mesh size (e.g. in Northern latitude coastal areas, 1mm should be used, whereas in similar situations in the Mediterranean, 0.5mm should be used), sample preservation and level of taxonomic identification (for standard monitoring practices in low sensitivity situations, a specific taxonomic resolution could be agreed, e.g. Family level; (Kingston and Riddle 1989; Karakassis and Hatziyanni 2000)). Rigorous analytical quality control procedures must be developed for all areas of practical work and implemented.


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5 Aquaculture Regulation Marine fin-fish aquaculture continues to expand, although expansion has to some extent plateaued in several of the developed countries where modern intensive aquaculture was pioneered. For example, in Scotland production increased steadily reaching a peak in 2003 but has since been variable (figure 1). There are several reasons for this including disease and market conditions, but one reason has been constraints placed on farmers by planners and regulators.

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Figure 1. Scottish annual salmon production, tonnes. (Frs, 2006) Planners must ensure that aquaculture developments meet aesthetic, social and economic criteria, and that there is harmonisation between new developments and local infrastructure capacity or other resource use e.g. tourism. Planners and regulators have duties to ensure that developments do not adversely affect the environment. The objectives of regulation can be separated into three areas: 1. protection of legitimate users of the environment, such as tourists or fishermen, such that

resources are fairly distributed. 2. protection of the environment for its biological structure including protection of

important/rare habitats and species 3. protection of ecosystem functions such as the recycling of nutrients and the maintenance

of oxygen levels The first of these is the subject of the evolving “discipline” of Integrated Coastal Zone Management (ICZM) which has 8 broad principles (Defra, 2006). It is worth presenting these here in full: a. A broad holistic approach

The objective of a holistic approach is to forego piecemeal management and decision making in favour of a more strategic approach which looks at the ‘bigger picture’,


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including cumulative causes and effects. This means considering the conservation value of natural systems alongside the human activities which take place on land and coastal waters. Taking a holistic approach will also involve looking at the problems and issues on the coast in the widest possible context, including looking at the marine and terrestrial components of the coastal zone and considering how different issues conflict or interact together.

b. Taking a long term perspective Successful coastal management must consider the needs of present and future generations. Therefore, administrative structures and policies required to manage the environmental, social and economic impacts now, must also be adaptable to take account of, and acknowledge, uncertainties in the future.

c. Adaptive management The … coastline has been subject to constant physical and economic changes over the years, and management of such a dynamic environment requires measures which are able to adapt and evolve accordingly. Successful management should reflect this principle by working towards solutions which can be monitored effectively.

d. Specific solutions and flexible measures Coastal management measures for each stretch of coast must reflect and accommodate the many variations in the topography, biodiversity and local decision-making structures. Integrated management should therefore be rooted in a thorough understanding of the specific characteristics of an area i.e. its local specificity.

e. Working with natural processes The natural processes of coastal systems are continual, so it becomes necessary in some instances to adopt a different approach which works with natural processes rather than against them. By recognising the physical impacts and the limits imposed by natural processes, decisions regarding the human impact on the coastal zone are made in a more responsible manner and are more likely to respond to environmental change.

f. Participatory planning In the past stakeholders may not have had sufficient opportunity to contribute towards the development and implementation of coastal management measures or programmes. Participatory planning incorporates the views of all of the relevant stakeholders (including maritime interests, recreational users, and fishing communities) into the planning process. It can also help to promote a real sense of shared responsibility and coastal stewardship by reducing conflict as real issues, information and activities which affect the coast can be aired more openly.

g. Support and involvement of all relevant administrative bodies Administrative policies, programmes and plans (land use, spatial, energy, tourism and regional development for example) set the context for the management of coastal areas and their natural and historical resources. Addressing the problems faced by … coastal zones will therefore require the support and involvement of all relevant administrative bodies at all levels of government to ensure cooperation, coordination and that commons


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goals are achieved. It is therefore essential to engage key bodies from the start so that decisions are consistent and firmly based on local circumstances.

h. Use of a combination of instruments Managing the different activities which take place on the coast requires the use of a number of different policies, laws and voluntary agreements. While each of these approaches is important, achieving the right combination is key to resolving conflicts, as these instruments should work together to achieve coherent objectives for the planning and sustainable management of coastal areas.”

The second objective, the protection of ecosystem structure, may be intimately linked to the third, protecting ecosystem function, especially where the structure of habitats have dominant functional roles. For example, mangroves have been shown to have key functional roles in flood protection, nutrient recycling and as nursery areas (Holmer, 2003; Primavera, 1998; Primavera, 2005). However, habitats may be deemed worthy of protection when their precise contribution to ecosystem function is unknown but they are considered to be rare or have rare species assemblages e.g. cold water corals and moves to protect them from trawling damage (Roberts et al., 2006). Interactions between aquaculture and sensitive habitats or species can be minimised by establishing aquaculture zones in areas with less sensitive/important/rare habitats, or by designations that more closely regulate developments with respect to their interactions with particular features of concern. In Europe, Special Areas of Conservation (SACs) have been established under the Habitats Directive (92/43/EEC) for the protection of specific habitats. As this is an important piece of legislation, with wide ranging impact, it is worth exploring this further. The UK Joint Nature Conservation Council gives the following background on its website (www.jncc.gov.uk).

“In 1992 the European Community adopted Council Directive 92/43/EEC on the Conservation of natural habitats and of wild fauna and flora (EC Habitats Directive). This is the means by which the Community meets its obligations as a signatory of the Convention on the Conservation of European Wildlife and Natural Habitats (Bern Convention). …The provisions of the Directive require Member States to introduce a range of measures including the protection of species listed in the Annexes; to undertake surveillance of habitats and species and produce a report every six years on the implementation of the Directive. The 189 habitats listed in Annex I of the Directive and the 788 species listed in Annex II, are to be protected by means of a network of sites. Each Member State is required to prepare and propose a national list of sites for evaluation in order to form a European network of Sites of Community Importance (SCIs). Once adopted, these are designated by Member States as Special Areas of Conservation (SACs), and along with Special Protection Areas (SPAs) classified under the EC Birds Directive, form a network of protected areas known as Natura 2000. The Directive was amended in 1997 by a technical adaptation Directive. The annexes were further amended by the Environment Chapter of the Treaty of Accession 2003. The Habitats Directive introduces for the first time for protected areas, the precautionary principle; that is that projects can only be permitted having ascertained no adverse effect on the integrity of the site. Projects may still be permitted if there are no alternatives, and there are imperative reasons of overriding public interest. In


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such cases compensation measures will be necessary to ensure the overall integrity of network of sites.”

In Loch Creran on the west of Scotland, for example, an SAC has been implemented to protect the unique reefs of the tube building polychaete Serpula vermicularis and reef forming horse mussels Modiolus modiolus. Several management principles are invoked in marine SACs (www.argyllmarinesac.org):

• Management should enable the natural habitat types and the species habitats concerned to be maintained or, where appropriate, restored at a favourable conservation status.

• Steps must be taken to avoid deterioration or disturbance of the habitats and species for which the site has been designated.

• Activities, plans or projects likely to have a significant effect upon the features of the site must be subject to an appropriate assessment. A development that would have an adverse effect on the conservation interests of the site should only be permitted if there is no alternative solution and there are imperative reasons of over-riding public interest, including those of a social or economic nature.

• Monitoring must be undertaken at each site to monitor the condition of the conservation features and to assess the effectiveness of management measures.

• Management of the site must take into account the economic, social, cultural and recreational needs of the local people.

Loch Creran was one of the original sites in Scotland developed for fish farming with continuing major aquaculture activity both from shellfish and salmon faming. However, any proposed change to the aquaculture activity must be assessed against these principles. In practice, the operator has found it possible to continue to farm in Loch Creran, despite the SAC, by being able to show that any proposed changes will have little impact on the conservation feature. In general, the science of interactions between sensitive habitats and aquaculture is not well developed, and it is only where catastrophic impacts occur on relatively visible habitats, such as mangroves or sea grasses (Holmer et al., 2003; Orth et al., 2006), that research effort is focussed. Thus for many habitats, neither the precise functional significance nor the sensitivity to aquaculture is known with any certainty and regulation is more problematic. In addition to loss of habitats, reduction in genetic diversity can also be regarded as having a structural component that may or may not also have a functional aspect. For example, the release of genes, disease organisms and parasites for salmon culture and their effects on wild populations has received considerable attention. Wild Atlantic salmon maintain their high degree of population structure by homing to natal rivers for breeding. The resulting genetic diversity has been shown to be important in terms of fitness and can be severely eroded by interbreeding with escaped cultured salmon or from intentional translocations of fish with maladapted genes (Mcginnity et al., 2003). Modelling work has shown that under high rates of intrusion of cultured fish significant changes to populations will occur that may not recover even if the intrusion rate is reduced (Hindar et al., 2006). For marine species, such as Atlantic cod and European sea bass, the consequences of genetic interactions on populations are not well established.


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The potential interactions between parasitic sea lice on wild and cultured populations have also received considerable attention and publicity. Salmon host ectoparasitic parasitic sea lice Lepeophtheirus salmonis and Caligis elongatus which feed on the skin and underlying tissue. In salmon culture these can increase rapidly to high abundance if left untreated and infestations can result in severe legions and mortality. Sea lice larvae are adept at find hosts and increasing infestations on juvenile Atlantic salmon and sea trout Salmo trutta were blamed on infection pressure from salmon farms. This in turn was linked to reduced survival and population declines of these salmonids. While causal links are extremely difficult to prove (Mcvicar, 1997), there seems little doubt that sea lice from farmed salmon can (Murray and Gillibrand, 2006) and do (Bjorn et al., 2001) infect wild salmonids and cause mortality, particularly to sea trout. Different approaches have been taken to regulating escapes and parasites in different countries. In Scotland, the government now require statutory declaration of escapes including the approximate number and size of fish, the location and date of escape (Scottish Statutory Instrument 2002 No. 193, The Registration of Fish Farming and Shellfish Farming Businesses Amendment (Scotland) Order 2002). In the planning process, developments of farms near important salmon rivers is discouraged, and farmers are obliged to show how they have taken steps to minimise escapes and improve containment. In Scotland, there is no statutory limit at present for the average lice infestation of salmon that is permitted before treatment with anti-parasitic medicines, although there are several voluntary agreements in place between the fishfarming sector and the fisheries sector (Area Management Agreements). This is in contrast to Norway where when statutory lice limits are exceeded, treatment is compulsory (Heuch and Mo, 2001). The situation is likely to change in Scotland as new legislation, aimed at ensuring farmers keep lice levels low is enacted (Aquaculture and Freshwater Fisheries Bill). This will require fish farmers to record levels of parasites and to ensure that burdens are kept low, but the precise mechanism for achieving this has yet to be announced. The state of knowledge of the structural effects of fish farming and their sound regulation is rather weak. In contrast, the third objective of regulation, to protect ecosystem function from the consequences of aquaculture, is often better understood and quantified. We now briefly review the current state of knowledge for the main environmental outputs from aquaculture, focussing on marine fish farming, and for each of these we comment on issues related to regulation. In general, the environmental interactions posed by intensive marine shellfish culture are fewer than for fin fish culture owing to the fact that shellfish are net extractive of nutrients. However, they do concentrate organic material and deposit wastes as faeces and pseudofaeces causing enrichment of the local benthos (Chamberlain et al., 2001) and, if cultured in sufficiently high density, in some areas can clear the water to such an extent that they reduce productivity (Smaal et al., 2001). Regulation on shellfish farming is, however, less well developed than for marine finfish farming owing to its lower perceived environmental risk.

6 Conclusions The regulation of aquaculture in developed countries has developed considerably over the past decade. This has been driven by the need to improve the scientific basis for


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management of this very high economic value sector. Regulators must base their decisions of good science in order to protect the environment and but at the same time allowing development with its economic benefits. This has particularly been the case for the major salmon growing countries, although Chile perhaps has still to catch up in a regulatory sense with its rapidly expanding industry. In the developing world, where aquaculture products are primarily for export, market pressures are increasingly brought to bear to ensure food safety, for example concerning residues, and there is also a growing awareness of environmental issues, particularly relating to habitat destruction related to shrimp farming. In the Philippines, regulation of the very large number of small scale fish farms where the market is local represents a significant regulatory challenge, but the potential environmental costs make rational regulation essential for the future of this industry. In this paper we have outlined some of the approaches being taken by aquaculture regulators in other countries. It is highly likely that some of these approaches will be relevant and adaptable to the Philippine industry. Acknowledgements This review is primarily based on papers by Kenny Black of the Scottish Association of marine Science, Yannis Karakassis et al of Biology Department of the University of Crete, and Sabina Cochrane of Akvaplan-niva.


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