Anthropogenic Nutrient Enrichment and Blooms of Harmful Micro-algae

Anthropogenic Nutrient Enrichment and Blooms of Harmful Micro-algae

Richard Gowen1, Paul Tett2, Eileen Bresnan3, Keith Davidson2, Alan Gordon4, April McKinney1, Steve Milligan5, David Mills5, Joe Silke6, Ann Marie Crooks1.

A report prepared

For

The Department for Environment, Food and Rural Affairs

September 2009

1 and Aquatic Ecosystems Branch, Agriculture Food and Environmental Science Division, Agri-Food and Biosciences Institute, Newforge Lane, Belfast, BT9 5PX. 2 Scottish Association for Marine Science, Dunstaffnage Marine Laboratory, Oban, Argyll, PA37 1QA. 3 Marine Marine Laboratory, P.O. Box 101, 375 Victoria Road, Aberdeen AB11 9DB 4 Biometrics Branch, Applied Plant Science and Biometrics Division, Agri-Food and Biosciences Institute, Newforge Lane, Belfast, BT9 5PX. 5Centre for Environment, Fisheries & Science, Pakefield Road, Lowestoft, Suffolk, NR33 0HT. 6 The Marine Institute, Rinville, Oranmore, Galway, Ireland. Contents

Page Summary iv

Part 1 Introduction 7

1.1 Introduction and Rationale ………….…………………..…...... 7 1.2 Methodology and Statistical Analysis ……………….………...... 8 1.3 Phytoplankton ………………………………………………………………...... 9 1.4 Blooms of Micro-algae ……………………..…………………………..……...... 14 1.4.1 Introduction ……………………………………………………………...... 14 1.4.2 Red , nuisance, noxious and harmful blooms ...... 15 1.5 Micro-algal Blooms and Their Effects ……………………………………...... 18 1.5.1 Introduction …………………………...... 18 1.5.2 Ecosystem effects of micro-algal blooms ...... 20 1.5.2.1 Pulse and press disturbance 1.5.2.2 Shading and smothering 1.5.2.3 Deoxygenation 1.5.2.4 Algal biotoxins 1.5.3 Micro-algal blooms and the human use of the ecosystem ...... 23 1.5.3.1 Algal biotoxins and human health 1.5.3.2 Mucilage and human health 1.5.3.3 The impact of micro-algal blooms on recreation and tourism 1.5.3.4 The economic impact of blooms on Fisheries and aquaculture

Part 2 An overview of Harmful Micro-algal Blooms and HAB Species in Coastal Waters of the and Republic of Ireland 29

2.1 General Introduction ……………………………………………………...... 29 2.2 Coastal Waters of the United Kingdom ……………………..……………...... 29 2.2.1 Introduction ……………………..……………...... 29 2.2.2 Species of Alexandrium ……………………..……………...... 30 2.2.3 Species of Dinophysis ……………………..……………...... 33 2.2.4 Species of Pseudo-nitzschia ……………………..……………...... 34 2.2.5 Karenia mikimotoi ……………………..……………...... 35 2.2.6 Other HAB species …..………………..……………...... 36 2.3 Coastal Waters of the Republic of Ireland ……………..………………...... 40 2.3.1 Introduction ..…………………………….……………..………………...... 40 2.3.2 Species of Alexandrium ………………………………..………………...... 40 2.3.3 Species of Dinophysis .………………………….……..………………...... 41 2.3.4 Species of Pseudo-nitzschia .……………….…………..………………...... 43 2.3.5 Karenia mikimotoi .………………..………………...... 44 2.3.6 Other HAB species .………………..………………...... 45

Part 3 Anthropogenic Nutrient Enrichment and Harmful Algal Blooms: A Literature Review 46

3.1 Introduction ……………..………………………………….…………...... 46 3.2 Nutrient Enrichment of Coastal Waters ………………………………….……………………… 50 3.3 Nutrient Enrichment and Blooms of Harmful Micro-algae ……..………………………………. 54 3.3.1 Introduction …………………………………………………………………………………. 54 3.3.2 The nutrient enrichment HAB hypothesis ……………………………………………… 55 3.3.2.1 Introduction

i 3.3.2.2 Historical and natural occurrence of HABs 3.3.2.3 Increased environmental awareness and monitoring of coastal waters 3.3.2.4 The influence of climate change 3.3.2.5 Introductions and transfers of new species 3.4 Case Studies ……………………………………………………………………………………... 69 3.4.1 Introduction …………………………………………………………………….. 69 3.4.2 Coastal waters of China ……………………………………………………………. 70 3.4.2.1 Introduction 3.4.2.2 Coastal waters of Hong Kong 3.4.2.3 Other coastal regions of China 3.4.2.4 The influence of the seasonal monsoon and climate change 3.4.2.5 Other human 3.4.2.6 Summary 3.4.3 Coastal waters of Japan …………………………………………………………... 86 3.4.3.1 Introduction 3.4.3.2 The Seto Inland Sea 3.4.3.3 Summary 3.4.4 The North Sea ……………………………………………..…………………………… 93 3.4.4.1 Introduction 3.4.4.2 Phytoplankton blooms in the wider North Sea 3.4.4.3 Phytoplankton blooms in coastal waters 3.4.4.4 Phaeocystis in the North Sea 3.4.4.5 The role of climate change and anthropogenic nutrient enrichment 3.4.4.6 Summary 3.4.5 Coastal waters of the continental United States of America ..……………………………….. 105 3.4.5.1 Introduction 3.4.5.2 Nutrient enrichment HAB relationships in coastal waters of the U.S. 3.5 Nutrient Ratios, Dissolved Organic and Particulate Nutrients …..………………………………. 109 3.5.1 Introduction ………………………………………………………………………………….. 109 3.5.2 Nutrient ratios ………………………………………………………………………………... 109 3.5.2.1 Theoretical considerations 3.5.2.2 Nitrogen to phosphorus ratio 3.5.2.3 Silicate limitation of diatom growth 3.5.3 Dissolved organic and particulate nutrients ………………………………………………….. 118 3.5.4 Nutrients and toxin production ………………………………………………………………. 120 3.6 Hypotheses Concerning the Occurrence of HABs ……………………………………………….. 122

Part 4 An Evaluation of the Distribution of HAB Species in UK and Irish Coastal Waters 124

4.1 Introduction ………………………………………………………..……………………………... 124 4.2 Methods ………………………………………………………………………………………….. 124 4.2.1 Nutrient data …………………………………………………………..……………………… 124 4.2.1.1 Riverine loadings 4.2.1.2 Winter nutrient 4.2.2 Phytoplankton data …………………………………………………………..……………….. 127 4.3 Statistics ………...……………………………………………………………..…………………. 128 4.4 Results ………………………………………………………………………………………….... 129 4.4.1 Statistical analyses ………………………………………...... 129 4.4.1.1 Data sets 4.4.1.2 Nutrient loadings and HAB species abundance 4.4.1.3 Ratios of nutrient loadings and HAB species abundance 4.4.1.4 Correlations between loadings and winter concentrations 4.4.1.5 Winter concentrations, ratios and HAB species abundance 4.4.1.6 Time-series analysis 4.4.2 The distribution of HAB species in UK and Irish coastal waters …………………………….. 138

ii 4.5 Discussion …………………………………………………………..…………………………..... 142 4.5.1 Introduction …………………………………………………………………………………... 142 4.5.2 Data sets and analysis ………………………………………………………………………… 142 4.5.3 Interpretation of results ………………………………………………………………………. 144 4.5.3.1 Introduction 4.5.3.2 HAB species abundance, nutrient loadings and winter concentrations 4.5.3.3 HAB species abundance and nutrient ratios 4.5.3.4 Time-series analysis 4.6 Conclusions …………………………………………………………………..…………………... 148

Part 5 Discussion and Synthesis 149 5.1 Introduction .……………………………………………..……………………………………….. 149 5.2 Ecohydrodynamic: Some General Principles ………………………………..…………………… 150 5.3 Ecophysiology: Phytoplankton Life Forms and Species Succession ………..…………………… 152 5.4 The Interaction between Ecohydrodynamics and Ecophysiology ….……..……………………… 154 5.4.1 Introduction ………..………………………………………………………..………………... 154 5.4.2 Small regions of restricted exchange ………..………………………………………………… 155 5.4.2.1 Introduction 5.4.2.2 Tolo and Victoria Harbour (Hong Kong) 5.4.2.3 A comparison between Tolo Harbour and small enriched RREs in UK coastal waters 5.4.3 Regional Seas ………..………………………………………………………..………………. 163 5.4.4 Summary ..………..………………………………………………………..…………………. 167 5.5 The Distribution of HAB Species in UK and Irish Coastal Waters ……………………………… 168 5.5.1 Introduction …………………………………………………………………………………… 168 5.5.2 Ecohydrodynamic conditions in UK and Irish coastal waters ..………………………………. 169 5.5.3 Species of Alexandrium ……………………………………………………………………….. 171 5.5.4 Species of Dinophysis ………………………………………………………………………… 172 5.5.5 The genus Pseudo-nitzschia ………………………………………………………………...... 173 5.5.6 Karenia mikimotoi ……………………………………………………………………………. 175 5.5.7 Prorocentrum minimum and P. lima ………………………………………………………….. 177 5.5.8 Lingulodinium polyedrum and Protoceratium reticulatum ………………………….……….. 179 5.5.9 Aquaculture and HABs ……………………………………………………………….………. 179 5.6 Synthesis ……………………………………………………………………………………..…… 182 5.6.1 Introduction …………………………………………………………………………………… 182 5.6.2 Does the occurrence of HABs imply eutrophication and is eutrophication always accompanied by HABs? ………………………………………………. 182 5.6.3 Has an increase in HABs been reported and is this increase real? ……………………………. 183 5.6.4 Does nutrient enrichment lead to more large-biomass HABs? ………………………………...183 5.6.5 Does nutrient enrichment lead to greater abundance of toxin producing species and hence an increase in low biomass HABs? ………………………………………... 184 5.6.6 Do shifts in nutrient ratios lead to more HABs? ……………………………………………… 185 5.6.7 Are toxin producing algae more toxic when nutrient ratios are perturbed in the sea? ………………………………………………………………………...... 186 5.6.8 Is the distribution of HAB species in UK and Irish waters related to niche requirements and ecohydrodynamics? …………………………………………………...186 5.7 General Conclusions ………………………………………………………………………...... 188

References

Annexes

Annex I Project partners and their affiliations Annex II Illustrations of phytoplankton species Annex III Acknowledgements

iii

Summary

Phytoplankton is the collective name for the microscopic organisms in lakes and seas. The word derives from the Greek phyton - plant and planktos - wandering, because these free floating plants are transported throughout the seas by currents and tides. Phytoplankton is as fundamental to life in the sea as grass or trees are to life on land. Like plants on land, all phytoplankters contain the green pigment chlorophyll which enables them to use the energy from sunlight to make organic matter from carbon dioxide, water and inorganic nutrients such as mineral salts of nitrogen and phosphorus (the process of carbon fixation called photosynthesis). These microscopic algae (micro-algae) therefore form the base of the marine food chain. Populations of individual species (of which there are ≈ 4000 world wide) are not fixed in time and space but are dynamic, particularly in coastal waters, and in many cases are highly seasonal. Most phytoplankters typically reproduce by binary division and consequently the normal pattern of growth involves an exponential increase in cells over a period of days. Under certain circumstances therefore, the abundance of phytoplankton as a whole or of one or more species in particular, can increase rapidly. Such an occurrence is often referred to as a ‘bloom’. Blooms are discrete events that can occur at any time during the production season and it is important to note that many, such as the spring bloom in temperate waters, are natural events. Some blooms of marine micro-algae (referred to as ‘Harmful Algal Blooms’ or HABs) can have a negative impact on the ecosystem and/or restrict the human use of the ecosystem. It is necessary to distinguish high (millions of cells L-1) and low (thousands of cells L-1) biomass HABs because their impact on ecosystem goods and services are often very different and there are different causal models to explain their occurrence. Bottom water deoxygenation leading to mortalities of fish and benthic organisms, and disruption to tourism due to water discolouration, foam on beaches and unpleasant odours are typically associated with high biomass HABs. The closure of shellfish harvesting areas because of elevated concentrations of shellfish toxins and mortalities of fish due to physical damage to gills are generally (but not uniquely) associated with low biomass HABs. Anthropogenic nutrient enrichment of coastal waters is often invoked as a reason for the occurrence of HABs. A link between harmful algal blooms and enrichment in some coastal waters is taken as evidence of a link in a wide range of coastal regions. This has led to the view that the occurrence of HABs diagnoses the undesirable consequence of anthropogenic nutrient enrichment and thus the occurrence of eutrophication as defined by the EC and OSPAR. A

iv number of assumptions are involved in this view, and there is a need to examine the associated scientific arguments and evidence if HABs and the occurrence of harmful algae are to be used as indicators of eutrophic conditions and counter-indicators of ecosystem health. This study examines some of the evidence and scientific arguments. The objectives of the project were to (i) review the scientific literature on the putative link between the occurrence and magnitude of HABs and anthropogenic nutrient enrichment of coastal waters and (ii) investigate the relationship between nutrients and HABs/ HAB species abundance by statistical analysis of data sets. It is important to separate the issue of eutrophication from the question of whether anthropogenic nutrient enrichment stimulates the occurrence (where none have occurred before), causes an increase in the frequency of occurrence, or promotes an increase in the duration or spatial extent of HABs. Blooms are discrete events and as such distinct from a more general increase in biomass and production fuelled by anthropogenic nutrient enrichment. The occurrence of HABs is not, in general, an indicator of eutrophication and HABs are not necessarily associated with eutrophication. However, an increase in HABs linked to anthropogenic nutrient enrichment may be one of several undesirable outcomes of the human driven eutrophication process. Based on a review of the scientific literature, it is evident that there is no consensus regarding the role of anthropogenic nutrients in stimulating the occurrence of HABs. Attempts to relate trends in HABs to nutrient enrichment are confounded by: the considerable spatial and temporal variability in naturally occurring HABs; the human mediated transport of HAB species between coastal regions; increased monitoring effort and the reporting of HABs; the influence of climate change (e.g. the North Atlantic Oscillation Index and the El Niño Southern Oscillation) on the occurrence of HABs and HAB species. For large biomass HABs, the hypothesis that nutrient enrichment can cause HAB is supported in some water bodies at the spatial scales of Tolo Harbour (Hong Kong) and the Seto Inland Sea of Japan but not in other water bodies with similar spatial scales (Carlingford Lough and the eastern Irish Sea). The global evidence for enrichment having brought about an increase in low biomass HABs of toxin producing species is more equivocal. To further examine the relationship between anthropogenic nutrient enrichment and HABs, data sets from coastal waters of the UK and Republic of Ireland were compiled and used to test the hypothesis that the occurrence of HABs and HAB species abundance increases with anthropogenic nutrient enrichment (proxy: riverine loading and mean winter concentrations of nutrients). On the basis of the statistical analysis carried out, the hypothesis is rejected and it is concluded that the UK and Irish data do not support the nutrient enrichment – HAB hypothesis.

v It is hypothesised that there is no single general hypothesis for changes in the occurrence of HABs but that their occurrence is the result of interactions between changes in specific pressures (including nutrient enrichment), the ecohydrodynamic conditions in particular water bodies and the adaptations of particular harmful algal species or life-forms. Rates of lateral exchange, mixing and dispersion within and between water bodies are considered to be one of the key determinants of algal blooms. A second crucial set of hydrodynamic characteristics involve the strength of vertical mixing and its consequences for stratification of the water column. Phytoplankton retained in near surface layers of stratified waters is well-illuminated throughout the year in tropical and subtropical waters and during spring and summer in temperate latitudes. Nutrient inputs to such layers (either natural, during , or anthropogenic, in urban waste water or enriched river discharges) are likely to stimulate algal blooms, unless planktonic animals or benthic filter-feeders consume the increased algal production. Distinct patterns are evident in the distribution of some HAB species in UK and Irish waters and it is concluded that the greater abundance of these species in waters to the west of Ireland and Scotland is the result of the intersection between the ecophysiology of individual phytoplankters and the ecohydrodynamic conditions in the water bodies in which they live. Thus, the seasonal development of thermo-haline stratification in coastal waters to the west of Ireland and Scotland favours the growth of dinoflagellates as the dominant life-form of pelagic primary producer. Advective processes such as downwelling at the coast serve to connect open shelf and coastal waters and thereby promote the transport of cells from populations that develop in seasonally stratifying offshore shelf waters. The Irish and Scottish coastal currents provide mechanisms for transporting populations along the coast. It is concluded that the occurrence of HABs and the abundance of HAB species should not be used to diagnose eutrophication unless a link to anthropogenic nutrient enrichment can be demonstrated. Furthermore, evidence of a link in one coastal region should not be taken as evidence of a general linkage in other coastal regions.

vi Part 1

Introduction

1.1 Introduction and Rationale

Over the last 25 years, a number of scientific publications have reported an apparent global increase in the occurrence of phytoplankton blooms, and incidents related to algal and cyanobacterial toxins; phenomena that can have a negative impact on ecosystem health1 and the human use of the marine ecosystem. This has led to a search for causes and anthropogenic nutrient enrichment of coastal waters is one that has received much attention. There is however no clear consensus. A link between harmful algal blooms and enrichment in some coastal waters is taken by some as evidence of a link in a wide range of coastal regions. This has led to the view that the occurrence of HABs can be used to diagnose the undesirable consequence of anthropogenic nutrient enrichment and thus the occurrence of eutrophication as defined by the EC and OSPAR2. A number of assumptions are involved in this view, and there is a clear need to examine the associated scientific arguments and evidence if HABs and the occurrence of harmful algae are to be used as indicators of eutrophic conditions and counter-indicators of ecosystem health. This study examined some of the evidence and scientific arguments. The objectives of this project were to: (i) review the scientific literature on the putative link between the occurrence and magnitude of HABs and anthropogenic nutrient enrichment of coastal waters; (ii) investigate the relationship between nutrients and HABs/ HAB species in UK and Irish waters by statistical analysis of data sets. With respect to the first objective the specific aims were to evaluate: (i) the role of increased nutrient availability and nutrient ratios; (ii) the importance of ecohydrodynamic3 conditions; (iii) whether nutrient/HAB relationships derived from other geographic areas can be applied to UK and Irish waters. To achieve objective two, the analysis of data sets was designed to determine whether there are: (i) coastal ‘hot-spots’ in the UK and Ireland which support large populations of HAB species; (ii) relationships between the abundance of HAB species and

1Ecosystem health is defined as the biological community’s vigour (energy flow) and organisation (structure) its resistance to disturbance and ability to recover from disturbance (resilience). See Tett et al. (2007). 2 OSPAR is the Oslo Paris Commission for the protection of the Marine Environment of the North Atlantic 3 Ecohydrodynamic refers to those features of the physical, chemical and biological environment to which the phytoplankters are adapted and which can differ between water bodies. - 7 -

enrichment (as nutrient loadings, winter concentrations of dissolved inorganic nutrients and their ratios). The title of this report makes it clear that the subject matter is micro-algae as distinct from macroalgae (seaweeds) which can sometimes form dense blooms in response to nutrient enrichment. Furthermore, the report focuses primarily on planktonic micro-algae but it is acknowledged that some harmful species of micro-algae are epiphytic or benthic. The literature review deals with species which cause harm through production of a large biomass as well as species known to produce biotoxins. The UK and Irish phytoplankton data are primarily from monitoring programmes undertaken to fulfil the requirements of the European Union, Regulation (EC) No 854/2004) for monitoring the occurrence of harmful phytoplankton species in shellfish cultivation and harvesting areas (OJEU 2004). The remainder of this introductory section describes the methodology used in the project, presents relevant aspects of phytoplankton ecology and considers the question ‘what is a micro- ’. The final part of this section discusses the effects of phytoplankton blooms on ecosystem health and the human use of the marine ecosystem, commonly referred to as ecosystem goods and services4 using examples from the scientific literature.

1.2 Methodology and Statistical Analysis

The work was undertaken by a group of experts (Annex I) who have collectively a broad knowledge of marine phytoplankton ecology, harmful algal blooms and toxin producing algae and who have contributed to the development of both regulatory and management policy and scientific understanding of harmful algal bloom dynamics. The report was based on a review of the relevant scientific literature and statistical analysis of nutrient and phytoplankton data derived from monitoring programmes in the UK and Ireland. The report has been fully supported by citations of the scientific literature and written in anticipation that it will be subsequently submitted for peer review and publication in an international scientific journal. A meeting of team members was held in Belfast on 19th June 2008 to agree the scope of the study and timetable. The literature review was undertaken by AFBI and a draft review written by R Gowen (AFBI) and P Tett (SAMS). Members of the team provided relevant data sets for statistical analysis which was undertaken by A Gordon (AFBI). A draft report on the results of the analysis was prepared by R Gowen and A Gordon (AFBI). Copies of the reports

4 Ecosystem goods and services (or ecosystem services) includes the ecosystems, together with the goods (e.g. fisheries) and services (e.g. waste assimilation) which humans derive directly and indirectly from ecosystem functioning (Costanza et al. (1997). - 8 -

were circulated to team members for review on 2nd December and the reports were reviewed and agreed at a meeting held in Belfast on the 16 and 17th December 2008. Data on the spring/ summer abundance of Alexandrium spp., Dinophysis spp., Pseudo- nitzschia spp., Prorocentrum lima, P. minimum, Karenia mikimotoi, Lingulodinium polyedrum and Protoceratium reticulatum were compiled from Irish and UK monitoring programmes and matched with nutrient data. For UK coastal waters, measured and modelled riverine nutrient loading data were also used in the analysis. These data referred to as RID (Riverine Inputs and Direct Discharges) are presented as annual loads per PARCOM area, based on monthly measurements. These reports have been available from 1992 onwards (see for example OSPAR 2001). For the UK, unmonitored areas account for ~39 % of the landmass and the issue of missing data needs to be addressed to avoid underestimation of the nutrient loads entering the marine environment. The original monthly data were therefore used to derive modelled loadings. The derivation process involved interpolation to cover missing data points, use of climatology (a standard year based on the available data) to fill in gaps in the data and correction factors to account for un-gauged areas. Linear regression analysis was used to identify relationships between HAB species abundance and anthropogenic nutrient enrichment. Riverine loadings (measured and modelled) and winter nutrient concentrations were used as proxies for the level of enrichment. The correlation coefficient was calculated to determine whether UK winter nutrient concentrations in coastal waters were related to riverine loadings (measured and modelled). Phytoplankton data from stations in Northern Ireland sea loughs (at which monitoring has been conducted for a minimum of 10 years) and PSP toxicity data from the northeast of England were examined to determine whether there were temporal trends in the data. The Mann-Kendall non-parametric test for monotonic trends (Hirsch & Slack, 1984; Hirsch et al. 1991) was used for this analysis. The relationship between the HAB time-series presented by Liu and Wang (2004) and the El Niño Southern Oscillation Index was also examined.

1.3 Phytoplankton

Phytoplankton is the collective name for the tiny organisms in lakes and seas. The word derives from the Greek phyton - plant and planktos - wandering, because these free floating plants are transported throughout the seas by currents and tides. A member of the phytoplankton is a phytoplankter. Phytoplankton is as fundamental to life in the sea as grass or trees are to life on

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land. Like plants on land, all phytoplankters contain the green pigment chlorophyll which enables them to use the energy from sunlight to make organic matter from carbon dioxide, water and inorganic nutrients such as mineral salts of nitrogen and phosphorus (the process of carbon fixation called photosynthesis). These microscopic algae (micro-algae) provide the food for small animals (zooplankton) which are food for larger animals such as larval fish. Thus, phytoplankton ultimately supports the fisheries (upon which humans depend) and the sea birds and sea mammals of our coastal waters. The species which make up the phytoplankton ( 4000 world wide, Sournia 1995) are commonly grouped into three functional categories of pelagic primary producers or lifeforms. These are diatoms, dinoflagellates and microflagellates although differentiation of dinoflagellates into a number of lifeforms may be appropriate as understanding of their complex nutrition increases. A more detailed discussion of lifeforms and species succession is presented in Part 5. Diatoms have a silicon cell wall (or frustule) which is divided into two halves, hence the name which is derived from the Greek ‘split in half’. The frustule often appears finely chiselled and may have spines or extensions (Annex II). All diatoms are photosynthetic. Dinoflagellates (Annex II) have two flagella with which they can swim short distances and from which their name is derived: Greek, dinos ‘whirling’ because of their swimming motion and the Latin word flagellum ‘whip’. There are two types of dinoflagellate, naked (without a cell wall) and armoured (processing a complex cell wall or theca made up of cellulose plates). Some dinoflagellates are photosynthetic, but others are mixotrophic5 and others are heterotrophic. Microflagellates are a taxonomically and nutritionally diverse group of small ( ≤ 20 µM) organisms which have one or more flagella (Annex II). Populations of individual species are not fixed in time and space but are dynamic, particularly in coastal waters. Smayda (1997a) has suggested that phytoplankton community dynamics are the consequence of cellular and population growth, each of which is controlled by a different suite of factors. Accordingly, at the cellular level, growth is the outcome of a balance between physiological fitness for growth and external control (e.g. nutrient and light availability). Population growth (and ultimately the size of a phytoplankter population) is, according to Smayda (1997a) the balance between cellular growth and environmental control.

5 Mixotrophs are autotrophic (fix carbon by photosynthesis) but are also capable of using organic matter. Heterotrophs require organic matter as a source of energy and nutrient elements. Any heterotroph which takes particulate food is a phagotroph. Saprotrophs use dissolved organic matter, perhaps obtained directly from within other organisms.

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The latter includes losses to advection/ dilution (discussed in more detail in Part 5) and herbivorous zooplankton grazing. The role of grazing in controlling HAB development is not discussed further in this report. In temperate-latitude seas, the onset and duration of the phytoplankton production season is controlled by the amount of light available (Sverdrup, 1953; Smetacek & Passow, 1990; Tett, 1990). This results in a pronounced seasonality (Figure 1.1A) which is largely determined by the cycle of solar radiation although, in waters deeper than  40 m, turbulent mixing which transports phytoplankton out of the illuminated surface waters also plays a role in determining the seasonal pattern of growth.

Figure 1.1 The seasonal cycles of A, phytoplankton biomass (as chlorophyll, mg m-3) and B, dissolved inorganic nitrogen (µM nitrate + nitrite,) in near surface (upper 20 m) offshore waters of the western Irish Sea. Nutrient data are from 2002 and chlorophyll data from 1992-2004. (Redrawn from Gowen et al. 2008). Winter nitrate (+ nitrite) converts to spring chlorophyll at 1-2 mg chlorophyll (mmol N)-1.

18 A 15 ) -3 12

9

6 Chlorophyll (mg m (mg Chlorophyll 3

0 01-Jan 22-Feb 14-Apr 05-Jun 28-Jul 18-Sep 09-Nov 31-Dec

10

9 B 8

M) 7  6

5 4

3 ( 2 1

0 01-Jan 22-Feb 15-Apr 06-Jun 28-Jul 18-Sep 09-Nov 31-Dec

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Figure 1.2 Cellular and population growth.

During the production season, it is the supply of mineral nutrients that largely determines how much growth occurs at the cellular and population level (Figure 1.2). Nutrients accumulate

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during the winter period when there is insufficient light for net growth, and the supply of nutrients (from remineralisation in the water column, efflux from sediments, run-off from land and atmospheric deposition) exceeds the demand by phytoplankton (Figure 1.1B). When the light climate improves in late winter/ early spring, these winter nutrients fuel spring growth and are incorporated into organic matter. In general these nutrients are not replenished and the low concentrations which prevail during the summer months constrain phytoplankton growth. It is widely accepted that in coastal waters, it is the availability of dissolved inorganic + - - nitrogen (ammonium, NH4 ; nitrate, NO3 ; nitrite, NO2 ) that is most likely to constrain phytoplankton growth (Ryther & Dunstan 1971) although diatoms (and silicoflagellates) require silicon (Si, as dissolved silica, Si(OH)4) for cell wall formation. Nitrogen limitation is the expected situation in northern European marine waters, although in some low salinity 3- environments such as the Baltic Sea, phosphate (PO4 ) is considered to be the limiting nutrient and P limitation has been assumed for the Eastern Mediterranean (Krom et al. 2004) and in the vicinity of the Pearl River estuary in China (Yin et al. 2001; Xu et al. 2008). The description and levels of nutrients and phytoplankton biomass (as chlorophyll) given above can be considered typical of UK and Irish shelf waters that seasonally stratify and are in a near pristine state. In shallow turbid estuaries or deep mixed waters, however, growth may be limited by light during the spring and summer months, thereby modifying the seasonal cycle and levels of biomass attained during the growing season. In other regions of the world, differences in the light climate and nutrient regimes result in different seasonal cycles of phytoplankton biomass and growth. The South China Sea, for example, is naturally nutrient poor (oligotrophic) and despite high levels of sub-surface irradiance, surface chlorophyll concentrations in offshore waters are  0.2 mg m-3 (Tang et al. 2004a). Similarly the Mediterranean Sea is regarded as being oligotrophic (Azov 1991; Ignatiades 1998). Some coastal regions are naturally enriched and productive. This is as a consequence of upwelling; a process whereby deeper nutrient rich water is entrained into surface illuminated waters and supports phytoplankton growth (see review by Smith 1968). Examples of coastal upwelling include coastal waters of the north western U.S. (Small & Menzies 1981), north west Africa (Jones & Halpern 1981) and Peru (Dengler 1985).

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1.4 Blooms of Micro-algae

1.4.1 Introduction

Most phytoplankters typically reproduce by binary division and consequently the normal pattern of growth involves an exponential increase in cells over a period of days. Under certain circumstances therefore, the abundance of phytoplankton as a whole or of one or more species in particular, can increase rapidly. Such an occurrence is often referred to as a ‘bloom’. The term 'bloom' has several meanings including: flowering (Gran & Braarud 1935) or blossoming of the organisms in the sea; a period of rapid increase in populations of some species; a period of abundant phytoplankton. In all these meanings there is the implication of a discrete event, an increase in abundance of one or more species, which stands out from what has happened before and the state to which the phytoplankton returns after a bloom. Phytoplankton blooms are natural events and part of the annual cycle of phytoplankton growth. For example, the spring growth of phytoplankton mentioned above is commonly called the spring bloom (Bigelow, 1926; Marshall and Orr, 1930; Mills et al. 1994; Gowen et al. 1995). This annual event is typical of many coastal waters in the temperate zone. It lasts for 2 to 3 weeks and is very important for the productivity (including the fisheries) of coastal waters. Following the spring phytoplankton bloom, nutrients become depleted in the near surface, well – illuminated, waters of the sea. The development of blooms later in the growing season therefore requires a replenishment of nutrients and summer blooms are often associated with a particular oceanographic feature or process that provides such a supply. One example of this is the summer blooms of the dinoflagellate Karenia mikimotoi6 at the Ushant tidal mixing front in the English Channel (Pingree et al. 1975). Tidal mixing fronts are the interface between tidally stirred and seasonally stratified waters and in this example the supply of nutrients is from the tidally stirred water (in which there is incomplete utilisation of nutrients because phytoplankton growth is constrained by the lack of light). In the Strait of Ombai7, Moore and Marra (2002) also observed elevated concentrations of chlorophyll (> 1 mg m-3 in these oligotrophic waters) occasionally extending 100’s of kilometres downstream of a frontal feature. Holligan et al. (1983) reported the regular occurrence of blooms (up to 8.5 x 106 cells L-1 in May 1982) of the coccolithophorid Emiliania huxleyi at the shelf break (edge of the continental shelf) front on the Celtic and Armorican shelf regions. Garcia-Soto et al. (1995) also report an E. huxleyi bloom of > 2.0 x 106 cells L-1 in the western English Channel in June 1992. In upwelling regions, such as

6 Formally know as Gyrodinium aureolum 7 Located between the Indonesian island of Alor and the northern coast of East Timor - 14 -

the southern Benguela upwelling system off the Atlantic coast of South Africa (see for example Armstrong et al. 1987 and references cited therein) nutrients are entrained into the euphotic zone and support enhanced phytoplankton production. In their study, Armstrong et al. (1987) measured chlorophyll concentrations of up to 25.2 mg m-3 at a depth of 10 m in the frontal zone. Human activities provide other sources of nutrients and the introduction of such anthropogenic nutrients into coastal waters can stimulate the productivity of planktonic algae in these waters. This could be good (more food for fish), but can also have undesirable effects on the ecosystem and human use of the ecosystem.

1.4.2 Red tides, nuisance, noxious and harmful blooms

On occasions and under the right conditions the abundance of cells can be so high that an algal bloom is visible to the human eye as a discolouration of the sea. Okaichi (1997) credited Okamura in 1916 with defining ‘red tides’ (red = ‘Akashio’ in Japanese) as a change in water colour due to outbreaks of microscopic plankton which sometimes causes the death of fish and other marine animals. The term has almost become synonymous with negative effects on the marine ecosystem (such as mass mortalities of fish, Rounsefell & Nelson 1966) but is rather vague and has been used for any large bloom whether it is red or not and even large diatom blooms have been referred to as red tides (Iwasaki, 1989). Despite this the term red tide is still widely used in SE Asia but has generally fallen out of use in Europe. Parker and Tett (1987) used the term ‘exceptional bloom’ to denote a large biomass bloom (> 100 mg chlorophyll m-3) and one which was not part of the regular seasonal cycle of phytoplankton. ‘Nuisance’ and ‘Noxious’ have also been used to describe both species and blooms which reportedly have a negative impact. Of the  4000 species of phytoplankton worldwide, some 300 are referred to as harmful (Sournia 1995) and blooms of these species are now widely referred to as Harmful Algal Blooms or HABs. The term ‘Harmful Algal Bloom’ and the acronym ‘HAB’ is part of the common language (jargon) amongst scientists working in this field and has been incorporated into the name of several international bodies and groups (e.g. GEOHAB, IPHAB, WGHABD8). In many ways the term is a misnomer, especially when used for the occurrence of low numbers of toxin producing species (see below). As suggested by Richardson (1997) in using the term, it

8 GEOHAB, Global Ecology and of Harmful Algal Blooms; IPHAB, the UNESCO, Intergovernmental Oceanographic Commission (IOC) Intergovernmental Panel on Harmful Algal Blooms; WGHABD, the ICES (International Council for the Exploration of the Sea)/ IOC Working Group on Harmful Algal Bloom Dynamics. - 15 -

is often the effect which is being referred to and not the abundance of the phytoplankter or phytoplankters involved. Smayda (1997b) argued that terms such as ‘nuisance’ and ‘harmful’ are subjective9 and that there is confusion over what a nuisance or harmful bloom is and whether such blooms should be defined by size (biomass) or effect. Richardson (1997) pointed out that:

“No common physiological, phylogenetic or structural feature has yet been identified that distinguishes “harmful” phytoplankton species from non- harmful and the scientific basis for treating harmful phytoplankton blooms as a distinct sub-set of algal blooms is not obvious.”

(see also Anderson and Garrison (1997). It is clear that a distinction needs to be made between harm in terms of a negative impact on ecosystem health and harm in terms of restricting the human use of the ecosystem (e.g. for fisheries and aquaculture, recreational activities, nature conservation). A further distinction needs to be made between harm which results from a large biomass (large biomass bloom) and harm which is induced by algal biotoxins or abrasive cell walls and spines. In general (but not exclusively) problems associated with biotoxins and physical damage to fish gills are associated with low biomass blooms, when the causative algae may be much less abundant than other phytoplankters and this does not accord with the idea of a bloom as a discrete event. Furthermore, a low biomass ( few hundred or thousand of cells L-1) would not necessarily require a source of additional nutrients (natural or anthropogenic) although large biomass blooms of toxin producing species can also occur which would require a nutrient supply. Fukuyo et al. (2002) made a similar distinction suggesting that two types of HABs were known in Japan:

“The first one is a noxious algal bloom associated with the mass mortality of marine organisms…..Most of the noxious blooms cause water discolouration, i.e. red tides…”. “The second type of HAB is a toxic algal bloom causing contamination of shellfish….”

Interestingly, Fukuyo et al. (2002) state that toxic blooms with toxin contamination levels lower than the level permitted by public health authorities are not recorded as HAB occurrences. In his study of the changes in frequency of HAB occurrence in the Seto Inland Sea, Okaichi (1989) distinguished between red tides associated with fish kills from other red tides but related changes in the occurrence of all red tide to pollution loading. In a study of HABs in the southern

9 ‘Nuisance’ is defined by the Concise Oxford Dictionary (8th edition, 1990) as “a person, thing, or circumstance causing trouble or annoyance” and defines ‘harm’ as “hurt or damage” and ‘harmful’ as “causing or likely to cause harm”. - 16 -

Benguela region (Southern Africa), Stephen and Hockey (2007) only considered HABs (i.e. blooms that had a harmful effect) in their analysis. It seems to us that under the right conditions any phytoplankter which blooms and reaches a sufficiently high biomass can impact ecosystem goods and services and might therefore be regarded as harmful. However, periodic deoxygenation of bottom water and mortalities of benthic organisms resulting from the senescence of a naturally-occurring bloom can be regarded as a natural perturbation much like a natural forest fire rather than a ‘harmful’ event. While a species may cause problems in one part of the world it may be benign or beneficial in other regions. This is exemplified by the diatom Skeletonema costatum10. In some northern European waters, this alga is a prominent component of the spring bloom which provides food for planktonic copepods and benthic organisms. Chen and Gu (1993) and Lin (1989) regarded S. costatum as a dominant red tide species. Similarly, Eucampia zodiacus, a common diatom that rarely exceeds an abundance greater than  103 cells L-1in UK waters, is regarded as a HAB species in Japan where it reaches 105 cells L-1 and has been associated with the indirect bleaching of the thallus (frond or leaf) of Porphyra (the Japanese ‘Nori’) through competition for nutrients (Nishikawa et al. 2007). Referring to all phytoplankton blooms (natural or otherwise) which result in an impact on or restrict the human use of the marine environment as HABs could misrepresent the scale of blooms stimulated by anthropogenic nutrient enrichment. For this reason it seems to us that a distinction should be made between natural blooms which may be beneficial, benign or inimical and ‘human induced blooms’ which impact ecosystem goods and services. Therefore, use of the term ‘Harmful Algal Bloom’ could be restricted to blooms that have unwanted consequences for human use of the ecosystem and which probably result from human activity especially nutrient enrichment. Such a distinction would remove some of the confusion surrounding the issue of anthropogenic nutrient enrichment and HABs. Furthermore, the idea that a HAB is a human induced event is consistent with the European Union Urban Waste Water Treatment Directive and OSPAR definitions of eutrophication in that to diagnose eutrophication, anthropogenic nutrient enrichment must promote the growth of phytoplankton to the extent that the biomass reached causes an undesirable disturbance which have a negative impact on ecosystem goods and services. In the case of toxin producing species which may involve low biomass, enrichment induced changes

10 The type species for Skeletonema costatum was found in Hong Kong waters; S. costatum s.s. is probably not present in NW European waters. - 17 -

in phytoplankton community structure which result in increased abundance of toxin producing species would represent an undesirable disturbance. For the purposes of this report we define a harmful algal bloom as:

“a discrete event associated with a 'bloom' of micro-algae or cyanobacteria that damages human use of ecosystem goods and services”

In this definition a ‘bloom' is defined as an increase in abundance relative to a normal background level which may be 1, low or high, depending on the organism.

1.5 Micro-algal Blooms and Their Effects

1.5.1 Introduction

HABs can often appear as unusual and spectacular discolorations of the sea that may also result in human health problems and economic loss at a local and regional scale. For these reasons HAB events are often emotively reported in the media. There is therefore a danger that HAB events are over reported and while acknowledging the effects that persistent and frequent HABs can have on ecosystem goods and services, it is important to keep a sense of perspective. Hallegraeff (1993) reported that worldwide  300 people a year die as a result of HAB related events. Hoagland et al. (2002) estimated the average annual cost (1987 – 1992) of HABs (in terms of the effects on public health, commercial fisheries and recreation/ tourism) in U.S. waters as $50 million. This compares to a gross national product of US$5 trillion generated by the coastal counties of the U.S. in 1995. Furthermore, Hoagland et al. (2002) report that in the U.S. morbidity resulting from the 6 leading food borne pathogens is between < 4 and > 7 million people per year with between 2,600 and 6,500 mortalities. In the UK, the last significant hospitalisations due to PSP were in 1968 (summarised by Ayres et al. 1982). The closure of shellfisheries is now strongly enforced by EU and UK regulation, and it is thought that this has successfully prevented almost all human illness due to algal biotoxins. Smayda (1997b) suggested at least eight mechanisms by which blooms of phytoplankters can have an impact on the ecosystem (Table 1.1). The wider effects of algal blooms on ecosystem health and influence the human use of the ecosystem are summarised in Figure 1.3. These effects are discussed in more detail in the remainder of this part of the report although at this point, the occurrence of HABs is not attributed to particular environmental conditions or e.g. climate change or anthropogenic nutrient enrichment.

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Table 1.1 Mortality modes and impact mechanisms of harmful species and their blooms (from Smayda 1997b).

Effect Species associated with effect Starvation Nutritional mismatch Size mismatch Excess prey density Mechanical “Bumping” Particle irritation Chaetoceros spp. Physical Viscosity barrier Karenia mikimotoi Gelatinous barrier Cerataulina pelagica Mucoid layer reduction Chattonella marina Anoxia Ceratium blooms Ammonium toxicity Noctiluca blooms Phycotoxins Direct vs. vectored toxicity Saxitoxin, Brevetoxin Hemolysins Cytotoxins Allelopathic11 Ambush predation Pfiesteria piscicida Unresolved Aureococcus anophagefferens

11 The inhibition of one plant by the release of a chemical by another. - 19 -

Figure 1.3 Algal blooms and their effects on ecosystem health and the human use of the marine ecosystem.

1.5.2 Ecosystem effects of micro-algal blooms

1.5.2.1 Pulse and press disturbance

Tett et al. (2007) suggested that ecologists distinguish episodic ‘pulse’ from sustained ‘press’ disturbances (see Bender et al. 1984). Local pulse disturbances are short-lived events and not considered to be a threat to ecosystem health and indeed may be an important part of natural ecosystem dynamics. Blooms can be spectacular in terms of their geographical extent and speed of development, and cause mortalities of other marine organisms. However, infrequent blooms fall into the category of pulse events and accordingly would not be considered a threat to ecosystem health. The 1988 Chrysochromulina polylepis bloom described below exemplifies this. In reviewing the long-term impact of the C. polylepis bloom (Gjøsaeter et al. 2000) asked the - 20 -

question ‘catastrophe or an innocent incident?’ On the basis of a comparison of coastal fish and benthic communities before, during and after the 1988 bloom, Gjøsaeter et al. (2000) found that populations of most organisms had recovered within months of the bloom. After one year there were few traces of the bloom’s impact and after 4 – 5 years all communities had recovered. Gjøsaeter et al. (2000) argued that seeding by planktonic larvae, and the short generation time of most coastal marine organisms, resulted in rapid recovery. They concluded that:

“even though algal blooms like the one in 1988 may reoccur, such perturbations are unlikely to leave any long-lasting profound effects”.

1.5.2.2 Shading and smothering

During the process of photosynthesis and the fixation of carbon by phytoplankton, light is absorbed by the pigment chlorophyll. The presence of high algal biomass (and chlorophyll) in the water column might therefore be expected to reduce the penetration of light with potential effects on the growth of submerged vegetation. Gallegos and Bergstrom (2005) reported a spatially extensive and large (≥ 107 cells L-1 and ≥ 300 mg chlorophyll m-3) bloom of Prorocentrum minimum in upper Chesapeake Bay during the spring of 2000, and estimated that the absorption due to phytoplankton at a wavelength of 676 nm increased from 0.92 m-1 at the start of the bloom to > 3 m-1 at the peak. As a consequence, light attenuation increased (up to eight fold) and Gallegos and Bergstrom (2005) concluded that the decrease in the depth to which sufficient light penetrated to support growth of submerged aquatic vegetation was the most likely reason for a decrease in submerged aquatic vegetation in 2000 relative to earlier years. However, blooms are transient events and the effects of shading would be equally transient and should not be confused with longer-term shading caused by an increase in suspended particulate matter. Perhaps the most striking example of the effects of smothering is that caused by blooms of Aureococcus anophagefferens. A bloom of this small (2-3 µm) picoplanktonic alga first occurred in Narragansett Bay in 1985, reached a maximum abundance of 1.2 x 109 cells l-1 and lasted for 5 months (Smayda & Villareal 1989). Blooms have reoccurred in subsequent years (Laroche et al. 1997) and apparently spread to other coastal regions of the eastern seaboard of the United States (Gobler et al. 2005). Blooms of A. anophagefferens (often referred to as brown tides) have been held responsible for a decline in submerged aquatic vegetation (eel grass) as a result of increased light attenuation (Laroche et al. 1997) and have had a significant effect on wild and cultured populations of bivalve molluscs. According to Bricelj and - 21 -

MacQuarrie (2007) brown tides have caused inhibition of feeding, inhibited growth and recruitment and caused mortality of several species of bivalve. Initially, these effects were thought to be caused by general smothering of the shellfish but Bricelj and MacQuarrie (2007) make reference to a toxic isolate of Aureococcus anophagefferens. Lomas et al (2004) cite Gainey and Shumway (1989) who suggest that A. anophagefferens may produce a dopamine- like toxin. With respect to coral reefs and the possible effects of shading and smothering, Szmant (2002) concluded that there was limited evidence for widespread decline in corals as a result of anthropogenic nutrient enrichment.

1.5.2.3 Deoxygenation

The consumption of organic matter by protozoans and bacteria as it settles to the seabed creates an demand. As a cosequence, in some coastal and shelf seas where the bottom water is isolated from surface waters (e.g. stratified water column) and there is a low rate of re-supply of oxygen, deep water may become naturally hypoxic or depleted in oxygen (e.g. Gillibrand et al. 1996). The rapid sinking of a large biomass bloom can have the same effect and in extreme cases affects fish and benthic organisms. A bloom (up to 0.5 x 106 cells l-1) of the armoured dinoflagellate Ceratium tripos caused deoxygenation of bottom water and mortality of shellfish in New York Bight during the summer of 1976 (Falkowski et al. 1980). Other examples of phytoplankton blooms and associated hypoxic events include a bloom (11.8 x 103 to 169.6 x 106 cells L-1) of Prorocentrum minimum in the lower Potomac River during which surface day time -1 oxygen was 1.3 mg O2 L and was associated with mortalities of fish (Tango et al. 2005).

1.5.2.4 Algal biotoxins

In addition to human health issues related to algal biotoxins (discussed below) it is evident that the toxins produced by a number of phytoplankton species can impact on other components of the marine ecosystem. During May and June 1988, a bloom of the prymnesiophyte Chrysochromulina polylepis caused widespread mortalities of a wide range of marine organisms in Scandinavian coastal waters. The bloom extended over an area of 75,000 km2 (Granéli et al. 1993) and reached a density of between 5 and 10 x 106 cells L-1 (Maestrini & Granéli 1991). C. polylepis produced a non selective toxin that affected membrane permeability and ion balance and was thought to have had an allelopathic effect on the growth of other phytoplankton (Maestrini & Granéli 1991). The bloom was held responsible for mortalities of molluscs, echinoderms, ascidians (sea squirts), cnidarians (jellyfish, anemones, hydroids), farmed and wild fish (Gjøsaeter et al. 2000). - 22 -

There is clear evidence of mortality of organisms feeding directly on toxin producing algae and transfer of toxins through the food chain. During the 1968 Alexandrium tamarense12 bloom off the north east coast of the UK (Ayres & Cullem 1978) there were coincidental mortalities of sand eels (Ammodytes spp.) and an estimated 80 % of the breeding population of shags (Phalacrocorax aristotelis) in Northumberland died (Adams et al. 1968; Coulson et al. 1968). White (1984) documented four cases of fish kills associated with paralytic shellfish toxins. In two of these, extensive mortalities of adult Atlantic Herring in the Bay of Fundy (Canada) in July 1976 and 1979 coincided with blooms of Alexandrium fundyense13 and in both cases there was sufficient toxin in the gut of the fish to have caused death. White (1984) concluded that zooplankton were the most likely vector for the transfer of the toxin from the alga. Mortalities of marine organisms occurred during a bloom (33 x 106 cells L-1) of Karenia brevisulcata in Wellington harbour (New Zealand) in mid-February to late March 1998 (Wear & Gardner 2001). According to these workers a toxin produced by this dinoflagellate caused widespread and almost total mortality of zooplankton, pelagic and demersal fish species and mortality of epibenthic and benthic invertebrates across all trophic levels. Scholin et al. (2000) report the deaths of over 400 California sea lions (Zalophus californianus) along the central Californian coast during May and June 1998. Coincident with these mortalities, a bloom of Pseudo-nitzschia australis was reported and domoic acid (a neurotoxin associated with amnesic shellfish poisoning (ASP) in humans, see below) was detected in planktivorous fish and in sea lion body fluids. Doucette et al. (2006) measured saxitoxin (a neurotoxin associated with paralytic shellfish poisoning (PSP) in humans, see below) in the faeces (up to 0.5 µg saxitoxin equivalents g-1 faecal material) of North Atlantic right whales (Eubalaena glacialis) during August/ September 2001 in the Bay of Fundy. At the same time, Doucette et al. (2006) measured saxitoxin in the herbivorous (feeding on phytoplankton) copepod Calanus finmarchicus, the main prey of these whales.

1.5.3 Micro-algal blooms and the human use of the ecosystem

1.5.3.1 Algal biotoxins and human health

Some species of phytoplankter produce chemical toxins. When the cells of these species are filtered by shellfish, the toxin can be retained by the shellfish and transferred through the food chain to humans and cause ‘shellfish poisoning’. The toxins are categorised by the type of toxic

12 Previously called Gonyaulax tamarensis. 13 Previously called Gonyaulax excavata. - 23 -

syndrome they cause: paralytic (PSP); diarrhetic (DSP); amnesic (ASP); azaspiracid (AZP); neurotoxic (NSP) shellfish poisoning; ciguatera fish poisoning (CFP). Illustrations of some of the species mentioned in this section are shown in Annex II. The genus Alexandrium14 are thecate (armoured) dinoflagellates, variable in size and usually found as single cells or in pairs in UK and Irish waters. The genus contains a number of species which are known to produce paralytic shellfish toxins (FAO 2004). There are toxic and non-toxic strains of some species, for example A. tamarense (Lilly et al. 2007; Higman et al. 2001; Medlin et al. 1998) and A. minutum (Lilly et al. 2005; Touzet et al. 2007). The toxins, saxitoxin and its derivatives, are potent neurotoxins and can cause headaches, nausea, and facial numbness and in severe cases respiratory failure and death. Ayres (1975) reviewed cases of poisoning associated with the consumption of mussels in the UK and concluded that for some of these cases (e.g. in Leith and Liverpool in 1827 and 1888 respectively) the medical reports were sufficiently detailed to assign the cause of illness to PSP. One of the first documented PSP events linked to a phytoplankter was in 1927 near San Francisco (U.S.) when 102 people became ill and 6 people died, although the causative species was not identified (Sommer & Meyer 1937). Azanza and Taylor (2001) cite Azanza (1999) as the source of results from a survey which indicate that globally, the armoured dinoflagellate Pyrodinium bahamense var. compressum was responsible for the greatest number (41 %) of Paralytic Shellfish Poisoning events between 1989 and 1999. This phytoplankter is known to produce saxitoxin (MacLean 1977) and can form spectacular bioluminescent blooms (Seliger et al. 1971). The first well documented occurrence of this dinoflagellate causing serious problems was in Papua New Guinea in 1972 when discolouration of the water occurred and three children were fatally poisoned (MacLean 1989a). This phytoplanker has caused severe economic and health problems in the Philippines, Malaysia, Brunei and Indonesia. The Philippines has been the most severely affected with 1,995 cases and 117 deaths linked to PSP toxicity between 1983 and 1999 (Azanza & Taylor 2001). Diarrhetic shellfish poisoning was first linked to the presence of Dinophysis fortii in Japan (Yasumoto et al. 1980) and to D. acuminata in Dutch coastal waters (Kat 1983). During 1989, human illness was recorded after consumption of mussels containing DSP toxins from the Northern Adriatic coast (Boni et al. 1992). To date, seven species of the genus Dinophysis are thought to produce okadaic acid or its derivative dinophysistoxins which cause DSP (Lee et al. 1989). The dinoflagellate Prorocentrum lima is also known to produce okadaic acid (Koike et

14 Previously called Gonyaulax. - 24 -

al. 1998). Symptoms of DSP in humans include diarrhoea, nausea and vomiting but no human fatalities have been associated with DSP toxins. For regulatory purposes, pectenotoxins and yessotoxins are classified within the DSP group. Pectenotoxins are produced by some of the Dinophysis species including D. acuta and D. acuminata (MacKenzie et al. 2005), and can cause liver and heart disease in humans. Yessotoxins induce similar symptoms but are produced by the dinoflagellates Lingulodinium polyedrum15 and Protoceratium reticulatum16 (Paz et al. 2004 and references cited therein). The neurotoxin domoic acid is produced by various species of Pseudo-nitzschia and Nitzschia. Domoic acid was identified as the toxin responsible for causing an outbreak of poisoning in humans after the consumption of blue mussels from Prince Edward Island (Canada) in 1987. During this incident, 107 illnesses and 3 deaths were attributed to the toxin (Todd 1990). Symptoms of ASP poisoning in humans include short and long-term memory loss. The pennate diatom Pseudo-nitzschia pungens17 was identified as the causative species in this case (Bates et al. 1989). The toxins known as azaspiracids were first identified in mussels from Ireland in 1995 (McMahon & Silke 1996). The azaspiracids belong to a novel group of polyether toxins which cause symptoms similar to those displayed by DSP (Twiner et al. 2008). A small (11-15 µm) armoured dinoflagellate named Azadinium spinosum has recently been identified as a producer of azaspiracids (Tillmann et al. 2009). Brevetoxin is the collective name given to toxins which cause NSP. Brevetoxins have caused considerable problems in Florida and the Gulf of Mexico and are primarily produced by the naked dinoflagellate Karenia brevis. Blooms of K. brevis have caused water discolouration and large scale fish kills (Magaña et al. 2003) and human illness (Kirkpatrick et al. 2004) in the Gulf of Mexico, Florida and North Carolina coastal regions although NSP has not been linked to fatalities in humans (van Dolah 2000). Magaña et al. (2003) present a chronology of K. brevis red tides in the western Gulf of Mexico where records of their occurrence date back to1648. In Florida, toxic red tides have been reported since the 1840s (Kirkpatrick et al. 2004). In 1987, an extensive red tide of K. brevis along the North Carolina coast resulted in 48 cases of shellfish poisoning after the consumption of oysters (Morris et al. 1991). A recent study demonstrated that exposure to a red tide of Karenia brevis resulted in increased reports of respiratory problems in asthmatics caused by inhalation of toxin in the form of an aerosol

15 Previously named Gonyaulax polyedra 16 Previously named Gonyaulax grindleyi 17 Previously named Nitzschia pungens - 25 -

(Milian et al. 2007). Similarly, in the Mediterranean, people visiting beaches have suffered from fever, conjunctivitis and respiratory problems. In 2005 for example, several hundred people who had spent the day on a beach in Genoa were hospitalised. These events were linked to an aerosol containing a toxin which was considered to be from the benthic dinoflagellate Ostreopsis spp. (Ciminiello et al. 2008). Closely related to brevetoxins are a group of toxins known as ciguatoxins (Naar et al. 2007). Ciguatoxins are produced by the benthic dinoflagellate Gambierdiscus toxicus and are transferred through the food chain by the consumption of tropical and subtropical fish. It is mostly confined to areas of the Pacific Ocean, Caribbean Sea and the Western Indian Ocean. Ciguatera poisoning is thought to affect 25,000 people annually (Terao 2000).

1.5.3.2 Mucilage and human health

Large areas of the northern Adriatic Sea were covered with gelatinous material or mucilage from June to September 1988 and June to August 1989 (Giani et al. 1992). The mucilage appears to originate from marine snow (aggregates of phytoplankton cells and detrital material) which accumulates at the pycnocline, aggregates and floats to the surface (Herndl et al. 1992). The human health concerns associated with widespread occurrence of mucilage were considered by Funari and Ade (1999) who concluded that direct effects of the mucilage on human health have not been observed but cite World Health Organisation reports which suggest that by preventing people from bathing in the sea, mucilage events impact on recreation and tourism, and therefore have a negative effect on good human health and well being.

1.5.3.3 The impact of micro-algal blooms on recreation and tourism

Discolouration of the water, surface slicks of mucilage, foam and unpleasant odours (which may be produced during the decay of large biomass blooms stranded on the shoreline) all have a negative effect on the aesthetic quality of the coastline and coastal waters. Species of Phaeocystis have a colonial life stage during which large numbers of cells are embedded in a mucilaginous matrix up to 2 cm in size (Davidson & Marchant 1992 and references cited therein). During Phaeocystis blooms, agitation of the water by wind and wave action can break up the colonies and the mucilage creates large quantities of foam. Lancelot et al. (1987) for example, reported foam caused by Phaeocystis pouchetii of up to 2 m thick on Dutch and German beaches. This is one reason why species of Phaeocystis are considered to be HAB species (but see Cadée & Hegeman 2002).

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1.5.3.4 The economic impact of blooms on Fisheries and aquaculture

Algal blooms and the occurrence of toxin producing algae continue to have a major economic impact on fisheries and aquaculture throughout coastal waters of the world. In addition to financial loss associated with periodic prohibition on harvesting of cultured and wild shellfish because of biotoxins, algal blooms cause financial loss through mortality of stock. There are many reports and scientific publications and the intent here is not to present an exhaustive list but provide well documented examples from different regions of the world which illustrate the range of effects and scale of impact in economic terms. Blooms of Phaeocystis are reported to clog nets and interfere with (Chang 1983). The study of Savage (1930) is widely cited as evidence of the way in which blooms of Phaeocystis can interfere with fishing. However, in considering the influence of Phaeocystis on fisheries, Savage (1930) concluded that this phytoplankter has two effects:

“(i.) It may partially or quite ruin a by acting as an almost impassable barrier to the shoaling of herring on the usual fishing grounds.

(ii.) It may actually divert more herring to the fishing ground.”

It is noteworthy that Savage (1930) also stated that:

“These observations are only put forward as suggestions and not as proved facts. Enough is not known about the migrations of the herring……It does seem probable, however, that the presence of these “weedy water” patches is one of the factors influencing the migrations of the herring, and is worthy of further investigation.”

In addition to nuisance caused by foam on beaches, species of Phaeocystis have also been associated with mortalities of farmed fish. In September 1997 for example, a bloom of P. globosa in Quanzhan Bay, Fujian province (China) extended over an area of 3,000 km2 and caused major losses of farmed fish estimated as US$ 7.5 million (Qi et al. 2004). Commercial fishing for the Bay Scallop (Argopecten irradians) in coastal bays of Long Island has suffered because of a major reduction in the fishery as a result of mortality attributed to brown tides of Aureococcus anophagefferens. The economic loss has been estimated as US$ 3.3 million per year (Gobler et al. 2005). A number of phytoplankton species have been associated with mortalities of cultured fish in different regions of the world. Blooms of the naked dinoflagellate Karenia mikimotoi have caused mass mortality of farmed fish in Norway (Tangen 1977, Dahl & Tangen 1993), Ireland - 27 -

(Raine et al. 1993), Scotland (Jones et al. 1982; Davidson et al. (2009), China (Hong Kong) Qi et al. (2004) and South Korea (Kim 1997). Many of the early reports of mortalities of benthic animals and farmed fish were attributed to deoxygenation but Roberts et al. (1983) reported toxin like damage to fish gills and cytotoxins have been identified from K mikimotoi (see also Satake et al. 2005). Blooms of Cochlodinium polykrikoides have been reported from several coastal regions of Japan including a large bloom in 2000 which caused losses of  US$ 36.4 million (Kim et al. 2004). This species has been associated with fish kills and mortalities of coral organisms in pacific coastal waters of Costa Rica (Vargas-Montero et al. 2006) and mortalities of fish in South Korea (Kim 1997). Blooms of microflagellates have also caused mortalities of farmed fish throughout the world. In Northern Europe, Kaartvedt et al. (1991) reported a bloom of the toxin producing Prymnesium parvum in a Norwegian fjord system which caused mortalities of farmed fish with an economic loss of US$5 million. Microflagellate (unknown identity) blooms have been linked to mortalities of farmed fish in Scotland (Gowen 1987). In Japan, blooms of several microflagellate species have been associated with fish kills. Chattonella antiqua blooms have frequently caused fish kills in Japan and in 1972, a bloom was associated with the mortality of  14.2 million Yellowtail and an economic loss of US$70 million (Okaichi 1989). Blooms of Heterosigma akashiwo caused serious damage to the aquaculture industry in the Seto Inland Sea during 1975 and 1981 (Yamochi 1989). This phytoplankter has also caused mortalities of farmed fish in coastal waters of the Canadian Pacific (Black et al. 1991 and references cited therein) and in New Zealand (MacKenzie 1991). Chaetoceros species are a common component of the marine flora in coastal waters throughout the world usually without harmful effects but occasional blooms have resulted in losses of farmed fish. In Scottish coastal waters, mortalities of farmed salmon were linked to blooms of Chaetoceros wighami in Loch Torridon (Bruno et al. 1989) and Chaetoceros debile in the Shetland Isles (Treasurer et al. 2003). Species of Chaetoceros have also been associated with mortalities of farmed fish on the pacific coast of Canada (E Black, pers. comm.). The cause of death is presumed to be gill damage (from the siliceous spines of the cell wall of Chaetoceros spp.) leading to asphyxiation.

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Part 2

An Overview of Harmful Micro-algal Blooms and HAB Species in Coastal Waters of the United Kingdom and Republic of Ireland

2.1 General Introduction

Historically the recording of harmful algal blooms in UK and Irish waters has been localised with a lack of a systematic format for recording events within and between countries. Programmes initiated in the late 1980s and 1990s to protect human health mean that toxin producing algal species and toxicity events are now more regularly reported at national and international (ICES, IOC) levels and in the scientific literature. The following is based on a review of the scientific literature, the grey literature and expert opinion18 within the group. Where available, details of species abundance during HABs and the species affected are given.

2.2 Coastal Waters of the United Kingdom

2.2.1 Introduction

After the 1968 PSP outbreak in the north east of England (see below) a biotoxin monitoring programme was established in England. Initially restricted to monitoring for the presence of PSP between April and September, the programme has gradually expanded to include biotoxin analysis and monitoring of toxin producing phytoplankton from all commercial shellfish production areas. In 1993, a programme was established in Wales to monitor for PSP and DSP toxins between March and September. In 1968 a small monitoring programme was introduced in Scotland. This was revised in 1991 to include testing for PSP and DSP toxins at 65 monitoring sites. In 1996, a phytoplankton monitoring programme was initiated followed by the introduction of testing for ASP in 1998. In Northern Ireland regular testing of commercial shellfish beds was carried out for the presence of

18 AFBI undertakes statutory monitoring of phytoplankton in Northern Ireland waters on behalf of the Food Standards Agency in Northern Ireland (FSA(NI)); Cefas undertakes similar monitoring in England and Wales on behalf of the Food Standards Agency in the UK (FSA(UK)); SAMS (and previously , Marine Laboratory) undertakes similar monitoring in Scotland on behalf of the Food Standards Agency in Scotland (FSA(S)); The Marine Institute in Galway (Ireland) undertakes regulatory monitoring on behalf of the Food Safety Authority of Ireland (FSAI) and the Sea Fisheries Protection Authority (SFPA).

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PSP toxins during the 1970’s and 1980’s (McCaughey &Campbell 1992). Following a series of negative results, testing ceased in the mid 1980s but was re-established again in 1990 (McCaughey & Campbell 1992). A more comprehensive monitoring programme was established in the 1990s and included monitoring for toxin producing phytoplankton. Testing shellfish tissue for DSP was introduced in 1992 using a rat bioassay but this was replaced by the mouse bioassay in January 2001. Testing for ASP was introduced for farmed scallops in 1997 and for wild scallops in September 1999. In the UK, the current monitoring of toxins in shellfish tissue and phytoplankton in the vicinity of commercial shellfish production areas is undertaken to fulfil the requirements of the European Union, Regulation (EC) No 854/2004 (OJEU 2004). A map of the UK showing the locations of places mentioned in the text is shown in Figure 2.1.

2.2.2 Species of Alexandrium

Species of Alexandrium (Annex II) are small to medium thecate (armoured) dinoflagellates. A. tamarense is 22 - 44 µm long and 20 – 36 µm wide (Dodge 1982); A. minutum is 15 -29 µm long and 13 – 21 µm wide (Taylor et al. 2003). Lebour (1925) provided the original description of Alexandrium tamarense (originally called Gonyaulax tamarensis) using cells collected from the Tamar estuary in Devon (UK). Toxicity in shellfish is frequently associated with low abundance (a few hundreds of cells per litre) of these dinoflagellates (i.e. they are low biomass HAB species) although they can also form large biomass blooms. For the UK national monitoring programmes, a threshold abundance (action level) of presence (20 cells L-1 the limit of detection assuming a settling volume of 50 ml, or 40 cells -1 for a 25 ml settling volume for microscopic analysis) has been set. The existence of toxic and non-toxic strains of the same species (Medlin et al. 1998; Higman et al. 2001) may be the reason why areas such as the Orkney and Shetland Islands demonstrate toxicity at low cell levels whilst dense blooms of Alexandrium spp. (> 10 x 106 cells L-1) along the south coast of England have little or no toxicity associated with them (Bresnan et al. 2007). The distribution of toxin producing A tamarense (Group I) described by Lilly et al. 2007 appears to be restricted to Scottish waters (Higman et al. 2001, John et al 2003, Collins et al. 2009; Brown et al. in press) although cysts of this strain have also been found in Belfast Lough (Neale et al. 2007). The non toxic A. tamarense (Group III) described by Lilly et al. (2007) is widespread around the UK coast ( Higman et al. 2001, John et al. 2003, Collins et al. 2009, Brown et al in press). It can form very high biomass blooms along the south coast.

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Reports of paralytic shellfish poisoning are rare in the United Kingdom. Based on a re- evaluation of previous accounts, Ayres (1975) concluded that between 1827 and 1968 there had

Figure 2.1 A map of the United Kingdom showing locations mentioned in the text. 1, ; 2, Farne Islands; 3, Northumberland; 4, Trow Rocks; 5, Staithes; 6, Southampton; 7, Plymouth; 8, Penzance; 9, Belfast 10, ; 11, Loch Fyne; 12, Loch Striven; 13, Loch Torridon.

been ten incidents involving PSP poisoning with approximately 14 fatalities. The first well documented case of PSP intoxication in UK waters was in 1968 when 78 people showed clinical

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symptoms of PSP toxicity after consuming mussels from an area in the North East of England (Ayres & Cullem 1978). There were coincident reports of dead sea birds and sand eels (Adams et al. 1968). Coulson et al. (1968) estimated that 80 % of the breeding population of shags (Phalacrocorax aristotelis) in Northumberland died during this event, which was significantly greater than any other species. The species responsible for the initial PSP poisoning event was identified as Gonyaulax tamarensis (var excavata) (Wood 1968), later reclassified as Alexandrium tamarense (Lebour) Balech (Balech 1995). Robinson (1968) used samples taken from Continuous Plankton Recorder (CPR) surveys to map the spatial distribution of A. tamarense during the period April-June 1968. A. tamarense cells were first detected in mid April to the east of the Firth of Forth and appeared to spread southwards until maximum concentrations were observed 16-24 km offshore between Eyemouth and the Farne Islands during mid May (Ayres & Cullem 1978). Areas of discoloured water were also reported from various sites in the region during this time with a maximum abundance of 74,000 A. tamarense cells L-1 recorded from a site off Staithes (Ayres & Cullem 1978). Alexandrium spp. regularly occur in numerous locations along the south and south-west coasts of England where concentrations in sheltered estuaries can reach levels in excess of 10.0 x 106 cells L-1 (FSA(UK) unpubl. data). In 1990, widespread toxicity was recorded in coastal areas of the North East of England and in Belfast Lough, Northern Ireland (Wyatt & Saborido-Rey 1993). The maximum recorded level of toxicity (3,647 µg saxitoxin (100 g shellfish tissue)-1) was in mussels from Trow Rocks on the north east of England (ICES C.M.1991/Poll:3). Between 1968 and 1990 there were no clinical cases of PSP although threshold levels were breached 17 times (Wyatt & Saborido-Rey 1993). In coastal waters of Northern Ireland, Alexandrium spp. occur infrequently and rarely exceeds 200 cells L-1. The exception to this was a bloom (~ 1 x 106 cells L-1) of A. tamarense in Belfast Lough in 1996 which was associated with toxicity in farmed mussels (Mytilus edulis) (FSA(NI) unpubl. data.). PSP levels exceeding the regulatory limit occur on an almost annual basis in Scotland (Turrell et al. 2007). The most seriously affected regions include the Orkney and Shetland Islands, Firth of Forth and the Scottish east coast where relatively low numbers of Alexandrium spp. (below 2,000 cells L-1) can cause toxicity in shellfish (Bresnan et al. 2007) and closures of regulated harvesting areas. In 1991, a PSP outbreak in Orkney extended into 1992 with prolonged closures of shellfisheries in the region (ICES C.M.1992/Poll:4).

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A. minutum from the Fal estuary (south west coast of England) has been identified as a PSP toxin producer (Percy 2006) while A. minutum from Scottish waters was not observed to produce PSP toxins under the culture conditions used (Brown et al. in press). A. ostenfeldii, identified from along the south coast and in Scottish waters, was observed to produce both spirolide and PSP toxins (the latter at low concentrations) (Percy et al. 2006; Brown et al. in press). The recently described A. tamutum (Montresor et al. 2004) has been identified along the east coast of Scotland (Neale et al. 2007; Alpermann et al. 2008) as well as in waters around Orkney and Shetland (Brown et al. in press). No PSP toxicity was identified in this species. The UK represents the most northly record of this species to date.

2.2.3 Species of Dinophysis

Species of Dinophysis (Annex II) are armoured dinoflagellates that range in size from small to medium and large robust species (D. acuminata 38 – 58 µm; D. acuta 54 – 84 µm; D. norvegica 48 – 67 µm, Dodge 1982). To date, seven species of Dinophysis are thought to produce okadaic acid or its derivative dinophysistoxins which cause DSP (Lee et al. 1989). Some species of Dinophysis (including D. acuta and D. acuminata) produce pectenotoxins (MacKenzie et al. 2005). Toxicity in shellfish is often associated with low abundance (a few hundreds – thousands of cells per litre) of Dinophysis, although large biomass HABs of these species can occur. For the UK national monitoring programme, the threshold abundance has been set at 100 cells of Dinophysis spp. L -1, above which additional samples are collected for phytoplankton analysis and shellfish tissue testing. Species of Dinophysis detected in UK coastal waters include D. acuminata, D. acuta, D. norvegica, D. fortii, D. dens, D. hastata and D. rotundata (Bresnan et al. 2007). In Scottish waters, Dinophysis spp. have been recorded as a regular component of the phytoplankton but normally at low levels (Tett & Edwards 2002). One exception to this was a bloom (0.94 x 106 cells L-1) of D. acuminata in Loch Long in July 1992 which caused discolouration of the water and toxicity in shellfish (MacDonald 1994). A number of incidents linked to DSP toxins have been recorded in Scottish waters. These include an extended period during 2001 along the west coast. The dominant species recorded in this period was D. acuta which reached a maximum abundance of 8,040 cells L-1 in August 2000 (ICES 2001/C:04). In June 2006, 171 people became ill after consuming mussels from a Scottish site. DSP toxins were confirmed as the cause and Dinophysis spp. as the causative genus. In 2007, an extended period of DSP toxicity

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(June to August) was recorded in shellfish from the Shetland Islands. Cell concentrations of Dinophysis spp. reached 5,300 cells L-1 in this region during July (ICES 2008/OCC:03). A number of human health problems have been associated with DSP in England and Wales but there are no reports in the scientific literature linking Dinophysis spp. to these incidents, although D. acuta and D. norvegica have been detected in high numbers off the North East coast of England in the 1990s (Bresnan et al. 2007). Cells of Dinophysis spp. are also recorded in coastal waters of the Isle of Man (T. Shammon pers comm.) and Northern Ireland. Diarrhetic Shellfish Toxins (DST’s) were detected in mussels from Belfast Lough during a bloom of D. acuminata in 1994 (FSA(NI) unpubl. data).

2.2.4 Species of Pseudo-nitzschia

Approximately 20 species make up the diatom genus Pseudo-nitzschia. Cells are needle shaped, range in size from 50 to 150 µm in length and form chains sometimes in excess of 1 cm (Annex II). These phytoplankters have a worldwide distribution and toxicity in shellfish is typically associated with medium biomass HABs (≥ 105 cells L-1). For the UK national monitoring programme, the threshold abundance (action level) has been set at 150,000 cells -1 for England and Wales and Northern Ireland and 50,000 cells L-1 for Scotland. Discrimination of Pseudo- nitzschia to the level of species by light microscopy is not possible. This has lead monitoring programmes to categorise cells by size and shape and as either ‘seriata’ type (diameter > 3µm) or ‘delicatissima’ type (diameter < 3µm). In Scottish waters, this demarcation has been shown to be of particular use as toxin producers have been identified within the former group (Fehling et al. 2006). There is also a distinct seasonality in the occurrence of these groups in Scottish waters with the spring bloom dominated by the ‘delicatissima’ type and the later summer/ autumn blooms dominated by the ‘seriata’ type (Fehling et al 2006). Species of Pseudo-nitzschia are a regular component of the phytoplankton community in coastal waters of the UK. Investigations into the diversity of this genus using transmission electron microscopy has identified 13 species in UK waters; P. americana, P. australis, P. caciantha, P. cf. calliantha, P. cuspidata, P. decepiens, P. delicatissima, P. fraudulenta, P. pungens, P. pseudodelicatissima, P. multiseries, P. seriata and P. subpacifica (Fehling et al. 2006, Bresnan et al. 2007). Three of these species have been confirmed as domoic acid producers in UK waters; P. australis and P. seriata in Scotland (Fehling et al. 2004) and P. multiseries from the English Channel (Percy 2006).

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A number of Pseudo-nitzschia blooms have been reported in Scottish waters. For example, in March/ April 2004, blooms of Pseudo-nitzschia delicatissima were observed and later in that year (August/September) Pseudo-nitzschia seriata blooms were recorded (ICES 2005/C:03). In April 2007, blooms of Pseudo-nitzschia spp. were reported and in early July a bloom of Pseudo- nitzschia spp. (2.5 x 106 cells L-1) was recorded in south west Shetland (ICES 2008/OCC:03). The most significant ASP incident to occur in Scottish waters was in 1999 when a large Pseudo-nitzschia spp. bloom (2.3 x 10 6 cells L-1) resulted in a temporary prohibition on scallop fishing due to high levels of domoic acid in tissue samples (Gallacher et al. 2000). There have been few records of domoic acid in blue mussels (Mytilus edulis) over the closure limit. Pseudo-nitzschia spp. are also widespread and found regularly in samples collected from around the coasts of England and Wales. They occasionally occur in concentrations that exceed the action level. During the period 2000 to 2005, Pseudo-nitzschia spp. exceeded the action level on one occasion but between June 2006 and March 2007, the threshold abundance was exceeded on 32 occasions (at eight sites) along the south-west coast of England. This increase coincided with increased monitoring effort.

2.2.5 Karenia mikimotoi

Karenia mikimotoi is a medium (24 - 40 µm long and 17 – 32 µm wide, Dodge 1982) photosynthetic naked dinoflagellate (Annex II). Mortalities of benthic animals and farmed fish are associated with large biomass (≥ 106 cells L-1) blooms of K. mikimotoi. This phytoplankter is recorded as part of the regulatory monitoring of toxin producing phytoplankters in Northern Ireland but not in England and Wales or Scotland. A number of K. mikimotoi blooms have been recorded in UK waters including the Eastern Irish Sea in 1971 (Helm et al. 1974), 1975 (Evans, 1976 cited in Ayres et al. 1982) and 1976 (Evans 1979) and along the south western coast of England in 1978 (Griffiths et al. 1979) and more recently the west coast of Scotland (Davidson et al. 2009). The K. mikimotoi bloom (0.5 to 5.2 x 106 cells L-1) in the eastern Irish Sea in Autumn 1971, caused mass mortalities of lugworm (Arenicola marina) and some deaths of the heart urchin Echinocardium cordatum (Helm et al. 1974). There were also reports of A. marina mortalities following a late summer/autumn bloom (0.92 x 106 cells L-1) in the eastern Irish Sea in 1975 (Evans, 1976 cited in Ayres et al. 1982) the following year a bloom was associated with localised mortalities of plaice (P. platessa), eels (A. Anguilla) and lugworms in the same area (Evans 1979).

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Griffiths et al. (1979) report that the sea urchin (Echinus esculentus), Devonshire cup coral (Caryophyllia smithii) and the spiny star fish (Marthasterias glacialis) were the most impacted species during a K. mikimotoi bloom in waters off Penzance in August 1978. In 1982, a bloom in the English Channel, off the south west coast of England, was reported to have caused fish and invertebrate mortalities (Ayres et al. 1982). In general, Karenia mikimotoi is found as a regular component of the phytoplankton in Scottish waters normally reaching only a few thousand cells per litre (Davidson et al. 2009). In 1980, however, a bloom (20.0 x 106 cells L-1) caused the deaths of farmed salmon in shore tanks on the shore of Loch Fyne (Jones et al. 1982). There then appears to have been a period of ~ 18 years (1981 - 1998) during which there were no reports of significant K. mikimotoi blooms in Scottish waters, although Gowen et al (1998) recorded an abundance of 0.3 x 106 cells L-1 at the Islay front on the inner Malin shelf in August 1995. According to Davidson et al. (2009), large K. mikimotoi blooms occurred in Scottish waters in 1999, 2003 and 2006. The first of these was in coastal waters of the Orkney Islands. The 2003 bloom (18.0 x 106 cells L-1) around the Orkney and Shetland Islands was responsible for the deaths of 53,000 farmed fish from four sites in the Shetland Isles. The 2006 bloom was the most extensive to date and covered an area that extended from the island of Mull on the west coast to the Shetland Isles and the north east coast (Stonehaven). Abundance over the region varied but reached 3.7 x 106 cells L-1 in a sample from in the Orkney Isles in mid August. Although no major fish kills were reported there were reports of mortalities of benthic organisms including lugworm, blue mussel (Mytilus edulis), common starfish (Asterias rubens) and king scallop (Pecten maximus). According to Davidson et al. (2009) there were also reports from the public of mortalities of crab and lobster and fish including sea scorpion (Myoxecephalus scorpius) and conger eel (Conger conger).

2.2.6 Other HAB species

Prorocentrum minimum is a small (14 – 22 µm long and 10 – 15 µm wide, Dodge 1982) armoured dinoflagellate with a worldwide distribution. This phytoplankter was initially believed to be the causative organism for ‘venerupin poisoning’ but the validity of this toxin syndrome has been discredited (Heil et al. 2005). The toxicity of P. minimum remains to be elucidated. Most clones examined appear to be non toxic, however clones isolated from the French Mediterranean and Japan have been shown to cause PSP and neurotoxic symptoms in mice (Heil et al. 2005). The toxicity of P. minimum in UK waters has yet to be confirmed although further

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studies are required. No threshold abundance (action level) has been set for this species in the UK. Blooms of P. minimum have been reported from coastal waters of Northern Ireland (FSANI unpubl. data), the east coast of Scotland, Shetland Islands and from shallow ponds on the south coast of England (Bresnan et al. 2007). In 2007, a bloom (2.4 x 106 cells L-1) was recorded in Shetland but there was no reported toxicity in shellfish. In August 1999, toxicity in farmed mussels (Mytilus edulis) from Belfast Lough was associated with an extensive bloom (> 5 x 106 cells L-1) of P. minimum (FSA(NI) unpubl. data) and resulted in public health warnings being issued. Prorocentrum lima is a medium sized armoured dinoflagellate (32 – 50 µm long and 20 – 28 µm wide, Dodge, 1982) that is known to produce okadaic acid (Koike et al. 1998). Unlike the other dinoflagellates described in this study, P. lima is an epiphytic benthic dinoflagellate. In the UK, the threshold abundance (action level) has been set at 100 cells L-1 in England and Wales and Northern Ireland but no threshold has been set for this phytoplankter in Scottish coastal waters. P. lima has been observed in water samples from a number of coastal areas around the UK coast (Bresnan et al. 2007; Stubbs et al. 2007). The abundance of this species in the water column is generally low, although this may be due in part to a problem with current sampling strategies that may underestimate the abundance of epiphytic and benthic species in the plankton community (Stobo et al. 2008). P. lima has rarely been directly linked to shellfish toxicity in UK waters although research suggests that this phytoplankter is the probable cause of DSP episodes in Fleet Lagoon in south western England (Foden et al. 2005). Lingulodinium polyedrum, previously named Gonyaulax polyedra is a medium sized armoured dinoflagellate (42 – 54 µm diameter, Dodge 1982) that has been shown to produce yessotoxins (YTX) in culture (Paz et al. 2004). These workers also refer to the presence of yessotoxins being detected in natural samples of phytoplankton when L. polyedrum was present. The extent to which this species causes toxicity in shellfish and the level of abundance required to induce toxicity are unknown. Furthermore, although this species is monitored as part of the UK programme to monitor the presence of toxin producing species, Northern Ireland is the only UK region to set threshold abundance (100 cells L-1). Protoceratium reticulatum, previously named Gonyaulax grindleyi is a small armoured dinoflagellate (28 – 43 µm long by 25-35 µm wide, Dodge 1982). Production of yessotoxins by P. reticulatum grown in culture was demonstrated by Paz et al (2004) who also mention the occurrence of yessotoxins in green shell mussels (Perna canaliculus) during a bloom of P.

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reticulatum in New Zealand in 1996 (although Paz et al. 2004 do not give details of the bloom or cite references to it). As in the case of L. polyedrum, the extent to which P. reticulatum causes toxicity in shellfish and the level of abundance required to induce toxicity are unknown. This phytoplankter is monitored as part of the UK monitoring programme but Northern Ireland is the only region to set a threshold abundance (100 cells L-1). Dodge (1982) stated that P. reticulatum was found all round the British Isles (common in the North Sea) and cites Reinecke (1967) as the source for this phytoplankter forming a toxic red tide in South Africa. This phytoplankter is generally found in low abundance in UK waters although relatively high abundance (0.148 x 106 cells L-1) was observed in the Scottish Loch Creran in 1983 (Lewis 1985). An extensive survey of algal toxins in Scottish shellfish identified the presence of YTX in low concentrations, but there have not been any closures as a result of high concentrations of this toxin group (Bresnan et al. 2007; Stobo et al. 2008). Azaspiracids have been observed in Scottish shellfish (Stobo et al. 2008) and the causative organism (Azadinium spinosum) identified from Scottish waters (Tillmann et al. 2009). It is evident that heterotrophic dinoflagellates and ciliates can act as a vector for this toxin and more studies are required to fully elucidate the mode of shellfish intoxication for this toxin group. The occurrence of Phaeocystis spp. are recorded as part of a number of monitoring programmes in the UK (Gowen et al. 2008). Blooms of Phaeocystis spp. were recorded in the eastern Irish Sea in 1957 and 1958. During the first of these blooms cell colonies were counted (> 4,500 colonies L-1) and during the 1958 bloom cell numbers were counted with a maximum of 193 x 106 cells L-1 along the North Wales coast (Jones & Haq 1963). In 1992, a bloom (12,000 colonies L-1) in the same area was linked to the deaths of fish and crustaceans possibly as a result of oxygen depletion (ICES C.M.1993/ENV:7). Blooms of Phaeocystis spp. in the English Channel are described as being annual events (Davies et al. 1992) and can often be dense and persistent (Boalch 1987). In 2005, a Phaeocystis spp. bloom (8.0 x 106 cells L-1) in the Shetland Islands that extended down the east coast of Scotland was also described as having an effect on farmed fish (ICES 2006/OCC:04). Records of harmful microflagellate blooms in UK waters are rare in the literature. One of the best documented are the blooms of an unidentified species known as ‘Flagellate X’ (possibly Heterosigma akashiwo) in Loch Striven in 1979 (Tett 1980) and again in 1982 in Loch Fyne (Gowen et al. 1982). These events were associated with mortalities of farmed salmon.

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Some diatom species, especially those belonging to the genus Chaetoceros, have been known to cause mortalities of fish due to physical damage. A bloom consisting predominantly of C. wighami in Loch Torridon as well as a mixed bloom of C. debile and the silicoflagellate Dictyocha speculum19 were collectively responsible for the deaths of farmed fish with a market value of several million pounds (Bruno et al. 1989). A bloom of the dinoflagellate Heterocapsa triqueta (1.0 x 106 cells L-1) in the Shetland Islands in May 2001 caused substantial losses to fish farms as did a Gymnodinium spp. bloom (~ 9 x 106 cells L-1 ) in the Orkney and Shetland Islands in August of that year (ICES 2002/C:03). Other phytoplankton species present in the coastal waters of the UK may cause discolouration of the water but have little known direct ecosystem effects. For example, Myrionecta rubra20 has frequently bloomed in Southampton Water (Crawford et al. 1997) causing red discolouration of the surface water. The coccolithophorid, Emiliana huxleyi is known to bloom in the waters off Shetland (Head et al. 1998) as well as waters off the south west coast of England. Holligan et al. (1983) reported regular blooms (up to 8.5 x 106 cells L-1 in May 1982) at the shelf break (edge of the continental shelf) front on the Celtic and Armorican shelf regions and Garcia-Soto et al (1995) also reported an E. huxleyi bloom (> 2.0 x 106 cells L- 1) in the western English Channel in June 1992. More recently an extensive bloom which covered up to 16,000 km2 at its peak, was reported in July 1999 off the south western coast of England (Smyth et al. 2002). Blooms of the dinoflagellate Noctiluca scintillans have been recorded in the Irish Sea in coastal waters of the Isle of Man (T. Shammon pers comm.), Northern Ireland (FSA(NI) unpubl. data) and the English Channel (Boalch 1987). One of the most extensive was a bloom in 1982 that was recorded from Plymouth to the French coast and caused the water to look like ‘tomato soup’(Boalch 1987).

19 Previously named Distephanous speculum 20 Previously named Mesodinium rubrum - 39 -

2.3 Coastal Waters of the Republic of Ireland

2.3.1 Introduction

Records of red tides in Irish waters are sparse prior to 1976 (Parker 1981). However, at least twelve phytoplankton species have been recorded as causing blooms: Myrionecta rubra, Nitzschia sp., Phaeocystis pouchetti, Flagellate X, Glenodinium foliaceum, Ceratium tripos, Prorocentrum micans, Glenodinium sp., Noctiluca scintillans, Lingulodinium polyedrum, Karenia mikimotoi and Dinophysis acuminata (Jenkinson 1987). Routine monitoring of shellfish growing waters and fin-fish sites in coastal waters of Ireland commenced in the mid-1980s. A comprehensive picture of phytoplankton species and their blooms has become available from this data. A map of Ireland showing the locations mentioned in the text is shown in Figure 2.2.

2.3.2 Species of Alexandrium

Routine monitoring of HAB species includes species of Alexandrium. Additional samples are collected if the presence of Alexandrium spp. is detected in a sample. The first reported occurrence of PSP toxicity in the Republic of Ireland was in wild mussels from Cork Harbour (south east coast) in 1996. This followed a bloom (> 0.1 x 106 cells L-1) of A. tamarense in the area (Furey et al. 1998). Recent studies suggest that A. tamarense co-occurs with A. minutum and it was the latter which has been responsible for historical episodes of PSP toxicity in the harbour (Touzet et al. 2007). In their 2008 paper, Touzet et al (2008) suggest that Belfast Lough in Northern Ireland is the only location where PSP toxicity has been associated with A. tamarense. It has also been suggested that the A. minutum present in waters to the south and west of Ireland are non toxic (Touzet et al. 2007).

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Figure 2.2 A map of the Republic of Ireland showing locations mentioned in the text. Locations: 14, Wexford; 15, Youghal; 16, Cork Harbour; 17, Dunmanus Bay; 18, Bantry Bay; 19, Kerry; 20, Donegal.

2.3.3 Species of Dinophysis

Prior to the early 1980s, incidents of shellfish toxicity leading to DSP were not reported in any significant numbers to require intervention by the regulatory authorities in Ireland. However, through the 1980s during the period of greatest development of the shellfish aquaculture - 41 -

industry, reports in the scientific literature were published demonstrating the association between Dinophysis sp and DSP accumulation in mussels (Kat 1983). A biotoxin and phytoplankton monitoring programme was established in 1984. DSP toxins and the causative species including D. acuminata, D. acuta, D. fortii, D. norvegica and D. rotundata were detected regularly in Irish coastal waters. D. tripos and D. sacculus have also been identified but are considered rare (Marine Institute unpubl. data). Problems due to toxicity in shellfish often lead to extended closures of harvesting areas (Jackson & Silke 1995). These closures can vary from year to year and have ranged from years where the shellfisheries were open all year, to years when there were closures in operation for up to 90 % of the year. Particularly bad years occurred in 2001 (86 %), 2002 (59 %), 2005 (73 %), 2006 (90 %) and 2008 (68 %) with closures in the south west mussel growing areas (Figure 2.3). The predominant toxins measured have been

Figure 2.3 Periods of closures due to lipophilic toxins in shellfish growing areas in the southwest of Ireland. Red blocks indicate closure orders were in place in shellfish production areas.

SW Jan Feb Mar Apr May Jun July Aug Sep Oct Nov De

Year/Week 1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20 21 22 23 24 25 26 27 28 29 30 31 32 33 34 35 36 37 38 39 40 41 42 43 44 45 46 47 48 49 50 1984 1985 1986 1987 1988 1989 1990 1991 1992 1993 1994 1995 1996 1997 1998 1999 2000 2001 2002 2003 2004 2005 2006 2007 2008

okadaic acid and dinophysis toxin 2 (DTX2) in the summer months (these have been associated with the presence of Dinophysis acuta and D. acuminata) and azaspiracid toxins originating

- 42 -

from an unconfirmed phytoplankton source typically in late summer but extending into the winter months. D. acuminata is typically present from late May, peaks in late June and abundance declines during July – September. This species is typically associated with an increase in okadaic acid in shellfish. D. acuta generally peaks 2 months later in August and the abundance of this species falls to background levels by October. This species usually results in more shellfishery closures as it contains both okadaic acid and DTX2 (Figure 2.4).

Figure 2.4 Plots of Dinophysis acuta (left graph) and Dinophysis acuminata (right graph) cell counts for the period 1991 to 2008.

2.3.4 Species of Pseudo-nitzschia

Members of the Pseudo-nitzschia genus are a common component of the phytoplankton community in Irish waters with eight species identified to date (Cusack et al. 2004). One of those isolated, Pseudo-nitzschia australis, has been shown to produce domoic acid in culture (Cusack et al. 2002). ASP toxicity is regularly detected in the digestive organs of king scallop (Pecten maximus) and less commonly in the other soft tissues of this species. While this is thought to be due to Pseudo-nitzschia spp. the monitoring programme does not sample phytoplankton from offshore areas. The first major ASP event in non-scallop shellfish documented in Irish coastal waters was in 2005 when a Pseudo-nitzschia australis bloom (1.0 x 106 cells L-1) was associated with toxicity (444.9 µg/g whole flesh) in farmed mussels (Mytilus edulis) (ICES 2006/OCC:04). Several Pseudo-nitzschia species are commonly observed in the monitoring programme by light microscopy and are present year round but are typically observed between May to September with a peak in July. High cell counts (> 1 x 106) were

- 43 -

recorded in 1991, 2000, 2005, 2007, and 2009 in inshore areas and molecular methods have shown that in 2005 and 2009 March / April episodes of ASP in mussels were associated with monospecific blooms of P. australis.

Figure 2.5 Plots of the monthly abundance of Pseudo-nitzschia abundance in Irish coastal waters between 1990 and 2007.

2.3.5 Karenia mikimotoi

The first record of the ichthyotoxic dinoflagellate Karenia mikimotoi causing extensive blooms in Irish coastal waters was in 1976 (Ottway et al. 1979) when a large bloom (500 x 106 cells L-1) caused red discolouration of the water and extensive mortalities of lugworms along the south coast from Wexford to Youghal in mid July. Since then a number of K. mikimotoi blooms have been reported in the literature (Doyle et al. 1984; Raine et al. 1993; Raine et al. 2001; Silke et al. 2005). In August 1978, a bloom (7.7 x 106 cells L-1) was reported in Roaringwater Bay (Roden et al. 1980) and the following year mortalities of littoral and sub-littoral organisms in the inshore waters of Bantry and Dunmanus Bays on the south western coast were reported following a K. mikimotoi bloom (Cross & Southgate 1980). However, the most extensive blooms occurred in 2005. The first of these (3.0 x 106 cells L-1) appeared in coastal waters off the north west of Ireland in May and the second bloom (3.7 x 106 cells L-1) occurred in waters off the south west

- 44 -

in July. These blooms were associated with widespread mortalities of farmed shellfish along the coasts of County Mayo and County Galway. Fish and crustacean mortalities were reported from the counties of Kerry, Donegal, Galway and west Cork and in July, mortalities of polychaetes and cockles were reported from County Donegal (Silke et al. 2005).

2.3.6 Other HAB species

Records of flagellate blooms in Irish waters are rare although the occurrence of both ‘Flagellate X’ (maximum abundance 2.5 x 106 cells L-1) in 1983 (Doyle et al. 1984) and Olisthodiscus luteus in 1985 (ICES 1986/L:26) were responsible for major mortalities of farmed fish on the west coast. In the 1983 episode, ≈ 74 tonnes of farmed fish (salmon smolts, rainbow trout and steelhead trout) were lost. There are few documented reports of Phaeocystis spp. blooms in Irish waters. The first major Phaeocystis spp. bloom recorded along the west coast of Ireland occurred in 1990 and was unusual in that it extended from Co. Donegal to Co. Clare. Maximum colony abundance (80,000 colonies L-1) was recorded from a site in Inner Galway Bay and water was reported as being a ‘very dense orange brown’ (Pybus & McGrath 1992). A bloom (28.0 x 106 cells L-1) of Phaeocystis pouchetti was recorded off the south and south west coast between April and August 2002 (ICES 2003/C:06). During April 2003, blooms of Phaeocystis spp. with cell abundance of 16.5 x 106 and 1 x 106 cells L-1 occurred in Castlemaine harbour and Bantry Bay respectively (ICES 2004/ C:08). Noctiluca scintillans has also been recorded in Irish waters and is often associated with water discolouration. Parker et al. (1982) reported that in 1977, extensive blooms of N. scintillans were recorded from the south/south west coast and Jenkinson (1987) reported the occurrence of a bloom in Bantry Bay in 1978. Jackson et al. (cited in Pybus & McGrath 1992) also report blooms of N. scintillans off the east coast of Ireland in 1990. In July 2002, there was a short lived (one week) N. scintillans bloom (3.0 x 106 cells L-1) on the east coast of Ireland (ICES 2003/C:06).

- 45 -

Part 3

Anthropogenic Nutrient Enrichment and Harmful Algal Blooms: a Literature Review

3.1 Introduction

As was stated in Part 1, it has been suggested that there has been a global increase in the occurrence of HABs, and that this increase is due to nutrient enrichment of coastal waters. However, this view rests on a set of assumptions and in this part of the report we examine some of the arguments that have been made about the increase in HABs and about their causes. Before addressing the evidence relating nutrients and HABs in the coastal waters of the UK and Ireland, some relevant studies carried out in other parts of the world are examined. The diagrams in Figure 3.1 (showing the links from nutrient enrichment and other pressures to HABs) and Figure 3.2 (showing hypotheses for HAB generation) set out the conceptual framework we bring to this analysis. The remainder of this part of the report reviews the literature pertinent to the enrichment of coastal waters and the nutrient enrichment - HAB debate. The nutrient enrichment → HAB hypothesis is examined using four case studies: coastal waters of China, the Seto Inland Sea of Japan and the North Sea are discussed in detail and a summary of HABs in continental coastal waters of the United States of America is presented. The final part of this section discusses hypotheses about the occurrence of HABs.

- 46 -

Figure 3.1 The interactions between nutrient enrichment and drivers and pressures (orange boxes) and the occurrence of HABs.

- 47 -

Figure 3.2 A general hypothesis: HABs may be distinct from eutrophication. Apart from this there may be no single hypothesis applicable to all HABS: it may be necessary to pose specific hypotheses for each HAB life-form.

The conceptual framework in Figure 3.2 introduces the term ‘eutrophication’ which is discussed briefly here. We believe there is still much confusion over this issue. Scientific definitions of the term eutrophication tend to focus on the ecological aspects i.e. eutrophication is a process by which a water body evolves as it becomes enriched either naturally or through human activity (Nixon 1995 and see also Gowen et al. 2008 who discuss how the meaning of the term has evolved). In a European context, legal definitions such as those of the EU Urban Waste Water Treatment Directive and the OSPAR strategy to combat eutrophication emphasise the

- 48 - undesirable consequences of nutrient enrichment. The diagnosis of eutrophication thefore requires evidence of undesirable consequences to the balance of organisms and water quality resulting from increased plant growth (algae and higher plants) fuelled by anthropogenic nutrient enrichment (Tett et al. 2007). It is evident however, that the term is still used to mean nutrient enrichment rather than an ecological process and this frequently leads to statements such as ‘the increase in HABs was related to eutrophication’. Furthermore, evidence of nutrient enrichment and increased primary production (part of the eutrophication process) has been taken as evidence for the putative global increase in HABs (see for example Hodgkiss & Ho 1997). There is clear evidence that enrichment of European coastal regions over the last 30 - 40 years has increased phytoplankton biomass and production (Radach et al. 1990; Schaub & Gieskes 1991; De Jonge et al. 1996; Gowen et al. 2000). There is particular concern about land- locked basins such as the Baltic Sea (Elmgren 1989; Jansson & Dahlberg 1999; Karlson et al. 2002) and freshwater enriched coastal zones of Belgium (Lancelot et al. 1987), northern France (Cugier et al. 2005), the Netherlands (van Bennekom, 1975; Postma 1985; Cadée 1990; de Vries et al. 1998; Cadée & Hegeman 2002) and the Italian Adriatic (Degobbis et al. 1979; Vollenweider et al. 1992). It is important to consider whether the occurrence of HABs necessarily implies eutrophication, and if eutrophication is always accompanied by HABs. As will be discussed in later parts of this report, we are of the opinion that the occurrence of HABs is not, in general, an indicator of eutrophication and that HABs are not necessarily associated with eutrophication. However, we recognise that harmful algal blooms may be one of several undesirable outcomes of the human driven eutrophication process. In our view it is important to separate the issue of eutrophication from the question of whether anthropogenic nutrient enrichment stimulates the occurrence (where none have occurred before), causes an increase in the frequency of occurrence, or promotes an increase in the duration or spatial extent of HABs. The reason for this distinction is that ‘blooms’ (red tides, HABs, noxious, nuisance, exceptional blooms) as we use the term (and as used in much of the scientific literature) are discrete events that are distinct from a more general increase in biomass and production fuelled by anthropogenic nutrient enrichment. Nevertheless, it is evident from the scientific literature that, increased primary production resulting from anthropogenic nutrient enrichment has been taken as evidence for an increase in HABs. For example, Hodgkiss and Ho (1997) state that:

“This considerable evidence of significant changes in phytoplankton species occurrences, biomass and productivity, as well as shifts in predominance, occurring in regions as far apart as the North Sea and Hong Kong support the

- 49 - hypothesis that phytoplankton blooms are increasing in coastal waters on a global scale and that they are linked to long term increases in coastal nutrient levels.”

We are of the opinion that it is inappropriate to assume that an increase in primary production provides evidence for an increase in HABs. This is because such an increase in primary production might result from a general increase in micro-algal biomass or in growth rate due to increased nutrient flux; the increase might be coupled to a more productive spring phytoplankton bloom; all of these causes might be benign or might give rise to undesirable disturbances. Only certain types of HAB may contribute a significant increase in biomass and primary production.

3.2 Nutrient Enrichment of Coastal Waters

It is clear that there has been a massive global increase in anthropogenic nutrient loading to the sea, although in some regions nutrient loads have begun to decrease as a result of economic recession and legislation. Human activity, particularly during the early part of the 20th century has, through increased population, industrialisation and intensification of agriculture increased the bio-availability of nitrogen with nitrogen (N) fertilisers considered the main source. Vitousek et al. (1997) report that  80 Tg21 of N is produced each year for fertilizer (citing a 1993 FAO report as the source of data) and more than 20 Tg y-1 are emitted as a result of burning fossil fuels. Galloway and Cowling (2002) give estimates of 85 and 21 Tg N y-1 from fertilisers and burning fossil fuels respectively and ≈ 30 Tg y-1 from cultivation. In total, human activity is responsible for the fixation of approximately 140 Tg of new N each year which is additional to natural terrestrial N-fixation (≈ 89 Tg y-1). Industrialisation and intensification of agriculture have also influenced the flux of phosphorus from land to seas and oceans and according to Howarth et al. (2002) and references cited therein, human activity has increased the flux of P by  14 Tg P y-1 to a current value of  22 Tg P y-1. As noted above, coastal waters in many regions of the world have become enriched as a result of human activity. Run-off from agricultural land, domestic and industrial waste, groundwater seepage into coastal waters and atmospheric deposition have, to varying degrees contributed to this enrichment (Jickells 2005). Time-series from a number of rivers show significant increases in concentration (Figure 3.3). For the Changjiang (Yangtze) River, the data presented by Li et al. (2007) shows that the loading of N (as ammonium, nitrate and nitrite) increased from ≈ 0.2 x 106 t y-1 in the early 1970s to  1.6 x 106 t y-1 by the late 1990s.

21 1 Tg is equal to 1012 g or 1 million (106) metric tones. - 50 - Figure 3.3 Time series of riverine concentrations. A, the mean annual concentration of nitrate (µM) in the Thames (from Heathwaite et al. 1996); B, annual mean nitrate (NO3) silicate (Si) and phosphate (PO4) concentrations (µM) in the Changjiang (Yangtze) River (from Li et al. (2007).

700 A 600

500 M)

 400 300 Nitrate ( Nitrate 200 100 0 1928 1934 1939 1945 1951 1957 1962 1968 1974 1979 1985

140 1.8 B 120 1.6 M) 1.4  100 M) 1.2 80 1.0 0.8 60 0.6 40 Si ( Phosphate 0.4 NO3 Nitrate and and silicate ( Nitrate 20 0.2 DIP 0 0.0 1959 1963 1967 1971 1975 1979 1982 1986 1990 1994 1998

According to Faeth and Greenhalgh (2002) and references cited therein, non-point sources (in particular from croplands) were the largest source of total nitrogen (82 %) and total phosphorus (84 %) to waterways in the Unites States. Atmospheric deposition as a transport mechanism for nutrients to coastal waters and shelf seas is important in part because of the amounts (Rendell et al. 1993; Asman et al. 1995) but also because deposition of material can occur in more open coastal waters some distance from shore (Jickells 1995). Particular sources of nutrients include agricultural emissions from livestock and motor vehicle exhaust fumes. Paerl et al. (2002) stated that atmospheric deposition of nitrogen accounted for between 10 and 40 % of new nitrogen loading to estuaries of the eastern United States and eastern Gulf of Mexico. According to Howarth (2008) the atmospheric deposition of oxidised nitrogen compounds in the north east United States was the single largest input of nitrogen to the region south of the watershed of Virginia which flows into Chesapeake Bay.

- 51 - For the greater North Sea and the Irish Sea the atmospheric deposition of N has been estimated as 0.35 x 106 (OSPAR 2000) and 0.043 x 106 t y-1 (Gillooly et al. 1992) respectively. Groundwater can be an important source of nutrients particularly in regions where there is no riverine transport (Jickells 2005). Giblin and Gaines (1990) considered groundwater to be a significant source of N for small coastal embayments and estimated the groundwater N input to Town Cove (a small embayment on the coast of Cape Cod, U.S.) as 300 mmol m-3 y-1. According to Giblin and Gaines (1990) when this figure was adjusted for the volume of the receiving water it represented a N source larger than the input of sewage - derived N in larger river estuaries. Not all of the riverine N and P ends up in coastal waters. Nitrogen can be denitrified22 in hypoxic sediments and P bound to particles. Billen et al. (1985) considered the possibility that ‘cleaning up’ estuaries would lead to greater N discharge and Howarth (2008) was of the opinion that only a small proportion ( 15-45 %) of the [net anthropogenic] nutrient input reached the coast with the remainder either retained in the landscape or denitrified as N2 or N2O gas. Hydes et al. (1999) suggested that wide shelf areas like the north west European shelf could be considered as extended estuaries within which the final stages of mixing between less saline coastal and more saline oceanic waters takes place. As such, considerable cycling and reprocessing of nutrients might be expected to take place in these regions. Using the LOICZ23 modelling approach, Smith et al. (1997) estimated that in the Northern North Sea, the loss of N by denitrification exceeded the land and atmospheric inputs. The equivalent rate of denitrification was 0.1 mol N m-2 y-1 with a rate of 0.2 mol N m-2 y-1 for the Southern North Sea (c.f. the Irish Sea for which Simpson and Rippeth (1998) estimated the rate -2 -1 as 0.3 mol N m y using the same technique). These estimates are in line with measurements of sediment denitrification rates (Lohse et al. 1996; Trimmer et al. 1999). Using nutrient salinity relationships based on the January 1989 NERC-NSP24 data, Hydes et al. (1999) showed that for the North Sea, predicted and measured concentrations of P were similar. For nitrate the relationship was suggestive of a significant loss (probably by denitrification, see also Gowen et al. 2002 for similar observations in the Irish Sea) and according to Hydes et al. (1999) for the NERC North Sea project area, by the end of winter the nitrate deficit is 580 ktonnes of nitrogen. The nitrate deficit in the southern North Sea was equivalent to a denitrification rate of 0.25 mol N m-2 y-1 assuming a flushing time of one year for the North Sea. Seitzinger and Giblin (1996) estimated that to balance the loss of nitrate by

22 Denitrification is the name given to the process carried out by various bacteria, during which nitrate ions act as an alternative electron acceptor to oxygen, resulting in the release of N2 gas. 23 Land Ocean Interactions in the Coastal Zone. 24 UK Natural Environment Research Council, North Sea Project. - 52 - denitrification, a net flux from the North Atlantic onto the North West European shelf of 16 x 1010 mol N y-1 (2.44 x 106 t N y-1) is required in excess of the inputs from rivers and the atmosphere. With respect to silicate and phosphorus in northern European shelf seas, both are conserved within the Irish Sea (Simpson & Rippeth 1998; Gowen et al. 2002). Dissolved inorganic nitrate is not the only form of available nitrogen released into coastal waters. Nitrite is generally a minor component but much of the nitrogen from domestic sources + is likely to be in the form of ammonium (NH4 ). Dissolved organic nitrogen (DON) may also be a significant fraction of the nitrogen; literature on this form of nitrogen has been extensively reviewed by Antia et al. (1991) and Bronk (2002). ‘Natural’ abiotic DON inputs include atmospheric (Cornell et al. 1995), rainwater (Seitzinger & Sanders 1999), and riverine (Meybeck 1993). In addition to these ‘‘natural’’ sources of N, increased anthropogenic nutrient fluxes to coastal waters, both from land-based agriculture and marine-based industries such as (Gowen & Bradbury 1987), have the capacity to introduce both inorganic and organic nutrients to the marine environment. Agricultural runoff may contain dissolved organic forms of nitrogen depending on the composition of the fertiliser used (see for example, Glibert et al. 2006) and organic nitrogen in the form of particulate detrital material. Phytoplankton utilization of inorganic N is well known, and DON utilization by phytoplankton has also been demonstrated by several authors including for example, the toxin producing Alexandrium tamarense (Stolte et al. 2002). The cycle of particulate material in coastal waters is important because: nutrient transport may become decoupled from water transport if a significant fraction resides in the particulate phase; particles might represent a reservoir of nutrients that erosion or resuspension could introduce to the water column where desorption or remineralisation might make the nutrients available. Consideration must also be given to the abiotic sorption of nutrient species to inorganic sedimentary particles which frequently occur in great numbers in energetic coastal waters (Jickells, 1998; Prastka et al. 1998; Jickells et al. 2000). Much of the riverine phosphorus is in particulate form or as dissolved inorganic phosphate 3- (PO4 ) bound to particles and the equilibrium between the dissolved and particulate phase is heavily biased towards the latter at moderate particulate loads. High removal of P occurs in the low salinity reaches of many estuaries (Prastka et al. 1998). The fate of this particulate material is important, since burial will remove P from the system and particle flushing may result in desorption in more saline waters where the particulate load is lower. One consequence of industrialisation is that not only are the particulate loads higher but sedimentation is lower, thus flushing is the likely fate. The desorption of large amounts of P in offshore regions means that the local DIN:DIP ratio may be a poor indicator of the potential limiting factors.

- 53 - The silicon cycle is not directly influenced by human activity since inputs to coastal waters are largely determined by catchment geology and weathering. However, water management (e.g. for domestic use, power generation and irrigation) can influence the loading of silicon to coastal waters. Slowing river flow allows increased riverine primary production and nutrient uptake and subsequent diatom sedimentation results in loss of silicate from the water column (Admiraal et al. 1990). In some regions of restricted exchange such as the southern North Sea, these factors have resulted in an increase in the coastal inorganic N:Si ratio (Aure et al. 1998). Humborg et al. (1997) present data on changes in silicate loading of the River Danube and concentrations in coastal waters of the Black Sea before and after the damming of the Danube. The silicate loading before completion of the dam was ≈ 800 x 103 t y-1 but only 230 – 320 x 103 t y-1 after the dam was build. Corresponding changes in silicate concentrations in coastal waters were 55 µM and 20 µM before and after construction respectively. A decrease in silicate concentration and flux has also been reported for the Changjiang (Yangtze) River (Figure 3.3). In some regions of the world, riverine inputs of nutrients (see above) and atmospheric inputs are expected to increase (Jickells 2005). Howarth (2008) has suggested that under some scenarios, nutrient input to Chesapeake Bay (eastern seaboard of the U.S.) is predicted to increase. However, reductions in the use of phosphates in detergents, tertiary treatment of sewage and the designation of areas of land as ‘nitrate vulnerable zones’ (within which the application of fertilizer is strictly controlled) have resulted in stable and in some catchments reductions in riverine inputs of nutrients to coastal waters of western Europe. In Dutch coastal waters, a decrease in total phosphorus from Lake IJssel and dissolved inorganic phosphate (DIP as PO4) concentrations in the western Wadden Sea has been reported but the total N load and dissolved available inorganic nitrogen (DAIN as NO3, NO2 and NH4) appear to be stable (Philippart et al. 2007, see also Cadée & Hegman 2002). Analysis of a  30 year time series of nutrient data from the Irish Sea shows that in recent years the winter concentration of DIP has decreased and DAIN is stable (Gowen et al. 2008).

3.3 Nutrient Enrichment and Blooms of Harmful Micro-algae

3.3.1 Introduction

Laboratory studies with cultured algae have shown how increases in the supply, or ambient concentration of a limiting nutrient, can lead to an increase in the population specific growth rate and that the final biomass under such conditions is proportional to the amount of limiting nutrient supplied (Droop 1968; Davidson et al. 1993). Under natural conditions, it is therefore likely that increases in nutrient availability will lead to enhanced growth and biomass, so long as

- 54 - nutrients are limiting and algal losses (due to dilution, sinking or grazing) do not increase in proportion to increased biomass. This is the argument for eutrophication and the link between nutrient enrichment and blooms requires that some algae would respond more than others to nutrient enrichment because in ecological theory each species has a different set of properties. If these algae are intrinsically or potentially harmful, then an increase in HABs might be expected. It might also be the case, however, that the species that respond are not harmful (although as we have suggested any phytoplankter which reaches a sufficiently high biomass and impacts ecosystem goods and services could be considered harmful). Furthermore, given the widespread enrichment of many coastal regions of the world (and in some regions continuing enrichment) and the putative global increase in HABs it is not difficult to see why anthropogenic nutrient enrichment of coastal waters is thought by some to be one of the main drivers for the apparent global increase in HABs.

3.3.2 The nutrient enrichment HAB hypothesis

3.3.2.1 Introduction

Whether or not anthropogenic nutrient enrichment has caused or influenced the occurrence, frequency of occurrence and spatial and temporal extent of HABs and HAB species is a complex issue and the nutrient enrichment → HAB hypothesis has been widely debated in the scientific literature. Many publications make reference to the link between anthropogenic nutrient enrichment and HABs but do not present any data or detailed assessment of the issue. A number of publications have considered the issue in detail and there have been several reviews (see for example, Anderson 1989; Hallegraeff, 1993; Richardson 1997; Smayda 1989, 1990, 2008; Anderson et al. 2002, 2008; Sellner et al. 2003). Recently, time-series have been assembled and used to examine the relationship between HABs and enrichment. Relevant sections of an ICES Workshop in 2006 on Time Series Data relevant to Eutrophication Ecological Quality Objectives (ICES, 2007) are discussed below. A number of studies presented at the workshop have been collated in a special issue of the Journal of Sea Research (Volume 61, Issues 1 – 2, 2009). Borkman et al. (2009) provide a summary of each of the papers and the findings of those studies more relevant to the nutrient enrichment → HAB hypothesis have been included as part of this review. In reviewing the scientific literature, it is clear that there is no scientific consensus on the relationship between the occurrence of HABs and anthropogenic nutrient enrichment. On the one hand, Anderson (1989) was of the opinion that:

- 55 - “it is now firmly established that there is a direct correlation between the number of red tides and the extent of coastal pollution….”.

In support of this statement Anderson (1989) cites the studies of Lam and Ho (1989) in Tolo Harbour (Hong Kong) and by Okaichi (1989) in the Seto Inland Sea. Konovalova (1989) shared the same view as Anderson (1989) and stated that for far eastern coastal waters of the former Soviet Union:

“Undoubtedly, the frequency and concentration of “red tides” are directly connected with increased eutrophication of coastal waters under the influence of anthropogenic factors.”

Park et al. (1989) was also of the opinion that there had been an increase in HABs stating that:

“Outbreaks of red tides in Korean neritic waters have remarkably increased in the last decade and have caused severe damage to cultured shellfish and other living organisms.”

On the other hand, at the same conference Smayda (1989) expressed the view that:

“This implicit concept of an anthropogenic trigger seems to be the favoured notion”

However he was not:

“ready to embrace this view, despite the widespread, provocative evidence.”

At a workshop during the 4th international conference on toxic marine phytoplankton in 1990 (see Smayda & White 1990) it was concluded that:

“Although it is generally suspected that toxic and noxious algal blooms have been increasing in frequency and intensity worldwide over the past 20 years or so, it was agreed that it is not possible to conclude this with certainty on a global level because the long-term data on the abundance of algae in the sea are insufficient in scales of both time and space.” and that:

“The causes and mechanisms of blooms may differ as there seem to be two major types of blooms, those in which nutrient additions to coastal systems are obviously implicated (for example, in Tolo Harbour (Hong Kong), in the Seto Inland Sea (Japan), and in the Aegean Sea in the vicinity of sewage outfalls) and those blooms that are not obviously associated with coastal enrichment (for example, Alexandrium, Pyrodinium, Dinophysis, etc.).”

Interestingly, the workshop considered that it would take a long time to answer the question and that an appropriate timescale for monitoring algal blooms was between 5 and 10 years. A

- 56 - further round table discussion of trends in the occurrence of harmful algal events (see Smayda & Wyatt, 1995) concluded that:

“Phytoplankton biomass and bloom frequency have increased in some regions in response to eutrophication, but this development has not favoured harmful species exclusively.”

“There has been an increased awareness of the problems caused by algae, by scientists, managers, and health authorities, which has led to the detection of new toxins and new toxic species, and of harmful algal events in areas where they were not previously reported.”

“Aquacultural operations seem to be closely linked with harmful algae events, but it is still not clear whether they simply detect species to which attention was not previously directed, or whether the environmental changes associated with such activities lead to increases in the biomass of problem algae.”

In her review, Richardson (1997) stated that:

“Indeed, in some areas-especially those with limited water exchange such as fjords, estuaries and inland seas-there does seem to be good evidence for a stimulation of the number of algal blooms occurring by eutrophication. However, the relationship between the occurrence of harmful phytoplankton blooms and environmental conditions is complicated and anthropogenic perturbation of the environment is certainly not a prerequisite for all harmful algal blooms.”

Hodgkiss and Ho (1997) were of the opinion that:

“This considerable evidence of significant changes in phytoplankton species occurrences, biomass and productivity, as well as shifts in predominance, occurring in regions as far apart as the North Sea and Hong Kong support the hypothesis that phytoplankton blooms are increasing in coastal waters on a global scale and that they are linked to long term increases in coastal nutrient levels.”

However, Anderson et al. (2002) concluded that:

“It is important to avoid ascribing the apparent global increase in HABs solely to pollution or eutrophication, although the public and the press often assume this linkage.”

Similarly, Sellner et al. (2003) were of the opinion that in relation to the occurrence of HABs:

“Blooms of these organisms are attributed to two primary factors: natural processes such as circulation, upwelling relaxation, and river flow; and, anthropogenic loadings leading to eutrophication. Unfortunately, the latter is commonly assumed to be the primary cause of all blooms, which is not the case in many instances.”

- 57 - Glibert et al. (2005) were of the opinion that anthropogenic nutrient enrichment was a key factor:

“Eutrophication is now recognised to be one of the important factors contributing to habitat change and to the geographical and temporal expansion of some harmful algal bloom (HAB) species (Smayda, 1990; Anderson et al., 2002).” and that the situation in coastal waters of China, northern Europe and the US provided clear examples and where according to Glibert et al. (2005):

“Since the 1970s, when escalation in use of chemical fertilizer began in China, the number of HAB outbreaks has increased over 20-fold, with blooms that now are of greater geographic extent, more toxic, and more prolonged (Anderson et al., 2002). Other examples…..In northern European waters, blooms of the mucus- forming HAB species Phaeocystis globosa have been shown to be directly related to the nitrate content of riverine and coastal waters (Lancelot, 1995). In the United States, a relationship between increased nutrient loading from the Mississippi River to the Louisiana shelf and increased abundance of the toxic diatom Pseudo-nitzschia pseudodelicatissima has been documented.”

Based on the findings of a meeting sponsored by the U.S. Environmental Protection Agency in 2003, Heisler et al. (2008) in their paper ‘Eutrophication and Harmful Algal Blooms: a scientific consensus’ concluded that:

“Degraded water quality from increased nutrient pollution promotes the development and persistence of many HABs and is one of the reasons for their expansion in the U.S. and the world;”

Perhaps the main reason for these differing views is the complexity of the relationship between nutrient enrichment and the factors controlling phytoplankton growth, the accumulation of biomass and formation of HABs. As the conceptual map in Figure 3.1 illustrates, there are multiple pressures (which may act in synchrony or antagonistically) and possible outcomes which are influenced by the ecophysiology of individual phytoplankters and the ecohydrodynamic conditions within which they live. Before reviewing particular case studies therefore, we consider some of the issues which impinge on the nutrient enrichment → HAB hypothesis.

3.3.2.2 Historical and natural occurrence of HABs

It is evident from the early scientific literature that many of the phytoplankters that are now referred to as HAB species have a wide geographical distribution which predates enrichment of coastal waters. It is also widely accepted that HABs are not a new phenomenon. According to

- 58 - Fukuyo et al. (2002) ‘The History of Great Japan’ that was edited more than 300 years ago, records the occurrence of red tides and that one in 1234 AD, caused a mass mortality of fish and human deaths from eating fish. Fukuyo et al. (2002) also note that in Northern Japan, local folklore advises not to eat shellfish during snow water runoff (which occurs in early spring) into the sea. In the UK there is similar folklore: only eat shellfish when there is an ‘R’ in the month, i.e. avoid the summer months (May to August). Perhaps as Fukuyo et al. (2002) suggest such folklore has arisen because:

“this indicates that toxin contamination of shellfish has repeatedly occurred almost every year over a long time, leading to many tragedies among the local people.”

Richardson (1997) suggests that perhaps one of the first documented HABs (although in freshwater) is from the Bible (Exodus 7: (20-21) and Anderson et al. (2002) make reference to the ships logs from voyages by Captains Cook and Vancouver which record discoloured water and poisonous shellfish. There are a number of well documented examples which show that the occurrence of HABs predate anthropogenic enrichment of coastal waters. For example, Brongersma-Sanders 1957 compiled records of red water and mass mortalities up to the mid 1950s (Figure 3.4). Some caution is needed here however, because it is known that in some coastal regions, enrichment of coastal waters has been taking place for longer than the last 30 to 40 years (Cugier et al. 2005; Kemp et al. 2005). In British Columbia, Canada (Gaines & Taylor 1985) and Norway (Yndestad & Underdal 1985) the first recorded outbreaks of Paralytic Shellfish Poisoning (PSP) were in 1793 and 1901 respectively. In a number of instances these early records together with investigations into toxic shellfish episodes from the 1930s and 1940s (e.g. Medcof 1985) pre date coastal enrichment and it can be concluded that toxin producing algae are naturally occurring and their harmful effects are not a recent phenomena.

- 59 -

Figure 3.4 The global distribution of discoloured water and mass mortalities. (Redrawn from Brongersma-Sanders 1957). Blue circle, red water, no mass mortality; green triangle, red water coinciding with mass mortality; pink star, mass mortality, probably coinciding with red water.

According to Hallegraeff (1993) there were ≈ 1300 cases of DSP in Japan between 1976 and 1982, > 5000 cases in Spain during 1981 and ≈ 3300 cases in France during 1983. Why DSP toxicity in humans was first recorded at these times is unclear. However, given that there have been no recorded fatalities associated with DSP poisoning and the symptoms are similar to bacterial induced gastroenteritis, it is possible that earlier cases were not identified as DSP. The sudden appearance of Dinophysis spp. in coastal waters of northern Europe seems unlikely. Species of Dinophysis were present in these waters well before the 1980s. Cleve (1900) reported D. acuta from waters in the northern North Sea, off Scotland and in the Irish Sea (and one cell in a sample from Puget Sound). Gran (1927, 1929) reported the presence of D. acuminata, D. acuta and D. norvegica in coastal waters of Norway during 1922 and 1926-1927 (only D. acuminata in 1927). Herdman and Riddell (1911, 1912) recorded the presence of Dinophysis sp. in the Scottish west coast sea lochs Hourn in July 1908 and 1909 and Torridon in July 1911 and in the Firth of Lorne in 1909. Lebour (1917) recorded the presence of D. acuminata in the English Channel (off Plymouth) during a study in 1915 and 1916. The presence of Dinophysis spp. in the North Sea was reported by Lucas (1942); Dodge and Hart-Jones (1974) recorded the presence of D. acuminata, D. acuta, D. norvegica and D. rotundata at a coastal station off the north east coast of England during a 14 month study during 1971 and 1972 and Dodge (1977) reports the

- 60 - wider distribution of these species in the North Sea. Interestingly, Gran and Braarud (1935) reported the presence of nine species of Dinophysis in the Bay of Fundy in 1932 when a maximum abundance of 1,560 cells L-1 of D. norvegica and 1,100 cells L-1 of D. acuminata were recorded. Species of Dinophysis were present in the Scottish sea loch Creran at least as early as the mid 1970s (P. Tett unpubl. data). With respect to Alexandrium tamarense, the original description (of Gonyaulax tamarensis) was based on cells collected from the Tamar estuary (Lebour 1925). In his book on marine dinoflagellates of the British Isles, Dodge (1982) described A. tamarense as a neritic species that was found in coastal waters of the UK but particularly in the west and north. Species of Pseudo-nitzschia came to prominence in 1987, when human illness and fatality were associated with consumption of blue mussels. Species of Pseudo-nitzschia are widespread however, not all species have been shown to produce toxins. Cleve (1900) noted that: Nitzschia delicatissima was widespread in the N Atlantic during spring 1898; N. pungens was noted from Japan and the Gulf of Bengal; N. seriata from the Azores to Shetland. In a series of studies of the plankton in Scottish west coast waters and the Irish Sea, Herdman and Riddell (1911, 1912) recorded the presence of Nitzschia seriata in Loch Hourn in July 1909; N. seriata in the Firth of Lorne in 1909 and 1910; N. seriata and N delicatissima (the latter up to 335 x 106 cells L-1 based on a net sample) in Loch Torridon in July 1911. During surveys of Norwegian coastal waters in 1922 and 1926 - 1927, Gran (1927, 1929) recorded the presence of Nitzschia delicatissima and N. seriata. Gran and Braarud (1935) also report Nitzschia delicatissima and N. seriata from the Bay of Fundy. The names of the Nitzschia species have recently changed and since electron microscopy or molecular methods (rather than light microscopy) are considered necessary for reliable identification, it may not be possible to determine whether these early records were of toxin producing species of Pseudo-nitzschia. Nevertheless, species of this genus are found throughout coastal waters of the world and domoic acid in shellfish tissue has been widely reported. The question why amnesic shellfish poisoning should first appear in 1987, in eastern Canada and in winter is not clear. Given that the effects of algal biotoxins on human health were well known in eastern Canada, at least since the early part of the 20th century (Medcoff 1985) and the effects of domoic acid are quite distinct from those associated with PSP and DSP, it seems unlikely that earlier ASP toxicity events would have been missed. The presence of dinoflagellate cysts25 in sediments can provide a means of investigating the spatial and temporal occurrence of particular dinoflagellate species over hundreds and in

25 As part of their reproductive cycle, some species of dinoflagellate (not all species are known to produce cysts) produce resting cysts which settle to the sea bed to be re-suspended at a later date and develop into - 61 - some cases thousands of years before human records. Based on the distribution of fossil cysts of Gymnodinium catenatum in sediment cores from Scandinavian waters, Dale et al. (1993) and Dale and Nordberg (1993) suggested that climate change (in particular the medieval warm epoch provides one explanation for why this species was historically much more abundant in Scandinavian waters. Mudie et al. (2002) undertook a detailed examination of sediment cores from the Pacific and Atlantic coasts of Canada and concluded that: the 10,500 year record from the Pacific clearly showed that the largest blooms corresponding to cysts of Protoceratium reticulatum and Gonyaulax spinifera occurred in the early Holocene period26; there were cycles of individual bloom species (with one species replacing another after a period of dominance) unlike the last 60 years during which one or more species co-occur; the sedimentary record from the Atlantic coast of Canada also shows that Alexandrium spp., P. reticulatum, G. spinifera and Lingulodinium polyedrum cyst abundance was an order of magnitude greater in the early Holocene sediment compared to recent sediments. Mudie et al. (2002) further concluded that:

“The similarity of pre-industrial age cyst records of ‘red tide’ histories in the oceanographically different Pacific and Atlantic regions of Canada indicates that climate change (including surface and storminess) is the main driving stimulating blooms.”

Changes in the abundance of cysts in the sediments have also been related to industrialisation and anthropogenic nutrient enrichment. For example, Kim and Matsuoka (1998) related changes in dinoflagellate cyst abundance and in particular an increase in the proportion of heterotrophic species to eutrophication in Omura Bay, Kyushu (Japan). Wang et al (2004) quantified cyst abundance in sediments from two basins in Daya Bay (Southern China) and concluded that an increase in the abundance and diversity of cysts, in particular an increase in cysts of heterotrophic dinoflagellates, was consistent with a change in water quality; the result of nutrient enrichment which began in the 1980s and increased significantly in the 1990s. However, in relation to this particular coastal region, Yu et al. (2007) related an increase in chlorophyll and HAB frequency in Daya Bay to increased water temperature in the Bay resulting from the discharge from a power station, which began operation in 1994.

3.3.2.3 Increased environmental awareness and monitoring of coastal waters

a vegetative cell thereby maintaining the planktonic population from one year to the next. A proportion of the cysts become buried in the sediment and in coastal regions where sediments accumulate and are relatively undisturbed (by tidal re-suspension and bioturbation (mixing) by burrowing benthic animals) these cysts become part of the sedimentary record.

26 The Holocene period began between 10,000 and 11,700 years ago and continues to the present. - 62 -

Over the last 20 to 30 years there has been an increase in the awareness of environmental issues by Government, NGOs, scientists and the general public. Food hygiene and safety have also been given broader consideration. As a consequence, there has been a substantial increase in environmental monitoring and monitoring of food hygiene including sea foods. Both Anderson (1989) and Hallegraeff (1993) make reference to the increased reporting, scientific investigation of harmful algal blooms and Hallegraeff (1993) used maps of the global distribution of DSP toxicity and Pseudo-nitzschia species as illustrations of increased reporting. Regulation (EC) No 854/2004 (OJEU, 2004) and formally the Shellfish Hygiene Directive requires member states of the European Union to monitor for the presence of toxin producing phytoplankters in the vicinity of natural and cultivated shellfish beds. This has resulted in an unprecedented level of monitoring HAB species in European waters. For example, prior to the introduction of the Directive in the mid 1990s, there was limited monitoring of toxin producing algae or levels of toxin in shellfish tissue in UK coastal waters and this was restricted to paralytic shellfish toxins (McCaughey & Campbell 1992; Joint et al. 1997). Toxin producing algae only came to public attention on the rare occasions when they led to outbreaks of shellfish poisoning or the deaths of wild or farmed animals i.e., when they were, evidently, HABs. Thus, until the 1990s, reports of human PSP poisoning from UK coastal waters were rare (Ayres 1975) and mainly restricted to the northeast coast of England (Ayres et al. 1982 and see Part 2). The presence of DSP and ASP toxins in shellfish was unrecorded and there were no records of human illness linked to these toxins. Since the introduction of the monitoring programme, PSP and DSP toxins above action levels have been recorded in shellfish throughout UK waters, with ASP posing a serious problem in Pecten maximus (King scallop) in Scottish waters in the late 1990s. However, as noted above, species of Dinophysis and Pseudo-nitzschia are not new to UK waters. The question therefore arises as to whether the current distribution of these toxin producing species in UK waters reflects geographical spreading, perhaps as a result of introduction/ transfer of cells between coastal areas, the increase in monitoring and reporting or an increase in response to anthropogenic nutrient enrichment (including aquaculture development)? Observations of Dinophysis and ‘Nitzschia seriata’ and ‘N. delicatissima’ from a century ago suggest that these genera have been part of the phytoplankton community for a long time. The simplest answer is that low numbers of these species particularly Alexandrium and Dinophysis spp.27 are a natural component of the summer phytoplankton in UK coastal waters and that

5In the UK the regulatory action level for Alexandrium spp. and Dinophysis spp. abundance is presence and 100 cells L-1 respectively. - 63 - increased monitoring is why in Northern Ireland the first recorded occurrence of biotoxins in the shellfish tissue was in 1994 for DSP and 1999 for ASP. Similarly in Scotland, DSP toxins were first recorded in 1992 and ASP toxins in 1998. An alternative explanation is that the occurrence of HAB species and toxic events has increased in response to anthropogenic nutrient enrichment. This hypothesis is tested in Part 4. Wang et al. (2008) report an increase in the level of HAB monitoring and sampling in the South China Sea between 1980 and 2003. Figure 3.5 (redrawn from Wang et al. 2008) shows that during this period of time the number of sampling stations increased from about 3 to 100.

Figure 3.5. The relationship between increased sampling effort and the occurrence of HABs in the South China Sea between 1980 and 2003. (Redrawn from Figures 1 and 10 of Wang et al. 2008). A, increase in the number of sampling stations and samples collected per year; B, the relationship between the number of HABs and monitoring frequency; C, the ratio of HAB occurrence to monitoring effort.

120 1200 70 A B 60 100 1000 50 80 800 40

60 600 30

HAB occurrence 40 400 20 Samples per year Number of stations 10 20 200 0 0 0 0 200 400 600 800 1000 1200 1980 1983 1986 1989 1992 1994 1997 2000 2003 Monitoring frequency

0.40 C

0.30

0.20

0.10

Ratio of HABs to monitoring 0.00 1980 1983 1986 1989 1992 1994 1997 2000 2003

Wang et al. (2008) compared the occurrence of HABs and the increased level of monitoring and concluded that:

“There is no consistent relation between the HAB occurrence and the monitoring frequency (R2= 0.0654, P > 0.05 Fig. 10A, B)”

and that:

- 64 -

“The increase of monitoring frequency may contribute only a small part (Fig. 10A) to the increase in HAB reports.”

One interpretation of the data plotted in Figure 3.5C, is that as monitoring effort has increased the number of HABs has decreased. One reason for this might be that monitoring was initially focussed on areas where HABs occurred frequently but as the monitoring effort increased it encompassed areas where HABs were less frequent. It is also evident from the data (Figure 3.5B) that there was a marked decrease in the occurrence of HABs between 1998 and 2003 despite a high level of monitoring. Considering the data prior to 1998, there is a significant positive correlation (R2 = 0.98) between the number of HABs reported and the level of monitoring effort suggesting that more blooms were reported as a consequence of increased monitoring effort.

3.3.2.4 The influence of climate change

In addition to the studies of fossil dinoflagellate cysts mentioned above, a number of publications have considered the relationship between future climate change and the occurrence of HABs and HAB species (see for example, Moore et al. 2008). There is a need to distinguish long term (century scale) trends driven by global warming from inter-annual fluctuations and decadal scale variation such as that caused by the North Atlantic Oscillation and El Niño Southern Oscillation. There have been some attempts to link climate change and the occurrence of HABs on a global scale. Hayes et al. (2001) proposed that the reported abrupt increase in marine ‘outbreaks’ (disease epidemics, mass mortalities, population explosions and HABs) since the mid 1970 coincided with a shift in the global climate regime. A consequence of this regime shift was a change in the biogeochemistry of iron (in particular, an increase in supply via atmospheric dust) that has brought about changes in the micronutrient factors that limit the growth of opportunistic organisms and pathogenic micro-organisms. By reducing iron limitation of cyanobacteria growth and nitrogen fixation, Walsh and Steidinger (2001) suggested that the deposition of Saharan mineral aerosols could indirectly support red tides of Karenia brevis28 in the eastern Gulf of Mexico. The idea of a global scale mechanism is consistent with the view expressed by Smayda (2008) that the apparent synchronicity in the putative world wide increase in HABs was suggestive of a general change in the plankton habitat. However, Hayes et al. (2001) make it clear that their arguments were: ‘honestly speculative’. There is more evidence of a linkage between HABs and climate change at regional and local spatial scales. With respect to the former, recent evidence has shown that some changes in the amount of phytoplankton and the balance of species are widespread in the North-Eastern

28 Previously called Gymnodinium breve - 65 - Atlantic and thus unlikely to be due to nutrient enrichment, which is most intense in certain near- shore waters. Proposed explanations include changes in climate, water circulation around the UK, and grazing by zooplankton (Edwards et al. 2001; Beaugrand et al. 2002; Brander et al. 2003). Edwards et al. (2006) used the North Sea CPR time-series to investigate changes in total phytoplankton biomass (1948 – 2002) and dinoflagellate and diatom abundance (1958 – 2002). Their findings were that diatom abundance has declined in the North Sea since the 1960s (although the winter assemblage of diatoms has increased since the 1990s) and dinoflagellate abundance has increased. Within the time-series, Edwards et al. (2006) looked in detail at decadal changes in the genus Prorocentrum and Dinophysis and the abundance of Ceratium furca and Noctiluca scintillans and found that since the 1960s: species of Prorocentrum have become more abundant in the North Sea (particularly in Dutch coastal waters and the German Bight) during the 1990s; Dinophysis spp. were more abundant along the east coast of the UK during the 1970s but that abundance in west coast waters of Denmark had increased in the 1980s; the region of high C. furca abundance was now further north than during the 1990s; the abundance of N. scintillans has increased along the Dutch and English Channel coasts and in the northern Irish Sea. In relation to the frequency of blooms29 in the North Sea, Edwards et al. (2006) conclude that the late 1980s was a period of high bloom frequency in the central and northern North Sea but with the exception of Norwegian coastal waters, there were no significant long-term trends in bloom frequency. Edwards et al. (2006) attributed many of the patterns evident in the CPR time-series to hydroclimatic changes (e.g. salinity, sea surface temperature and stratification) and identified German and Dutch coastal waters and the northern Irish Sea as regions which might be sensitive to hydroclimatic influence and concluded that:

“Phytoplankton structural changes and blooms attributed to climate change could therefore be reinforced or accentuated by anthropogenic nutrient input into these areas.”

A number of other published studies link the occurrence of HABs and HAB species to regional scale climate induced changes in physical processes. It has been suggested that the intensity of Gymnodinium catenatum blooms on the NW coast of Spain could increase as a result of a climate driven increase in the intensity of coastal upwelling (Fraga & Bakun 1993). Tester et al. (1993) considered whether the distribution of Karenia brevis might be influenced by global warming causing changes in transport processes in the Gulf of Mexico and South Atlantic Bight.

29 A bloom was defined as species abundance greater than 2 standard deviations above the long-term (1958 – 2002) monthly mean and on the basis of statistical analysis. - 66 - The occurrence of Pyrodinium bahamense blooms in SE Asia has been related to the El Niño - Southern Oscillation (ENSO) which causes prominent inter-annual variability in weather and climate (MacLean 1989b; Azanza & Taylor 2001). Yin et al. (1999) related a series of HABs in coastal waters of Hong Kong during 1998 to what was described as one of the strongest El Niño years in the 20th century and suggested that:

“The entire event [red tides] coincided with the dramatic change in the oceanographic conditions of the northern portion of the South China Sea between 1997 and 1998…..The differences [in oceanographic conditions] are believed to be due to El Niño and responsible for setting up the physical oceanographic conditions which were favourable for the formation of harmful algal blooms along the south China coast.”

On a local scale, Belgrano et al. (1999) studied the variability in phytoplankton and primary production in the Swedish Gullmar fjord and found that the abundance of Dinophysis (acuminata, acuta and norvegica) was significantly related to the North Atlantic Oscillation Index, sea surface temperature and salinity and concluded that:

“There was an indication that higher densities of toxic phytoplankton species may be associated with the positive oscillations of the NAO as, for example, during the late 1980s as well as warmer SST-conditions and increased surface salinity…”

Moore et al. (2008) suggest that global warming may increase the geographical range of some warm water HAB species and the period of the year during which they occur might be expected to increase. As an example of the latter, these authors consider the seasonal occurrence of Alexandrium catenatum (responsible for PSP in shellfish) in Puget Sound (U.S., northwest coast). in excess of 13° C promote blooms of this species during late summer and early autumn and there is a seasonal window of  68 days. According to Moore et al. (2008), an increase of 2° C would almost double the seasonal window and if the predicted maximum increase of 6° C was realised, water temperature in the Sound would exceed 13° C on 259 days. Finally, using batch cultures, Peperzak (2003) simulated the expected effect of climate in 2100 on conditions in Dutch coastal waters (+ 4°C temperature rise and increased salinity stratification). The growth rate of two species, the diatom Skeletonema costatum and the cryptophyte Rhodomonas sp. was unchanged, two HAB species (the prymnesiophyte Phaeocystis globosa and the diatom Pseudo-nitzschia multiseries did not survive the culture conditions but the growth rate of the dinoflagellates Prorocentrum micans and P. minimum and the raphidophytes Fibrocapsa japonica and Chattonella antiqua exhibited higher growth rates (double) compared to present day conditions. Peperzak (2003) concluded that given the experimental constraints and uncertainties regarding future climatic conditions, his study - 67 - suggested that climate change will increase the likelihood of harmful dinoflagellate and microflagellate blooms.

3.3.2.5 Introductions and transfers of new species

The potential for human activity to inadvertently transfer a range of plants and animals from one region of the world to another is well documented and not a new phenomenon. Medcof (1975) reported the presence of a variety of invertebrates (including planktonic copepods, amphipods and polychaetes larvae) in the ballast water of a ship that travelled from Japan to Australia. One of the earliest examples relating to phytoplankton was documented by Ostenfeld (1908) who reported the sudden appearance of the tropical/ sub-tropical diatom Odontella30 sinensis in the North Sea in 1903 and concluded that:

“As far as I can judge there is no other explanation of B. sinensis’ sudden appearance in the North Sea than the following: it has been drawn in from afar by the aid of man, that is to say carried along from distant oceans by some ship, for instance attached to the outside or growing in the water of the hold.”

More recently, Hallegraeff and Bolch (1991) investigated the presence of dinoflagellate cysts in the sediment found in cargo vessel ballast tanks entering Australian ports and found that of the sediment samples collected from 80 vessels, 40 % contained viable dinoflagellate cysts of non toxic species and 6 % contained the cysts of the toxin producing species Alexandrium catenella and A. tamarense. In a further study, Hallegraeff and Bolch (1992) surveyed 343 vessels entering Australian ports and found that more than 200 of the vessels contained sediment in the bottom of their ballast tanks and of these 50 % contained dinoflagellate cysts. A similar study undertaken in Scotland (MacDonald & Davidson 1998) gave comparable results. Of 127 vessels that were in ballast entering Scottish ports, motile dinoflagellate cells were found in the ballast water of 76 % vessels and cysts were found in 61 % of sediment samples (a total of 92 sediment samples were collected). These studies clearly show that motile cells and resting cysts can be transported in ballast water and sediment. Hallegraeff and Bolch (1991) were able to grow a viable culture of Alexandrium tamarense from cysts collected from one vessel but pointed out that it was difficult to determine how often such an introduction would result in particular phytoplankters becoming established. With respect to the presence of Gymnodinium catenatum in coastal waters around Hobart in Tasmania (Australia), Hallegraeff and Bolch (1991) concluded that evidence from surveys of cysts in sediment cores and genetic studies pointed to the ‘distinct possibility’ that this species had been introduced.

30 Previously known as Biddulphia. - 68 - Hallegraeff (1993) used maps of the global distribution of PSP in 1970 and 1990 as an example of the global spread of a HAB species arguing that in the 1970s, PSP was restricted to Europe, North America and Japan but by 1990 was widespread throughout the southern hemisphere. Van Dolah (2000) prepared similar maps and suggested that they represented composite pictures which reflected an increase in reports of toxic events, geographical expansion, increased monitoring, research and improvements in detection have collectively resulted in the detection of toxins (and toxin producing species) in coastal areas where there were no previous records. Our view is that such maps should be used with caution and represent ‘snap shots’ of the reported distribution at a given point in time. The movement of cultured shellfish has been linked to the spread of HAB species in Japan. Pearl oysters are often relocated to areas considered better for growth and to avoid red tides and following one such relocation, Honjo et al. (1998) reported the sudden occurrence of Heterocapsa circularisquama blooms in coastal regions where the species was previously unrecorded. On the basis of a series of experiments in which pearl oysters were exposed to cultures of H. circularisquama, Honjo et al. (1998) concluded that viable H. circularisquama cells could be inadvertently transported with consignments of pearl oysters.

3.4 Case Studies

3.4.1 Introduction

A number of publications have been much cited in the scientific literature as evidence of a link between the occurrence of HABs and anthropogenic nutrient enrichment. Of these the papers by Anderson (1989), Hallegraeff (1993) and Smayda (1990) stand out. In his 1989 paper, Anderson cited the studies by Lam and Ho (1989) in Tolo Harbour (Hong Kong) and by Okaichi (1989) in the Seto Inland Sea of Japan. Smayda (1990) cites these two examples and gives the Baltic and Black Seas and the Southern North Sea as additional examples. Hallegraeff (1993) cited the Tolo Harbour, Seto Inland Sea, North Sea and Black Sea examples and gives an additional example: that of the brown tide phytoplankter Aureococcus anophagefferens in waters of Long Island Sound (Eastern seaboard of the United States). More recently, Anderson et al. (2002) repeated some of the examples noted above but also present examples from coastal waters of China (in addition to Hong Kong waters) and the United States. The following evaluation of case studies has been limited to four: coastal waters of China, Japan (primarily the Seto Inland Sea), the North Sea and continental coastal waters of the United States.

- 69 - 3.4.2 Coastal waters of China

3.4.2.1 Introduction

The occurrence of red tides in coastal waters of China has been widely reported and over the last few decades has resulted in substantial financial loss, particularly to the mariculture industry. However, this does not appear to have always been the case. Over twenty years ago, Holmes and Lam (1985) reported that there were few reports of red tides in the subtropical and tropical waters of South East Asia and that:

“Their impact in most parts of the region are so far, not as serious and widespread as in northern temperate waters.”

Holmes and Lam (1985) cite several authors as evidence of changes in the occurrence of HABs and conclude that:

“the frequency of occurrence, the variety of dinoflagellate species involved and the presence of toxic species are observed to have an increasing trend in recent years.”

3.4.2.2 Coastal waters of Hong Kong

Holmes and Lam (1985) present data on the occurrence of HABs from coastal waters of Hong Kong (Figure 3.6) and suggest that there has been an increase in HABs during the 1970s and up to 1983. They argue that while the increase could simply reflect increased monitoring, the data were collected as part of a routine (biweekly sampling) monitoring programme established in 1976 and made this unlikely. The increase was particularly apparent in Tolo Harbour with approximately half of the incidents occurring in this inlet and in relation to red tides; Holmes and Lam (1985) concluded that:

“Tolo Harbour has suffered particularly, and it may be concluded that urbanisation of this inlet’s catchment has played a role in this situation.”

- 70 - Figure 3.6 A map of Hong Kong showing the location of Tolo and Victoria Harbours.

An updated time series of red tide incidents in Tolo harbour was presented by Lam and Ho (1989) who report a two fold increase in nutrient loading (from livestock and domestic sewage) from 800 to 2,000 Kg N and 200 to 450 kg P d-1 between 1976 and 1986. The annual median concentration of dissolved inorganic nitrogen and phosphorus also increased. Nitrogen (as - -1 -1 nitrate, NO3 ) for example, increased from 0.005 mg l (0.36 µM) in 1977 to 0.135 mg l (9.6 µM) in 1986. Lam and Ho (1989) related the increase in red tide incidents to increased human population (Figure 3.7) and concluded that:

“The increase of red tides in Tolo Harbour is therefore a consequence of accelerated eutrophication in the marine bay following intensive urban development in the catchment.”

Lam and Ho (1989) reported blooms of the dinoflagellates Noctiluca scintillans, Prorocentrum triestinum, P. dentatum and P. sigmoides, an unidentified Gymnodinioid and a wide range of microflagellates and Yin (2003) included the dinoflagellates Gonyaulax polygramma and Prorocentrum minimum together with the diatom Skeletonema costatum and the photosynthetic ciliate Myrionecta rubra amongst the six most frequently occurring red tide species in the harbour. Most of the problems caused by blooms of these species have been

- 71 - associated with deoxygenation rather than biotoxins (Holmes & Lam 1985) indicating that the problems are associated with large biomass blooms.

Figure 3.7 The relationship between the number of red tide incidents per year in Tolo Harbour and the increase in human population (millions). Redrawn from Lam & Ho (1989).

18 0.6 Red tide incidents 16 Population 0.5 14 12 0.4

ncidents 10 0.3 8

Population 6 0.2 Red tide i Red tide 4 0.1 2 0 0 1976 1978 1980 1982 1984 1986

The 1989 paper by Lam and Ho is one of the earliest to present data on an increase in red tide incidents and nutrient enrichment. Accepting as argued by Holmes and Lam (1985) that the increase in HABs in Tolo Harbour was not due to increased monitoring and reporting, the data set does provide prima facie evidence for an increase in the occurrence of HABs having been driven by anthropogenic nutrient enrichment. However, it is important to consider this example in relation to the ecohydrodynamics of Tolo Harbour (especially if the harbour is to be used as an example for other coastal regions) and to put the data collected during the 1980s within the context of a longer time series. A 10 year data set is too short to determine whether the increase was part of a longer-term trend. Tolo Harbour is a long (15 km) narrow sea inlet which is 1 km wide at its entrance and has slow tidal exchange (Lam & Ho 1989). The surface area of the harbour is approximately 50 km2. The depth in the inner region is 2-3 m compared to 20 m in the outer part and the overall mean depth is  12 m (Li et al. 2004). The volume of Tolo Harbour is 0.6 km3. The euphotic zone typically extends to the sea bed such that there is generally sufficient light to support phytoplankton growth throughout the water column and throughout the year and the flushing time of the harbour is between half and one month (Li et al. 2004). Furthermore, Li et al. (2004) found that in years when there was a high occurrence of HABs total global radiation was higher. The physical characteristics of the Harbour and light climate are likely to provide conditions in

- 72 - which anthropogenic nutrient enrichment stimulates the production of high phytoplankton biomass and HABs. Li et al. (2004) present the Tolo Harbour time series for the period 1980 to 1999 and the data in Figure 3.8 (redrawn from Li et al. 2004) shows that the number of red tide incidents per year reached a peak of  30 incidents in 1988. Thereafter, the frequency declined. From 1991 to 1999 there were < 10 incidents per year and in 1998 and 1999, the number of incidents (3 and 2 respectively) was similar to the situation between 1980 and 1982.

Figure 3.8 The number of HAB incidents per year in Tolo Harbour Hong Kong between 1980 and 1999. (Redrawn from Li et al. 2004).

35

30 25

20

incidents 15

Red tide Red tide 10

5 0

1980 1981 1982 1983 1984 1985 1986 1987 1988 1989 1990 1991 1992 1993 1994 1995 1996 1997 1998 1999

An obvious question is whether there has been a corresponding reduction in riverine nutrient loading to Tolo Harbour and nutrient concentrations in the Harbour. According to data collected by the Hong Kong Environmental Protection Department, the total nitrogen load from rivers discharging into Tolo Harbour was stable at ≈ 6,000 Kg d-1 between 1986 and 1992, but by 2000 had fallen to < 2,000 t d-1. Figure 3.9 shows the data on red tide incidents given by Li et al. (2004) together with data on the annual mean concentration of DAIN (ammonium, nitrate and nitrite) in the inner harbour and recent red tide events in Tolo Harbour based on data from the Hong Kong Environmental Protection Department. The data in Figure 3.9 show that the relationship between the occurrence of red tides and the concentration of DAIN and DAIP in the inner harbour is complex with the general decrease in red tide frequency (between 1988 and 1993) taking place when the annual concentration of DAIN and DAIP remained high.

- 73 -

Figure 3.9 Temporal changes in red tide incidents and the annual mean concentration (µM) of DAIN (as ammonium, nitrate and nitrite) and DAIP (dissolved available inorganic phosphate) in Tolo Harbour between 1976 and 2006. A, red tide incidents and annual mean concentration of DAIN; B, red tide incidents and annual mean concentration of DAIP. Data from Li et al (2004) and the Hong Kong Department of Environmental Protection. (Nutrient data from stations in the inner part of Tolo Harbour.

35 30 A HABs 30 DAIN 25 ` 25 20 20 15 15

d tide incidents 10 nual DAIN mean

Re 10

An 5 5

0 0 1976 1979 1982 1985 1988 1991 1994 1997 2000 2003 2006

35 3.0 B HABs 30 DAIP 2.5 ` 25 2.0 20 1.5 15 1.0 ed tide Incidents ` ed tide Incidents

10 nnual DAIP mean

R A 5 0.5

0 0.0 1976 1979 1982 1985 1988 1991 1994 1997 2000 2003 2006

Other bays in coastal waters of Hong Kong also experience HABs but the frequency of occurrence is less than that reported for Tolo Harbour. In Victoria Harbour, there were only 21 red tides between 1983 and 1998 compared to 288 in Tolo Harbour during the same period (Yin 2003). It has also been suggested that nutrient input into some of these coastal areas cannot sustain the cell densities reported. In a recent study of a Scrippsiella trochoidea bloom in Port Shelter, Yin et al. (2008) concluded that ambient nutrient concentrations in the Bay were 6 -1 -3 insufficient to support the high biomass (15 x 10 cells L and 56 mg chlorophyll m ). It was assumed that a physical mechanism was responsible for concentrating the cells and causing the bloom in the Bay although Yin et al. (2008) suggested that dissolved organic nutrients may also play a role in bloom formation.

- 74 - 3.4.2.3 Other coastal regions of China

There is also concern regarding the occurrence and apparent spreading of HABs in other coastal regions of China (see Figure 3.10 for locations mentioned in the text). According to Qi et al. (1993a) red tides were first recorded in Chinese waters in 1952, although Tang et al. (2006) state that the first documented HAB event in China was in 1933 when a bloom of Noctiluca scintillans and Skeletonema costatum in coastal waters of Zhejiang killed marine organisms.

Figure 3.10 A map of China showing locations mentioned in the text.

On the basis of incomplete records, Qi et al. (1993a) estimated that there had been 169 red tides between 1980 and 1990 and concluded that:

- 75 - “There has been an increase in the frequency, magnitude and geographic extent of red tides along the coast of China recently.”

Qi et al. (1993a) stated that the Chiangjiang (Yangtze) estuary area was a region of high HAB occurrence. In August 1982, a bloom of Noctiluca scintillans extended over an area of 10 km 2, but a bloom of Rhodomonas sp. and Myrionecta rubra covered an area of 300 km2 in 1986 and in August 1988 a N. scintillans bloom extended over 6,100 km2. Tseng et al. (1993) report 91 red tide species (including 11 toxin producing species) from Chinese coastal waters. In June 1990, a bloom of Cochlodinium “type 90” occurred in coastal bays of Fujian province (Figure 3.10) which lasted ≈ 10 days, and caused substantial mortalities of marine organisms including benthic and pelagic fish and shellfish (Qi et al. 1993b). A large bloom of a Gymnodinium species occurred in coastal waters and shrimp ponds (20 x 106 cells L-1 in the ponds) in Bohai Bay, northern China during August and September 1989 and resulted in an economic loss of ≈ US $40 million (Xu et al. 1993). Qi et al. (2004) reviewed the occurrence of HABs in Chinese waters. An unusually high number of HABs occurred during late 1997 and 1998 in coastal waters of Guangdong province (including coastal waters of Hong Kong) in southern China and Qi et al. (2004) give the following details. In September 1997, a bloom of Phaeocystis globosa bloomed in Quanzhou Bay (Fujian province) and spread south. The bloom which lasted 6 months and covered an area of 3,000 km2 caused major losses of farmed fish estimated as US$7.5 million. During March and April, blooms of Gymnodinium mikimotoi and Gyrodinium aureolum31 occurred in coastal waters of the Pearl River estuary and Hong Kong respectively. Both resulted in mortalities of farmed fish with the Hong Kong bloom causing an economic loss of  US$12 million. In September, there were blooms of Scrippsiella trochoidea (0.63 x 106 cells L-1) and Ceratium furca and blooms of N. scintillans and Myrionecta rubra occurred in November. None of the September and November blooms resulted in fish kills although Qi et al. (2004) note that during the C. furca bloom a beach was closed. A number of coastal areas in China are enriched with anthropogenic nutrients and the variety of hydrographic regimes that occur can exacerbate the influence of nutrients on HAB species. For the period 1982 to 1987, Chen and Gu (1993) give a mean concentration of inorganic nitrogen from the mouth of the Changjiang (Yangtze) River of between 13.3 and 23.6 mg l-1 and an average concentration in Hangchow Bay of between 30.1 and 42.8 mg l-1[32]. Qi et al. (2004) stated that in 1997, 2.8 billion tonnes of sewage was discharged into the Pearl River estuary but only 10 % was given primary treatment. Qi et al. (2004) report that during a Karenia

31Both species are considered to be Karenia mikimotoi. 32 These values appear particularly high given that for the same period, Li et al. (2007) give annual mean nitrate concentrations of between 33 and 53 µM in the Changjiang (see Figure 3.3). - 76 - mikimotoi bloom that occurred at the mouth of the Pearl River in April 1998, the concentrations of total inorganic nitrogen and phosphorus were 211 µg l-1 (15.1 µM) and 7 µg l-1 (0.23 µM) respectively33. At the beginning of a Gyrodinium instriatum bloom in Shenzhen Bay, the nitrogen level was up to 977 µg l-1 (69.8 µM) while the concentration of ammonia nitrogen at the mouth of the Shenzhen River was 4,500 µg l-1 (321 µM). Qi et al. (2004) also suggest that intensive fish farming could also contribute to coastal nutrient enrichment and note that there were 110,000 fish cages in Guangdong province. As an example of the nutrient output from cage farming, Qi et al. (2004) refer to the situation in Ya- qian Bay, a small 23,500 m2 bay in the larger Daya Bay. In 1997, fish production was estimated as 132 t which required 1,056 t of feed and resulted in the release of 48.9 – 131.8 kg of nitrogen per tonne of fish produced or between 6.5 and 17.4 t of nitrogen per year. While there can be no doubt that HABs in coastal waters of China are a serious problem causing major financial loss and there has been anthropogenic enrichment of coastal waters, it is difficult from the studies discussed above to obtain an overall picture of the role of anthropogenic nutrients given the range of genera involved in HAB events in areas with different oceanographic conditions. It is likely however, that the continuous input of nutrients into these coastal waters will influence phytoplankton dynamics (species abundance and community structure) in this region. Qi et al. (1993a) were of the opinion that HABs were restricted to embayments and river mouths and that: “Nutrient enrichment and eutrophication of coastal water and estuaries are often the reasons for red tide initiation…”

Chen and Gu (1993) also considered anthropogenic nutrient enrichment one of a number of key variables in the formation of HABs:

“Factors causing red tides in the East China Sea are very complicated; but, in general, eutrophication combined with suitable physical factors (temperature, salinity, upwelling, light etc.) appears to be involved.”

Qi et al. (2004) suggest that a number of factors are involved in HAB initiation:

“Among these are: (1) climate change and temperature, (2) meteorological and oceanographic features and (3) anthropogenic influences in the form of excess nutrient loading.”

With respect to temperature, Qi et al (2004) noted that in Hong Kong waters a K. mikimotoi bloom was associated with a temperature increase from 21 to 25° C; a bloom of the same species in Nan-au, Dapeng Bay (east of Hong Kong) was associated with a further

33 We have assumed that the concentrations given are µg N or P and that the conversion to µM is by dividing by 14 and 31 for N and P respectively. - 77 - temperature change and a G. instriatum bloom occurred in Shenzhen Bay when the temperature rose from 22 to 28° C. According to Qi et al. (2004) some of these blooms followed increased precipitation and run off:

“In Shenzhen Bay, a heavy downpour from 26 to 27 April, 1998 recorded 126 mm of precipitation. This rain greatly reduced the salinity of the water around Shenzhen Bay to brackish level, and together with high nutrient levels and high temperature, induced the outbreak of Gyr. instriatum bloom.”

Recently, studies by Liu and Wang (2004), Tang et al. (2006) and Wang et al. (2008) have attempted to investigate long-term changes in HABs in Chinese coastal waters. Liu and Wang (2004) investigated the relationship between red tide outbreaks and urban development in coastal waters of Guangdong Province. Time-series of red tides in coastal waters of Guangdong province were related to increases in human population and productivity by Liu and Wang (2004) who concluded that:

“The faster the urban development is the more are the red tides. The red tide outbreaks match the urban development of the coastal cities not only spatially but also temporally. In the last two decades, the red tide outbreaks reached the first peak in the period of 1987-1992 and the second peak in 1998-2001.”

It is clear from Figure 3.11, however, that despite an increase in GDP following the first peak in red tide occurrence, there was a period of 5 years (1993 – 1997) when the frequency of red tides was similar to that in the early 1980s when GDP was at its lowest. Furthermore, the second peak in red tides was much lower than the first despite a higher GDP. A decrease in monitoring effort or a major reduction in nutrient input to coastal waters would explain these patterns although as noted above, of the 2.8 billion tonnes of sewage discharged into the Pearl River estuary in 1997 only 10 % was given primary treatment (Qi et al. 2004). An alternative explanation is that some other pressure is overriding the effect of enrichment and Liu and Wang (2004) suggest that:

“Three out of the four causes of red tide blooms on the coasts of Guangdong, described in the last section, are natural factors with the second one related to human activities. Red tide outbreaks are related to nutrient enrichment. Once there are suitable seawater temperature, salinity and weather conditions, red tides will take place.”

Although these natural factors do not appear to be related to the El Niño Southern Oscillation since the frequency of red tides was not significantly related (R2 = 0.0078) to the ENSO index (Figure 3.11C).

- 78 -

Figure 3.11 Changes in the occurrence of red tides and cultural development (as gross domestic product, GDP) in middle Guangdong province (China) and the relationship between the frequency of red tide occurrence and the El Niño Southern Oscillation Index (ENSO). A, red tide incidents; B, GDP; C, red tide incidence and ENSO Index. (Red tide and GDP data redrawn from Lui & Wang (2004); ENSO Index data from http://www.longpaddock.qld.gov.au/SeasonalClimateOutlook/SouthernOscillationIn dex/SOIDataFiles/index.html).

16 14 A 12 10 8 6

Red tide incidents 4 2 0 1980 1982 1984 1986 1988 1990 1992 1994 1996 1998 2000

2500 B 2000 1500

GDP 1000

500

0 1980 1982 1984 1986 1988 1990 1992 1994 1996 1998 2000

18 C 16 14 12 10 8 6 4 incidents Red tide 2 0 -15.0 -10.0 -5.0 0.0 5.0 10.0

ENSO Index

Tang et al. (2006) compiled information on a total of 435 records of HAB events between 1933 and 2004 in coastal waters of the southern Yellow Sea and East China Sea. The data set (Figure 3.12) were divided into time periods of pre 1980, the 1980s, 1990s and 2000 – 2004, to examine changes in the time of occurrence, location, causative species and bloom area.

- 79 -

Figure 3.12 Changes in the frequency of HAB occurrence in the southern Yellow Sea and East China Sea between 1980 and 2004. (Redrawn from Tang et al. 2006).

100 90

r 80 70 60 50 40

ber of HABs per yea 30

Num 20

10 0 1980 1985 1990 1995 2000 2005

According to Tang et al. (2006) there was a total of 32 causative species but three were dominant: Noctiluca scintillans in the 1980s; Skeletonema costatum in the 1990s (these two species were responsible for 73.2 % of HABs before 2000); Prorocentrum dentatum (donghaiense) between 2000 and 2004. The timing of HAB occurrence had also shifted from August to October prior to 1980, to July – August in the 1980s, May – July in the 1990s and May to June during 2000 to 2004. Most of the HABs occur in the region of the Changjiang (Yangtze) River mouth and two regions further south that are influenced by upwelling. Tang et al. (2006) consider nutrient enrichment to be a likely reason for the high number of HABs off the Yangtze River mouth (see data from Chen & Gu 1993, reported above). For the other two areas of high HAB occurrence, Tang et al. (2006) note that they coincide with coastal upwelling areas and suggest that:

“Upwelling could provide rich inorganic nutrients for HABs to form; it might also make HABs last for a long period and cover large areas……”.

However, Tang et al. (2006) were also of the opinion that:

“Nutrient-rich water from Yangtze River meets with the warm water from the Taiwan Strait to form a convergence zone that is usually favourable for the phytoplankton growth and HAB occurrences”.

Nutrient input to the coastal area from the Changjiang (Yangtze) River is also considered by Tang et al. (2006) to explain the shift in timing of HAB occurrence and cite studies by Han et al. 2003, Li et al. 2003 and Wang and Huang 2003 as evidence for the following scenario. High

- 80 - N:P ratios in the vicinity of the Changjiang (Yangtze) River mouth indicates that phytoplankton growth is phosphorus limited and Prorocentrum dentatum (donghaiense) has a competitive advantage over Skeletonema costatum in P limiting situations and a lower temperature range (18 to 22° C) compared to 25° C for S. costatum34. Therefore according to Tang et al. (2006) P. dentatum (donghaiense) out competes S. costatum in May when the phosphate concentration is low (and water temperature is lower) but with increased precipitation (and temperature) in June, there is an increase in the concentration of nutrients allowing S. costatum to grow quickly. Tang et al. (2006) conclude that:

“the HAB occurrence frequency has been accelerated…”.

We take this to mean that the frequency of HABs has increased, although in our opinion the data compiled and presented by Tang et al. (2006) and presented here in Figure 3.12 is not entirely consistent with this view. Assuming a constant level of monitoring between 1980 and 2004, the data clearly show an increase in the frequency of HABs between 1986 and 1992 (with a peak occurrence in 1990). This was followed by a decrease almost to pre 1985 levels followed by a marked increase from 1999 to a maximum in 2003. It is also interesting that the unusual number of HABs reported by Qi et al. (2004) for the more southerly coastal waters in 1998 is not reflected in the data set presented by Tang et al. (2006). Such a pattern is not consistent with an increasing trend in HABs driven by anthropogenic nutrient enrichment but is suggestive of other pressures which override the effects of nutrient enrichment. Wang et al. (2008) compiled data on the occurrence of HABs in the South China Sea between 1980 and 2003 but excluded data from Hong Kong and noted that HABs were not officially reported in Vietnam before the early 1990s. The data are presented here in Figure 3.13 (redrawn from Figure 5 of Wang et al. (2008) but without data from the western coastal region). Since the South China Sea is considered to be oligotrophic, it is presumed that these blooms were recorded in coastal waters. With a surface area of 3.5 x 106 km2 and a maximum depth of 5,000 m, the South China Sea is the largest semi-enclosed sea in the western tropical Pacific Ocean. According to Wang et al. (2008) and references cited therein, the summer (June to August) is dominated by the south western monsoon and an associated anticyclonic (clockwise) circulation and in winter the South China Sea is dominated by a strong north easterly monsoon which drives a cyclonic (anticlockwise) circulation. Freshwater input from the main rivers (Han, 16.9; Pearl, 336; Red, 137 and Mekong 475 km3 year-1) is 965 km3 year-1.

34 As noted in Part 1 it is likely that Skeletonema costatum in Chinese waters may well be S. costatum s.s. which is distinct from the S. costatum in European waters. - 81 -

Figure 3.13 Changes in the occurrence of HABs in different regions of the South China Sea between 1980 and 2003. (Redrawn from Wang et al. 2008).

50 Northern region 40 30 20 10 0 0 2 8 0 6 8 0 98 99 99 00 198 1 1984 1986 198 1 1992 1994 199 1 2 2002

50 Eastern region 40 30 20

10 0

4 0 2 8 9 9 982 988 9 996 9 002 1980 1 198 1986 1 1 199 1994 1 1 2000 2

50

40 Southern region

30 20 10 0

0 4 6 0 2 6 0 2 98 98 99 99 00 1 1982 1 198 1988 1 199 1994 1 1998 2 200

70

60 Total 50 40 30 20 10 0

980 982 984 986 988 990 992 994 996 998 000 002 1 1 1 1 1 1 1 1 1 1 2 2

For the South China Sea as a whole, Wang et al. (2008) found that most HABs (369) occurred in the southern region where the climate is tropical and HABs occur all the year round. For the northern region, most HABs were observed in coastal waters near the Pearl River estuary. Wang et al. (2008) were of the opinion that anthropogenic nutrient enrichment from land based sources and aquaculture promote the occurrence of HABs and stated that:

- 82 -

“waters are enriched by high inorganic nutrients in freshwater run off, sewage discharge, agricultural fertilizers, and nearby high density coastal aquaculture.” and that:

“Intensive aquaculture causes self-pollution as a result of excessive feeding and fish feces, causing eutrophication of the aquaculture area, thus providing suitable environmental conditions for algae to grow and blooms to occur in the region.”

As an example, Wang et al. (2008) state that in the western coast of Sabah (a Malaysian state on the northern part of the Island of Borneo) aquaculture production increased from 160,000 t in 1994 to 2 million t in 1998. However, with respect to the western region of the South China Sea, Wang et al. (2008) state that:

“HABs occur frequently in July – September along the coast of Binh Thuan Province of Vietnam, where eutrophication is not so serious but where there is strong regional upwelling of nutrients (Tang et al., 2004a, b) that has contributed to HABs.”

According to Wang et al. (2008)

“The occurrence of HABs increased and the affected area spread substantially from 1980 to 2001.” and:

“From 1990, some previously unobserved HAB species were observed.”

These authors also concluded that:

“eutrophication appears to be the key factor for HABs in coastal and bay waters, such as the Pearl River estuary (A in Fig. 11), Manila Bay and the northwest coast of Sabah….”

In a study of sediment cores from the South China Sea, Hu et al. (2008) found there had been a gradual increase in phytoplankton abundance from 1940. This increase accelerated after 1965 especially during the period from 1980 to 2000 and that chemical markers for domestic sewage exhibited a similar temporal change. This provides evidence for the influence of anthropogenic nutrients on phytoplankton abundance (part of the eutrophication process) but is not evidence of a link between enrichment and an increase in the frequency of HABs. Interpretation of the time-series presented by Wang et al. (2008) is complicated by the increase in monitoring effort which increased from ≈ 3 monitoring stations in 1980 to approximately 100 sampling stations and just over 1000 samples in 2004 (see Figure 3.5). Furthermore, as with the

- 83 - data presented by Lui and Wang (2004) and Tang et al. (2006), the data (Figure 3.13) presented by Wang et al. (2008) suggest that the occurrence of HABs is not solely determined by nutrient enrichment and that other factors override the influence of nutrient enrichment. The data show that for the Northern region there was a peak in HAB occurrence during the early 1990s with the highest frequency ( 25) in 1991. This was followed by a decrease to a frequency similar to that observed in the early 1980s and a second increase from 1998 to 2003. For the eastern region, there appeared to be a gradual increase to a peak occurrence of  17 in 1998 but between 2000 and 2003 there were fewer HABs than in the late 1980s. The occurrence of HABs in the southern region shows a similar bimodal peak as in the northern region, with a peak in the late 1980s and a second peak in the late 1990s. Combining the data for all of the regions (excluding the western region) suggests an increase in the frequency of occurrence of HABs in the mid 1980s and 1990s, with 1991 and 1998 as years of particularly high occurrence ( 58 and 67 respectively) but a reduction in the frequency of occurrence (≤ 20 per year) since 1999.

3.4.2.4 The influence of the seasonal monsoon and climate change

Yin and Harrison (2007) suggested that both nutrient and phytoplankton dynamics (including the occurrence of HABs) in coastal waters of Hong Kong are influenced by the seasonal monsoons and discharge from the Pearl River. As the second largest in China, the Pearl River is 2,214 km in length, has a catchment of 452,000 km2 and an average discharge of 10,524 m3 s-1, 80 % of which is discharged during the wet season between April and September. During summer when nutrient concentrations in surface waters of the upper region of the estuary are typically 90 µM

NO3, 120 µM Si and < 0.5 µM PO4 (Yin et al. 2001) coastal waters of Hong Kong are dominated by the Pearl River plume to the west of Hong Kong Island although the influence is least on the eastern side. According to Yin et al. (2001) during summer, phytoplankton growth in the Pearl River plume is limited by the availability of P (summer biomass is low, only occasionally exceeding 10 mg chlorophyll m-3, Yin 2003) although close to the estuary, waters are turbid and light limitation may also play a role. Whether or not phytoplankton growth in the plume is nutrient or light limited, it is evident that coastal areas most affected by the Pearl River discharge have fewer HABs than the coastal embayments of Tolo Harbour, Mirs Bay and Port Shelter (Yin 2003). During winter, the north east monsoon results in downwelling at the coast and the retention of coastal water within the coastal embayments and it is during this period (and early spring) that most red tides occur (Yin 2003). In addition to the seasonal monsoons, large scale climatic variation is also considered to influence the occurrence of HABs in this region. The exceptional Karenia mikimotoi bloom in Hong Kong waters in 1998 was related to the 1997/ 1998 El Niño (widely regarded as the most

- 84 - intense of the 20th century (see Isoguchi et al. 2005 and references cited therein). According to Yin et al. (1999) typical oceanographic conditions in coastal waters of Hong Kong result in the waters being well flushed but in 1998:

“… the coastal waters of the south China coast including Hong Kong became trapped along the coast. Given local eutrophied conditions of the China coast, the outbreak of harmful algal blooms occurred over a coast-wide scale (~ 400 km) in winter 1997 and spring 1998.”

Qi et al. (2004) also noted that 1998 was an unusual year for HABs in coastal waters of Guangdong Province and Hong Kong and Wang et al. (2008) noted that in addition to 1998, 1991 and 1995 were also El Niño years during which there was a high occurrence of blooms. Isoguchi et al. (2005) related phytoplankton bloom events (although these were characterised by -3 -3 chlorophyll concentrations > 1 mg m relative to typical concentrations of ≤ 0.5 mg m ) in the vicinity of the Spratley Islands in the South China Sea to the 1997/ 1998 El Niño and concluded that: “The long-term reanalysis winds over the eastern SCS [South China Sea] demonstrates that wind jet formation and associated offshore cooling/ bloom are expected to occur in most cases of the subsequent El Niño years”.

Finally, in a review of Pyrodinium bahamense blooms in Southeast Asia, Azanza and Taylor (2001) related the occurrence of P. bahamense blooms to El Niño events:

“Records show that the rise and initiation of Pyrodinium blooms in the Indo- West Pacific coincided with El Niño, e.g. 1976- Malaysia, Brunei; 1982-1983 Samar-Leyte, Philippines; 1987-1988 Manila Bay events…”.

3.4.2.5 Other human pressures

Yu et al. (2007) investigated the effects of warm water effluent from a nuclear power station on the occurrence of HABs in the small (550 km2) semi enclosed Daya Bay in the northern South China Sea by examining time series of temperature, chlorophyll and HABs. An increase in chlorophyll (yearly means of 1.9 and 3.8 mg mg-3 pre and post 1994 respectively) was associated with an increase (1.1° C) in annual mean water temperature and monthly mean temperature differences (pre and post 1994) of up to 3.5° C in May. According to Yu et al. (2007) before 1994 most HABs occurred in spring and autumn but after 1994 they occurred all year round and lasted longer. Yu et al. (2007) also reported that during the period 1986 to 1999, nutrient concentrations increased (total inorganic nitrogen from 0.021 mg l-1 (1.5 µM) to 0.068 (4.9 µM) and the N:P ratio shifted from 1.5 to  60. Yu et al. (2007) concluded that:

- 85 - “HAB frequency increased remarkably after 1994, particularly in May, which was associated with the increasing of water temperature and eutrophication around the Daya Bay.”

3.4.2.6 Summary

Data from Tolo Harbour provide evidence that anthropogenic nutrient enrichment caused an increase in the occurrence of red tides (large biomass blooms). The potential for P or light limitation of phytoplankton growth appears to limit the effects of enriched Pearl River water on HAB development in western coastal waters of Hong Kong. It is likely that the occurrence of HABs is modified by the seasonal monsoons and El Niño events. In some coastal areas, inorganic nutrient concentrations are insufficient to support large biomass blooms suggesting other processes such as physical concentration are important in HAB formation. For more open coastal regions of China the situation is less clear. Waters are clearly enriched in many coastal regions and HABs occur frequently and often result in major economic loss. Many studies provide detailed descriptions of blooms and their effects and allude to (nutrient enrichment [eutrophication]) as a cause of HABs. The studies by Liu and Wang (2004), Tang et al. (2006) and Wang et al. (2008) present time series of HABs that show marked changes in the occurrence of HABs but in our opinion, the time-series do not provide unequivocal evidence of long-term increases in HABs driven by nutrient enrichment. Interpretation is complicated by the increased level of monitoring (e.g. Wang et al. 2008) and the patterns evident in the data sets suggest that other factors such as coastal upwelling and climate change (ENSO) override the influence of anthropogenic nutrient enrichment on the frequency and magnitude of HABs in these coastal regions. On a local scale, an increase in HABs was related to human induced increases in temperature and nutrients in one coastal embayment.

3.4.3 Coastal waters of Japan

3.4.3.1 Introduction

Harmful algal blooms are not a recent phenomenon in coastal waters of Japan. There are historical accounts of blooms and both red tides and toxic episodes occur throughout coastal waters of Japan (Fukuyo et al. 2002). According to Kotani et al. (2001) < 20 % of red tides caused harmful effects. Fukuyo et al. (2002) report a similar figure and show that for the 60 – 110 red tides reported in the Kyushu area each year between 1979 and 1998, only 35 caused damage to fisheries > 10 x 106 yen (US$ 93,000, based on 1 US$ = 108 yen). A wide range of species are considered to be harmful. Fukuyo et al. (2002) list 12 main red tide species some of which (e.g. Heterocapsa circularisquama) are considered to be ‘novel’ that is, previously unrecorded in a particular coastal region (Yamaguchi et al. 2001). Toxin producing species - 86 - responsible for PSP (Alexandrium tamarense and A. catenella), and DSP (Dinophysis fortii and D. acuminata) are widespread in Japanese coastal waters (Fukuyo et al. 2002). Figure 3.14 is a map of the Seto Inland Sea showing the locations mentioned in the text. Time-series of red tide occurrence in three areas of western Japan: Kyushu area, Tosa Bay and the Kumano-nada are presented by Fukuyo et al. (2002) who concluded that there was little evidence of any change in the frequency of red tide events in these areas during the 1980s and 1990s. The Seto Inland Sea has been the focus of particular attention with respect to the occurrence of HABs and the increased frequency in occurrence in the Inland Sea is one of the two examples cited by Anderson (1989) and one of three examples cited by Hallegraeff (1993) as evidence of a link between an increase in the occurrence of HABs and coastal pollution. The Seto Inland Sea is the focus of this case study.

3.4.3.2 The Seto Inland Sea

Since 1957, when a large (15 km2) mixed bloom of Karenia mikimotoi, Coscinodiscus spp. and Chaetoceros spp. resulted in mass mortalities (presumably due to K. mikimotoi) of marine organisms, red tides have been increasingly reported from the Seto Inland Sea (Okaichi, 1997). Blooms of Chattonella antiqua occurred frequently (1972, 1977 – 1979, 1982, 1983, 1987) in the Harima-nada (Figure 3.14) of the Seto Inland Sea (Yanagi 1989 and see Okaichi 1997) and

Figure 3.14 A map of Japan showing the Seto Inland Sea and locations mentioned in the text.

- 87 - caused large scale mortalities of cultured Yellowtail fish (Ishio et al. 1989). Okaichi (1989) reported a large bloom in 1972 that was associated with the mortality of  14.2 million Yellowtail and an economic loss of 7.1 billion Yen ( US$ 70 million). Blooms of Heterosigma akashiwo caused serious damage to the aquaculture industry in the Seto Inland Sea during 1975 and 1981 (Yamochi 1989) and more recently, blooms of Cochlodinium polykrikoides have been reported from several coastal regions of Japan including the Harima-nada (Kim et al. 2004). Finally, Yamaguchi et al. (2001) reported the occurrence of Heterocapsa circularisquama as a relatively new (since 1988) harmful species, blooms of which have increased in intensity and geographical distribution and that this species has replaced others such as Chattonella antiqua, C. marina, and Heterosigma akashiwo as the dominant phytoplankter causing HABs in Japanese coastal waters. The following description of the Seto Inland Sea (Figure 3.14) is from Takeoka (1997, 2002). The Seto Inland Sea is semi enclosed, has a surface area of 21,827 km2 and mean depth of 37 m and a volume of 816 km3. There are approximately 600 small islands in the Seto Inland Sea and the sea is divided into many narrow channels (“Seto”) where tidal flows are strong (up 5 m s- 1 during spring tides). The bays are called “Nada”. Tides from the Pacific Ocean enter the two main channels, where volume transport is highest and meet in the central region of the sea (Hiuchi-nada) where volume transport is negligible. Residual currents are generally weak and tend to result in closed circulation within each of the bays and this limits exchange between them. During the winter (December to February) an eastward flow of water through the Seto Inland Sea is generated by the north 10 -3 westerly monsoon and results in a total volume transport of 3.3 x 10 m ( 4 % of the total volume of the sea). The residence time of water in the Sea is  1.2 years. The catchment area is relatively small and freshwater inflow of 44 km3 per year results in a mean salinity of 33. As a consequence of the low freshwater inflow, estuarine circulation is weak except in Osaka and Hiroshima Bays where larger rivers discharge. During summer, the water column in the bays is stratified but remains mixed in the channels. Tidal mixing fronts separate stratified and mixed waters and are considered important regions where warm oxygenated water is transferred into bottom waters and nutrient rich bottom water into the upper layer. High summer phytoplankton biomass can be associated with these tidal mixing fronts. The average euphotic zone depth is  18 m. Takeoka (1997) concluded that the Seto Inland Sea has efficient biological and fisheries production as a result of:

“the sea’s enclosed structure which keeps the nutrient concentrations high and by the role of the many straits as bypasses of heat, nutrients and oxygen which contribute to the rapid and repeated utilization of the nutrients.”

- 88 - Coastal areas of the Seto Inland Sea are amongst the most industrialised in Japan. Approximately 30 million people live in the Sea’s watershed and significant industrial development took place during the 1960s and 1970s. For example, in 1973, 40 and 44 % of the total Japanese production of refined oil and iron respectively took place around the Seto Inland Sea (Yanagi & Okaichi 1997). There has also been significant aquaculture development and Takeoka (1997) cites Nakanishi (1993) for the estimate of 380 x 103 t for the annual production of cultivated fish. As noted above, the enclosed nature of the Sea and the retention of nutrients was one reason given by Takeoka (1997) for the high productivity and high fisheries yield in the Seto Inland Sea. However, these features are likely to make it more vulnerable to anthropogenic nutrient enrichment. Many of the regions within the Seto Inland Sea were regarded as polluted and Hashimoto et al. (1997) considered the Harima-nada to be one of the most organically polluted regions. Okaichi (1989) presents data showing an increase in the annual occurrence of red tides between 1968 and 1986 (Figure 3.15A redrawn from Figure 2 of Okaichi, 1989). The data which are based primarily on a monitoring programme established in 1973 and operated throughout the Seto Inland Sea until 1986 (Note, Okaichi (1989) gives a date of 1976 in the text, but his Figure 2 shows data up to 1986) shows that there was a remarkable increase in red tides with a peak of  299 in 1976. Since the monitoring programme did not begin until 1973, it is unclear where the data from 1968 to 1972 are from and whether they are directly comparable with the later data. Okaichi (1989) does not provide any details of what criteria were used to define a red tide although according to Ichiro Imai (pers. comm.) the data are based on observations made every week by each of the 11 Prefectures (local governments) around the Seto Inland Sea with water discolouration being scored as a ‘red tide’. The data were subsequently compiled by the Fisheries Agency of Japan. There may have been some over reporting but the high numbers of red tides reported during the mid 1970s are considered to be reliable (Ichiro Imai pers. comm.). Okaichi (1989) also presents the number of red tides which caused fish kills over the same period. This increased three fold, from 12 in 1968 to 39 in 1971 but had declined to 13 by 1986. Fukuyo et al. (2002) present data on the number of red tides associated with fisheries damage for Japan as a whole (Figure 3.15B) and concluded that over the last three decades the annual occurrence had been stable. This low incidence of fish kills relative to the total number of red tides that occurred each year is consistent with the views of Kotani et al. (2001) and Fukuyo et al. (2002) that red tides which cause actual harm only represent  20 % of the total number of red tides in any one year.

- 89 -

Figure 3.15 The temporal trend in the occurrence of red tides in Japan. A, in the Seto Inland Sea between 1968 and 1986 (Redrawn from Okaichi 1989); B, red tides associated with fisheries damage in Japan between 1971 and 1998 (redrawn from Fukuyo et al. 2002).

350 A r 300

250

200 150 100

Number of red tides per yea 50 0 1968 1971 1974 1977 1980 1983 1986

60 B 50

40

30

ber of incidents 20 Num 10

0

5

1971 1973 1975 1977 1979 1981 1983 1985 1987 1989 1991 1993 199 1997

Imai et al. (2006) present the red tide time series of Okaichi (1989) but include data collected from 1967 and from 1987 to 2004. Figure 3.16 which is redrawn from Figure 4 of Imai et al. (2006) and includes data from before 1968 from Fukuyo et al. (2002) shows the marked increase in the occurrence of red tides and that since 1986, the number of red tides occurring in the Seto Inland Sea each year has been stable at approximately 100. Okaichi (1989) did not provide any detailed information on temporal changes in nutrient loading or concentrations but instead stated that:

“red tides increased in proportion with the development of industries in this region.”

- 90 - It is evident that the increase in red tides coincided with substantial increases in domestic and industrial waste input to the inland sea. The chemical oxygen demand (COD)35 loading increased from 925 t d-1 in 1962 to 1900 t d-1 in 1969 (Yanagi & Okaichi 1997) and the annual discharge load (the amount of nutrient load reaching the Seto Inland Sea) of nitrogen increased from  240 t d-1 in 1957 to  750 t d-1 in 1972. The discharge of phosphorus increased from  16 to 80 t d-1 (Sekine & Ukita 1997). The loading data of Sekine and Ukita (1997) together with more recent loading data from Imai et al. (2006) show that the increase in loading preceded the increase in red tides and that following the peak loading in the early 1970s, the N load to the Seto Inland Sea has remained stable at approximately 600 t d-1 (Figure 3.16). The total phosphorus loading has decreased from ≈ 65 t d-1 in 1979 to ≈ 40 t d-1 (Imai et al. 2006).

Figure 3.16 Changes in the occurrence of red tides (number per year) and nitrogen run-off load (kg d-1). The red tide data are from Figures 2 and 4 of Fukuyo et al. (2002) and Imai et al. (2006). The N-loading data are from Figures 6.4 of Sekine & Ukita, 1997 (blue filled circles) and Imai et al. (2006) (open blue circles) respectively.

300 800 Red tide frequency 700 250 N-load

y 600 200 500

150 400 N load 300 100 Red tide frequenc 200 50 100

0 0

1950 1952 1954 1956 1958 1960 1962 1964 1966 1968 1970 1972 1974 1976 1978 1980 1982 1984 1986 1988 1990 1992 1994 1996 1998 2000 2002 2004

The relationship between nutrient enrichment of the Seto Inland Sea and the occurrence of HABs is more complex than the simple enrichment HAB paradigm. For example, according to Suzuki (2001), Mikawa Bay has experienced the most serious dissolved oxygen deficiency compared to other bays but had a low level of nutrient input: between 1955 and 1970, the N and P load to the Bay doubled and tripled respectively; red tides only became a notable feature some

35 Chemical oxygen demand (COD) is commonly used to measure the amount of organic matter in water by measuring the amount of oxygen (mg L-1) generated when the organic matter is oxidised. COD is commonly used as a measure of water quality.

- 91 - five years after this increase; oxygen depletion increased markedly from 1975. Suzuki (2001) concluded that the reason for the “intensified eutrophication” was land reclamation during the 1970s when ≈ 1,200 ha of shallows including tidal flats were reclaimed and that completion of a canal resulted in an average of 20 % of the Toyokawa River flow into the Bay being diverted. The loss of tidal flats and filtration capacity of shellfish (because of land reclamation) was approximately equivalent to 19 to 31 % of the water exchange and diversion of the River inflow reduced the dilution rate of the Bay (as a result of reduced density flow) by 20 – 40 %. A number of studies report temporal changes in riverine nutrient loadings and concentrations and nutrient concentrations in the Seto Inland Sea. Hashimoto et al. (1997) suggested that despite the introduction of legislation to reduce organic pollution in the Seto Inland Sea in 1973, there was little evidence to suggest that there have been significant changes in the loadings of dissolved and particulate N and total P. The N loading data presented in Figure 3.16 are consistent with this view. In 1994, a new law was introduced to reduce N and P levels in the water of the Seto Inland Sea (Hashimoto et al. 1997) and effluent control of total-N was introduced in 1996 (Imai et al. 2006). Yamamoto et al. (2002) showed that a significant decrease in dissolved (NH4, NO3 and NO2) and total nitrogen in the Ohta River that flows into Hiroshima Bay only occurred after 1995 but that DIP and total phosphorus had decreased between 1980 and 1998. The DIP river load showed a significant decrease of 66 % when the mean 1980-1982 concentration was compared to the mean 1996-1998 concentration but for DIN, the 1995-1998 mean was 31 % less (18 % in the case of total-N) than the mean for the period under study. Assuming the situation in the Ohta River is reflected in other rivers discharging into the Seto Inland Sea, it is only relatively recently that the DIN loading has decreased. Imai et al (2006) report changes in dissolved inorganic nutrients in the north eastern region of the Harima-nada. For surface and mid water depths: ammonium decreased from  5 µM in

1975 to < 2 µM by 1979; NO3 remained relatively stable during the early to mid 1970s, increased to  6 µM by the late 1970s but decreased to approximately 2 µM by 2000; DIP decreased during the period 1978 to 1984 but increased up to 1992 and has since remained at a stable concentration of ≈ 0.4 µM; silicate concentrations varied between 5 and 10 µM between 1973 and 2000. The differential control of N and P inputs to the Seto Inland Sea has raised questions about whether altered nutrient ratios and the balance of dissolved and organic nutrients have resulted in the selection of particular HAB species (Yamamoto 2003). Imai et al. (2006) report trends in the dominant HAB species: blooms of Karenia mikimotoi, Heterosigma akashiwo and Noctiluca scintillans were most frequent during the 1970s but decreased thereafter; species of Chattonella (antiqua and marina) formed frequent red tides during the 1970s and 1980s but decreased during

- 92 - the 1990s (although blooms of these two species and a ‘novel’ species C. ovata appear to have increased in recent years). Imai et al. (2006) also suggest that the overall long-term trend is for a reduction in the frequency of red tides, although red tides of Heterocapsa circularisquama and Cochlodinium polykrikoides have increased in recent years. Yamamoto et al. (2002) concluded that:

“phosphorus reduction measure could have changed the species composition in the bay to those can be advantageous to survive even in such a low DIP concentration.” and further suggested that changes in the relative availability of inorganic and organic nutrients in particular and the ability of some species (e.g. Alexandrium tamarense and Gymnodinium catenatum) to utilise dissolved organic phosphorus (DOP) provides one explanation for the recent blooms of species which can utilise DOP. Some of the issues relating to the influence of nutrient ratios and organic nutrients on HABs and HAB species abundance are discussed in section 3.5 of this chapter.

3.4.3.3 Summary

The times series of HABs in the Seto Inland Sea and associated changes in nutrient loadings provides evidence for anthropogenic nutrient enrichment having led to an increase in the frequency of HABs. Efforts to reduce nutrient inputs to the Seto Inland Sea have resulted in a reduction in the number of HABs occurring each year, although the number (≈ 100 per year) remains much higher than before industrialisation in the 1960s. It has been suggested that nutrient ratios and the availability of dissolved and organic nutrients may have brought about changes in the species of phytoplankter causing red tides. The role of nutrient enrichment in the occurrence of ‘novel’ species (i.e. a species occurring in a coastal area which it had previously been unreported) is difficult to assess, in part because of the potential transfer of motile and resting stages in the ballast water of ships and through the practice of relocating shellfish during cultivation.

3.4.4 The North Sea

3.4.4.1 Introduction

The North Sea (Figure 3.17) is a large semi-enclosed sea in north-western Europe, receiving water from the Atlantic Ocean through the narrow English Channel and across a broad open boundary to the north. It also exchanges with the landlocked Baltic Sea, from which low-salinity water flows northwards as the Norwegian Coastal Current (NCC), entraining water from the

- 93 - North Sea proper before exiting to the ocean. Underneath the NCC is a trench bringing water from the continental slope. The remainder of the North Sea is a continental shelf water body with depths mostly less than 100 m (except on the northern margins), and less than 50 metres over much of its southern part.

Figure 3.17 A map of the North Sea showing regions mentioned in the text and a generalised circulation (white arrows).

The North Sea has been much studied for more than a century, and many reviews (e.g. Lucas 1941, 1942 and see also Reid et al. 1990 and references cited therein) and assessments of fishery resources (e.g. Savage 1931) are available making this regional sea one of the most intensely studied in the world. We draw particularly on the work of Rodhe et al. (2006) to distinguish the following open water regions: • a western and southern region in which strong tidal currents prevent seasonal

- 94 - stratification but in which freshwater discharges from large rivers create intermittent stratification and has been referred to as a region of freshwater influence ('ROFI'); • a central and northern region that displays seasonal thermal stratification; • an eastern region influenced by the Baltic outflow and containing the NCC, with persistent haline stratification.

Nearshore waters include:

• those in the broad, shallow, turbid, tidally drying estuaries found on the British coast and that of parts of France, Belgium, the Netherlands and Germany; • the long shallow Wadden sea, lying along the coasts of the Netherlands, Germany and Denmark, and sheltered by an island chain; • the deep, clear, haline-stratified fjords on the Norwegian coast.

Each of these water types has different physical conditions and shows (under pristine conditions) different seasonal patterns of phytoplankton biomass and composition, and each has, we believe, a different propensity to certain types of HAB. Much confusion has been caused in the scientific literature by failures to distinguish the different regimes (e.g. Smayda 1990, who assumed that changes in phytoplankton in Dutch and Nordic coastal waters could be extrapolated to the wider North Sea – see below). Under pristine conditions it seems likely that the source of nutrients in the North Sea was almost solely from Atlantic inflows. Even under present-day, nutrient enriched conditions, some three-quarters of the nitrogen supply (excluding nitrogen fixation) and nine-tenths of the phosphorus supply comes from this source (OSPAR 2000). It is widely accepted that there is substantial anthropogenic enrichment of the western-southern and eastern open-water regions, and of the Wadden Sea, many of the estuaries, and some of the fjords. There is evidence that the maximum winter nutrient levels resulting from this enrichment are falling as a result of increasingly stringent EC policies on the treatment of urban waste water and the control of diffuse inputs from agriculture.

3.4.4.2 Phytoplankton blooms in the wider North Sea

Early studies of North Sea phytoplankton reported the occurrence of HAB species. During 1898, Cleve (1900) observed Dinophysis acuta to the south of Iceland, northwest of the Faroe Islands, around the north of Scotland and in the Irish Sea. Lucas (1941, 1942 observed Pseudo-nitzschia and Dinophysis spp. in the wider North Sea. Dodge (1977) reported the presence of a number of Dinophysis species including acuta and norvegica together with Alexandrium tamarense in open waters of the North Sea in 1971. With respect to Phaeocystis, Lucas (1942) noted considerable inter-annual variability and concluded:

- 95 - “The whole series of records seems to show a clear-cut cycle. After being fairly abundant on the two southern lines [ships tracks] Phaeocystis became scarcer on the whole and then even more abundant…..in 1938-39. Whilst limited mainly to the east at first, it extended over to the west and then was limited to the east again.”

There have been a number of large and spectacular phytoplankton blooms in the North Sea. For example, a bloom of Ceratium furca occurred between July and October 1981 and extended from Belgian to Swedish coastal waters. In the German Bight, the bloom reached a peak of ≈ 0.5 x 106 cells L-1 (Gillbricht 1983). One of the largest in terms of geographical extent was the 1988 bloom of Chrysochromulina polylepis which extended over an area of 75,000 km2 (Granéli et al. 1993) and reached a density of between 5 and 10 x 106 cells L-1 (Maestrini & Granéli 1991). The spatial and temporal variation in phytoplankton blooms was evaluated by Reid et al. (1987) using Continuous Plankton Recorder (CPR) data from the North eastern Atlantic (including the North Sea) for the period 1958 to 1983. Blooms were defined on the basis of a species being present in 14 out of 20 microscope fields and colour (greenness) of the silk used to collect the phytoplankton and not cell abundance. Reid et al. (1987) concluded that:

“Within the area sampled by the CPR there has been a general decline in the incidence of phytoplankton blooms over the last 26 years; Gyrodinium aureolum may be an exception to this generalization.”

In their review of North Sea phytoplankton, Reid et al. (1990) concluded that there was no evidence for an increase in bloom frequency although the exception to this might be Phaeocystis and that some years (1968, 1971, 1977, 1978, 1982 and 1984) stood out. Reid et al. (1990) also stated that the CPR time-series (1931 to 1989) for the North Sea shows a progressive decline in Phaeocystis from 1948 to the present. In his review of HABs and nutrient enrichment, Smayda (1990) considered the wider North Sea and noted the lack of time-series measurements but considered that:

“Nonetheless, the more subjective trend analyses when combined into a regional overview suggest a long-term nutrient-production-bloom pattern for much of the North Sea analogous to that described for the Dutch and Nordic coastal waters.”

In our view, such an extrapolation is unjustified. First, the CPR data provide no evidence for a long-term increase in the occurrence of HABs. Second, such an extrapolation is not comparing like with like particularly in respect of Dutch coastal waters which represent a particular set of ecohydrodynamic conditions being shallow, turbid and tidally stirred.

- 96 - It is apparent that the phytoplankton of the North Sea undergoes cyclical change. Lucas (1941, 1942) alludes to such change possibly as a result of changes in flow through the English Channel and also inter-annual variability in the occurrence of Phaeocystis. The plankton of the North Sea has undergone a number of regime shifts, the most recent of which were in the late 1980s and late 1990s. According to Reid et al. (2001) after 1987, there was a marked change in North Sea plankton and increased phytoplankton colour (a proxy for phytoplankton biomass) which was attributed to incursion of oceanic waters into the North Sea from the north or through the English Channel. Such changes need to be considered when attributing long-term changes in HABs to anthropogenic nutrient enrichment in the North Sea and its coastal waters.

3.4.4.3 Phytoplankton blooms in coastal waters

For the German Bight of the North Sea, Radach et al. (1990) presented a detailed analysis of a 23 year time series (1962 – 1984) of meteorological, nutrient and phytoplankton data for the German Bight. According to Radach et al. (1990) both air and sea surface temperature had increased by  1° C during the time-series but there were no trends in other meteorological and hydrographic variables. The winter concentration of nitrate exhibited an upward trend and in 1984, the winter maximum was  25 µM. In contrast, there was a decreasing trend in silicate concentration between 1966 and 1984. As a consequence of these changes, Radach et al. (1990) reported an increase in the molar N:Si ratio from 1-2 in the late 1960s to 4-8 in the early 1980s. Winter molar ratios of N:P were in the range 16-32 in the early 1960s; 4-16 in the 1970s and 16- 32 in the mid-1980s. These changes in nutrient concentrations and nutrient ratios have been associated with changes in phytoplankton biomass (Gillbricht, 1988; Radach et al. 1990; Hickel et al. 1993) and seasonal succession. In the 1960s, diatom abundance peaked in July, followed by a peak in flagellates which reached approximately the same biomass. In the early 1980s, diatoms peaked in April at nearly three times the former biomass. The flagellate peak remained in August, but was also  3 times its previous level. The flagellates included dinoflagellates, especially Ceratium spp. and small naked types (Radach et al. 1990). Hickel (1998) reanalysed the Helgoland time-series (1962 - 1994) and was of the opinion that there was clear evidence of nutrient enrichment but that the expected long-term trends in phytoplankton were not always clearly represented. Recurrent 3 - 5 year cycles of diatom and flagellate biomass were reported by Hickel (1998) but separation of the phytoplankton into two size fractions (nanoplankton < 20µm and microplankton > 20 µm) showed that the three fold increase in phytoplankton was largely due to an increase in the winter biomass of nanoplankton. Since light limits phytoplankton growth during the winter in the German Bight, Hickel (1998)

- 97 - was of the opinion that: the flagellates were mostly heterotrophic and mixotrophic species < 5µm in size; their increase was not significantly correlated with inorganic nitrogen but some other compound. The explanation for the increase in flagellates was not known but coincided with other large scale effects and Hickel (1998) concluded that:

“It became apparent that neither diatom biomass, nor dinoflagellate biomass without the nanoplankton component showed a clear long-term upward trend, possibly due to the enormous inter-annual variations which might have masked minor trends.” and was of the opinion that the large inter-annual variability was due to the hydrography of the region which is dominated by a convergence between continental coastal water and North Sea water. During the summer, the latter is stratified and according to Hickel (1998) is the site of large dinoflagellate blooms and whether or not the blooms reached Helgoland was dependant on the proximity of the fronts. Detailed observations on the occurrence of HABs and HAB species in the German Wadden Sea between 1989 and 1992 are summarised by Nehring et al. (1995). Blooms of Phaeocystis spp. (up to 100 x 106 cells L-1 and associated with foam and mucilage) and Noctiluca scintillans (up to 7 x 104 cells L-1) were considered to be a regular occurrence. Oxygen depletion and mortalities of fish were associated with large blooms of Ceratium spp. but toxin producing species generally occurred in low abundance. Nehring et al. (1995) considered that these blooms were associated with more offshore stratified and frontal regions, which is consistent with the view of Hickel (1998). In fact, on the basis of his earlier work, Hickel (1998 and references cited therein) suggested that the enrichment of the inner Bight can only enhance phytoplankton stock if stratification (and creation of stable well illuminated surface layer) extended over large areas of the Bight. The dynamics of Phaeocystis pouchetii blooms in the German Wadden Sea was studied by Weisse et al. (1986) in 1975, 1976 and 1981. Post spring diatom blooms of P. pouchetii occurred in each year ( 28 x 106 cells L-1 in 1975; 48 x 106 cells L-1 in 1976; 32 x 103 colonies L-1 in 1981) and Weisse et al. (1986) were of the opinion that Phaeocystis pouchetii formed regular blooms in spring and summer in the German Wadden Sea; that the blooms developed when nutrient concentrations were low and:

“when particularly inorganic phosphate had dropped to minimal concentrations.”

- 98 - 3.4.4.4 Phaeocystis in the North Sea

According to Cadée and Hegeman (1986) Phaeocystis was discovered in 1882 when it bloomed in waters off northern Norway. Cleve (1900) noted that Phaeocystis pouchetii was very common around the Faroes and west of Scotland in 1898. In a study of phytoplankton species in coastal waters of the English Channel (2.5 miles off Plymouth) during 1915 and 1916, Lebour (1917) found that:

“…Phaeocystis, which is so abundant here in May and June that it interferes with everything, clogging up all the nets.”

In 1923, Orton (1923) described the aesthetic effect of Phaeocystis on the oyster beds of the Thames as ‘baccy [tobacco] juice’ and Savage (1931) wrote that in April 1926:

“The predominant feature of the plankton was the almost universal presence of the flagellate Phaeocystis pouchetii in large quantities. The meshes of the plankton nets were clogged with it at most of the stations, and as a consequence it was difficult to estimate the numbers of the plankton species with any degree of accuracy.”

Savage (1930) presented details of Phaeocystis distribution in the southern North Sea during April 1924 and 1926 and November 1927 (Figure 3.18 based on the data in Savage 1930). It is important to comment on sampling methodology and the quality of the data on which these plots are based. Each plot is the settled volume of all plankton collected by vertical hauls (April 1924 and November 1927) and horizontal hauls (April 1926). The data collected by vertical haul were divided by the depth over which the net was hauled to give settled volume per metre and the data collected by horizontal tows divided by an average volume per 10 minute haul. Furthermore, in November 1927 a Henson rather than an international standard net was used and the two nets had different collecting efficiencies. According to Savage (1930) the Henson net was considered to be 10 times more efficient and the data derived from using the international stand net were adjusted accordingly. Finally, the estimate of Phaeocystis abundance was as settled volume of all the plankton but according to Savage (1930) because Phaeocystis was very flocculent it was possible to identify the area where Phaeocystis was most abundant. However, as acknowledged by Savage (1930) the plots are a rough approximation of Phaeocystis abundance. In April 1924, Phaeocystis was clearly most abundant to the south east of the southern North Sea close to the Dutch coast. According to Savage (1930) in April 1926:

“The enormous quantities present on this occasion caused the surface of the sea to present a muddy appearance, and great difficulty was experienced in working the nets in the centre of the zone.”

- 99 - Savage (1930) estimated that in November 1927, Phaeocystis was distributed over a broad area of approximately 160 km and noted that Hardy (Hardy 1925 cited in Savage 1930) recorded the presence of Phaeocystis in the North Sea during November 1922.

Figure 3.18 Plots of the distribution of Phaeocystis spp. in the southern North Sea. A, April 1924; B, April 1926; November 1927. (From the data in Tables 1 and II of Savage 1930).

54 A B

54

53

53 Latitide

52

52

51 0.3 1.2 2.1 2.9 3.8 4.7 0.31.22.12.93.84.7 Longitude Longitude 55 C

55

54

54 Latitude 53

53

52 0.31.22.12.93.84.7 Longitude

Despite these early records, there is clearly some confusion in the literature regarding Phaeocystis in the North Sea. Smayda (1990), Hallegraeff (1993) and Anderson et al. (2002)

- 100 - make reference to the sudden appearance of Phaeocystis blooms in relation to nutrient enrichment and changes in nutrient ratios. Smayda (1990) stated that:

“Mass occurrences began in 1977...... ”

Hallegraeff (1993) referred to:

“The remarkable increase of foam-producing blooms of the prymnesiophyte Phaeocystis pouchetii (Hariot) Lagerheim, which first appeared in Dutch coastal waters in 1978, is probably the best-studied example of this phenomenon (Lancelot et al. 1987).” and Anderson et al. (2002) stated that:

“Mass occurrences of this species began in 1977 in the North Sea (Cadée and Hegeman, 1986)”

In reference to this, it should be noted that the study of Cadée and Hegeman (1986) was undertaken in the tidal inlet to the Marsdiep region of the Wadden Sea and not open waters of the North Sea. Furthermore, the statements referring to the sudden appearance of Phaeocystis blooms in the late 1970s by Smayda (1990), Hallegraeff (1993) and Anderson et al. (2002) are clearly incorrect. In reference to Phaeocystis spp., Cadée and Hegeman 2002 pointed out that:

“Apparently such blooms are a natural phenomenon in the Marsdiep area, and Phaeocystis is not a novel bloom-forming alga since the 1970s as suggested by Smayda (1998).”

With respect to the German North sea coast, Weisse et al. (1986) stated that:

“Conspicuous mass occurrenes of Phaeocystis pouchetii are known since the last centuary (Pouchet, 1892; Lagerheim, 1896), and have also been reported from the German North Sea area (Scherffel, 1899, 1900).

Anderson et al. (2002) may have been referring to summer blooms since they cite Riegman (1995) who discusses summer blooms and not to the spring Phaeocystis blooms that occur after the spring diatom bloom and which Cadée and Hegeman (2002) considered to be a natural phenomenon in the Marsdiep. Riegman (1995) stated that:

“Novel summer blooms of Phaeocystis appeared in the late seventies.”

This is also confusing because during a study between February 1974 and September 1976, Cadée and Hegeman (1979) recorded Phaeocystis abundance (> 5 x 106 cells L-1) in April, May, June and July. The data in Table 1 of Cadée and Hegeman (1986) shows that summer (June – September) peaks in Phaeocystis spp. abundance occurred every year between 1973 and 1985

- 101 - (except 1977 and 1981) and that in some years there were two summer peaks. Cadée (1990) also reported that Phaeocystis spp. formed blooms during the summer although they were smaller than the spring Phaeocystis spp. blooms. It is also clear from Savage (1930) that the occurrence of Phaeocystis spp. in the southern North Sea was not restricted to the spring. The occurrence of summer Phaeocystis spp. blooms before the late 1970s is also noted by Peperzak (1993). Cadée and Hegeman (1986) present data collected between 1973 and 1985 on the seasonal and inter-annual variability in the spring bloom of Phaeocystis pouchetii in the Marsdiep tidal inlet to the western Wadden Sea. During the first few years of the time-series the peak of the bloom was  4 x 107 cells L-1 but after 1977, peak abundance was higher (≈ 108 cells L-1 in 1978, 1980, and 1984 and 1.9 x 108 cells L-1 in 1985). Cadée and Hegeman (1986) also noted that P. pouchetii blooms were absent from the Marsdiep in 1969 which is consistent with Gieskes and Kraay (1977). The timing of the bloom (i.e. after the diatom spring bloom) was assumed by Cadée and Hegeman (1986) to be related to limitation of diatom growth by silicate with sufficient phosphorus and nitrogen remaining for P. pouchetii to grow and the bloom to continue until N or P became limiting. Increased P. pouchetii abundance during the spring peak, the duration of the spring P. pouchetii bloom and increased summer abundance over the 12 year time-series, led Cadée and Hegeman (1986) to conclude that given nutrient enrichment and elevated production in continental coastal waters:

“It seems justified, but difficult to prove, to relate the Phaeocystis increase to eutrophication.”

Lancelot et al. (1987) expressed a similar view:

“Blooms of the planktonic alga Phaeocystis pouchetii in the continental coastal zones of the North Sea have been observed to occur more and more frequently and intensively over the past twenty years, probably as a result of nutrient enrichment from river discharge.”

A further review of the Marsdiep time-series but including data up to 2000 was undertaken by Cadée and Hegeman (2002). Figure 3.19 (redrawn from Figure 4A of Cadée & Hegeman, 2002) shows a marked increase in the duration of blooms (defined as the number of days when Phaeocystis abundance was > 106 cells L-1) occurred in the late 1970s, reached a peak during the late 1980s/early 1990s and has since declined. The data presented by Cadée and Hegeman (2002) provide convincing evidence of an increase in spring Phaeocystis blooms in response to nutrient enrichment. More recently, using a combination of observational data and model simulations, Lancelot et al. (2009) have shown (their Figure 4) a positive link between the simulated maximum of Phaeocystis cells and DIN and DIP inputs to Belgian coastal waters. Interestingly, Lancelot et al. (2009) concluded that - 102 - although the changing nutrient loads had modified annual primary production and the relative contribution of diatoms and Phaeocystis spp. to annual primary production and its transfer to higher trophic levels, the sustained pressure of anthropogenic nutrients had not substantially modified the structure and function of the ecosystem. The relationship between nutrient ratios and Phaeocystis spp. blooms is discussed later in this section.

Figure 3.19 Changes in the duration of Phaeocystis spp. blooms between 1974 and 2000. (Redrawn from Cadée & Hegeman 2002) who defined a bloom as the number of days when Phaeocystis spp. abundance was > 106 cells L-1).

200 180

) 160 140 120 100 80 60 40

Duration bloom(days of 20

0

1974 1976 1978 1980 1982 1984 1986 1988 1990 1992 1994 1996 1998 2000

3.4.4.5 The role of climate change and anthropogenic nutrient enrichment

The enrichment - Phaeocystis spp. bloom hypothesis does not provide a complete explanation for the observed changes in the ecology of Phaeocystis spp. On the basis of an evaluation of CPR data from an area of the southern North Sea close to Dutch coastal waters, for the period 1948 to 1975, Gieskes and Kraay (1977) were of the opinion that:

“The record of Phaeocystis poucheti (Fig. 6) suggests that this colony-forming µ-flagellate (Haptophyceae) has become less abundant during the last decade……”

In reviewing the same data set, Cadée and Hegeman (1986) stated that:

“Whereas marked Phaeocystis peaks occurred from 1948 to 1964, they were less pronounced or even absent from 1969 to 1975.”

This is the period of rapid enrichment of Dutch coastal waters (see above). It is also evident that if it is assumed that natural Phaeocystis blooms were between 50 to 60 day duration (Cadée &

- 103 - Hegeman 2002) then, at a time when nutrient concentrations were increasing (the 1970s) Phaeocystis spp. blooms were of shorter duration than natural blooms (Figure 3.19). Furthermore, the large increase in bloom duration which took place at the end of the 1970s, occurred after the main increase in nutrients. Gieskes et al. (2007) considered the role of climate in long-term changes in Phaeocystis spp., using CPR data covering the last 5.5 decades and were of the opinion that:

“The higher frequency of occurrence and longer growing season of Phaeocystis before 1965, when eutrophication was still quite low (Philippart et al. 2000), is particularly striking (Figs. 2, 5). Actually, colony abundance was already reported to be extensive more than 80 years ago (Savage 1930). This also suggests that Phaeocystis abundance is not determined by anthropogenic nutrient input only.”

However, the counter to this is that Savage (1930) was referring to blooms in open southern North Sea waters and not the Dutch Wadden Sea. Gieskes et al. (2007) concluded that:

“Frequency was especially high before the 1960s and after the 1980s, i.e., in the periods when anthropogenic nutrient enrichment was relatively low.” “Changes in eutrophication have obviously not been a major cause of long- term Phaeocystis variation in the southeastern North Sea, where total phytoplankton biomass was related significantly to river discharge.”

Gieskes et al. (2007) presented evidence to support their view that the abundance of Phaeocystis species in the southern North Sea is to a large extent determined by the amount of Atlantic Ocean water flushed in through the Dover Strait. The question is whether climate change has played a role in the changes that are apparent in coastal phytoplankton production and species assemblages. Breton et al. (2006) considered that:

“Whether long-term changes in Phaeocystis colony blooms in the Southern Bight of the North Sea are due to climate (Owens et al. 1989) and/or human- induced nutrient enrichment of coastal waters (Cadée and Hegeman 1991) is still the subject of debate.”

and were of the opinion that their analysis of data from Belgian coastal waters provided evidence that climate (driven by the North Atlantic Oscillation, NAO) and enrichment combined to influence the magnitude of spring blooms of Phaeocystis spp. Thus, Breton et al. (2006) suggest a cascade effect with large scale variation in the NAO influencing local (Belgian coastal region) meteorology that in turn influences local hydrographic conditions including the

- 104 - geographical spread of riverine nutrient loads in the coastal region and therefore determines the level of winter nitrate enrichment. Finally, results from the International Council for the Exploration of the Sea (ICES) workshop on Time Series Data Relevant to Eutrophication Environmental Quality Objectives (ICES, 2007) also raise serious doubts over a generic link between species of Phaeocystis and anthropogenic nutrient enrichment. Key conclusions from the workshop include:

“There is no convincing evidence, except in Belgian coastal waters, that harmful algal blooms and red tides, either in their intensity or bloom-species selection, are generally linked to eutrophication processes, to elevated nutrient concentrations, or to altered nutrient ratios at the time series locations evaluated.” and that:

“Blooms of Phaeocystis globosa in the Belgian coastal waters and in the Wadden Sea are an arguable, and possibly unique exception to this general finding….”

3.4.4.6 Summary

There is no evidence for an increase in the occurrence of HABs in the northern and central (seasonally stratified) North Sea. For the inner German Bight at Helgoland it is likely that HABs are fuelled by enrichment but that hydrographic conditions override the effects of nutrients. Blooms of Phaeocystis species in the southern North Sea are natural events and whether there was a sudden appearance of summer Phaeocystis spp. blooms in the mid to late 1970s seems unlikely. For Dutch coastal waters, an increase in the duration of spring Phaeocystis spp. blooms has been related to nutrient enrichment. Similarly, for Belgian coastal waters, nutrient enrichment has brought about an increase in the size of Phaeocystis spp. booms. Climate change has resulted in regional variation in the abundance of Phaeocystis species in the North Sea.

3.4.5 Coastal waters of the continental United States of America

3.4.5.1 Introduction

The issue of coastal eutrophication and the occurrence of harmful algal blooms are viewed as an important socio-economic and environmental problem in the U.S. (see for example Heisler et al. 2008). Under some scenarios, nutrient inputs are expected to increase. Howarth (2008) for example, has suggested that:

“The Susquehanna is the single largest source of nitrogen to the Chesapeake Bay. Given the climate change predictions for increased precipitation (Najjar

- 105 - 1999; Najjar et al. 2000) and assuming no change in NANI36 or land use, an increase in nitrogen flux down the Susquehanna of 17 % by 2030 and 65 % by 2095 is predicted…”

This is in contrast to other regions of the world such as northwest Europe where there is evidence that nutrient loadings are stable or decreasing. The HAB issue has attracted the interest of many scientists in the U.S. for over four decades and is still a widely researched and debated topic. HABs are not a new phenomenon in the U.S. (see for example Rounsefell & Nelson 1966) and a number of studies allude to an increase in the occurrence of HABs. Hoagland et al. (2002) were of the opinion that:

“During the last several decades, harmful algal bloom (HAB) events have been observed in more locations than ever before throughout the United States……Whatever the reasons, virtually all coastal regions of the U.S. are now regarded as potentially subject to a wide variety and increased frequency of HABs.” and more recently a meeting sponsored by the U.S. Environmental Protection Agency in 2003, (reported by Heisler et al. 2008) concluded that:

“Degraded water quality from increased nutrient pollution promotes the development and persistence of many HABs and is one of the reasons for their expansion in the U.S. and the world. ”

3.4.5.2 Nutrient enrichment HAB relationships in coastal waters of the U.S.

The relationship between HABs and anthropogenic nutrient enrichment in U.S. coastal waters has recently been reviewed in detail by Anderson et al. (2008) and the following is largely from their review. For north eastern coastal waters of the Gulf of Maine, large scale and small scale blooms occurred. The large scale blooms are supported by an oceanic supply of nutrients without significant input of nutrients from anthropogenic sources. For the small scale blooms the linkage to nutrient enrichment seems more likely but Anderson et al. (2008) concluded that:

“a separate analysis in each area would be needed to assess whether eutrophication is affecting A. fundyense blooms.”

With respect to brown tides of Aureococcus anophagefferens in the northeast and mid Atlantic coastal region, these were considered by Anderson et al. (2008) to be an indirect result of enrichment through the input of dissolved organic nitrogen. In the mid Atlantic region (Chesapeake Bay), blooms of Prorocentrum minimum and several other species were considered to be the result of nutrient enrichment. Marshall et al. (2005) were of the opinion that blooms of

36 Net anthropogenic nutrient input - 106 - Cochlodinium polykrikoides in the Bay and tidal tributaries and Microcystis aeruginosa in tidal tributaries had increased over the last decade. For the Gulf of Mexico, and the occurrence of Karenia brevis blooms, Anderson et al. (2008) presented both sides of the argument concerning the possible stimulation of blooms by anthropogenic nutrients and concluded that:

“clear evidence to support hypotheses about increased bloom frequency and biomass on the west Florida shelf is still not yet available.”

In contrast, Anderson et al. (2008) were of the opinion that:

“Blooms of K. brevis along the Texas coast, which are influenced by major nutrient loads from the Mississippi River, have been more clearly linked to stimulation from land-based sources.”

For the occurrence of Pseudo-nitzschia and Alexandrium catenella in Californian coastal waters, Anderson et al. (2008) concluded that:

“There is no consistent evidence that Pseudo-nitzschia blooms are correlated with run-off events, nor is there direct evidence for trace metal limitation or stimulation of DA [domoic acid] during most blooms….”

Coastal waters of California are dominated by upwelling which introduces nutrients to the near coastal region and for both species, bloom events are linked to large scale oceanic processes, particularly upwelling events (Anderson et al. 2008). With respect to Heterosigma akashiwo in the Pacific northwest (Puget Sound) of the U.S., Anderson et al. (2008) stated that:

“In the absence of directed studies to test influences of anthropogenic nutrient enrichment, linking nutrient loading to blooms of H. akashiwo remains an elusive possibility in the Pacific Northwest.”

With respect to Alexandrium catenella, Trainer et al. (2003) investigated long-term changes in the occurrence of paralytic shellfish toxins (PSTs) in Puget Sound (U.S.) by comparing mean decadal levels of PSTs (µg saxitoxin equivalent/ 100g tissue) based on data collected as part of a Washington State Department of Health surveillance programme. Although the programme has changed (number of sampling sites and the species used to determine levels of toxicity in shellfish tissue), Trainer et al. (2003) found that compared to the 1950s to 1970s, the level of PSTs were significantly higher (t-test, P = < 0.001) during the 1980s and 1990s. Furthermore, the maximum mean decadal PST level was significantly correlated (r2 = 0.987) with the human population in all counties bordering Puget Sound. As pointed out by these workers, statistical correlation does not establish cause and effect but is suggestive of pressure associated with human population growth (such as increased nutrient supply) having influenced the increase.

- 107 - On the basis of their analysis, Trainer et al. (2003) were of the opinion that the increase in shellfish closures was not an artefact of the number of samples collected and concluded that:

“1) There has been a significant increase in the magnitude of PSTs in Puget Sound shellfish with time. 2) The geographical scope of shellfish closures caused by high levels of PSTs in Puget Sound has increased over the past four decades.”

In considering the cause of this increase, Trainer et al. (2003) were of the opinion that:

“Global climate changes, such as the Pacific decadal oscillation and increased eutrophication in nearshore areas, are possible explanations for the increased magnitude of PSTs in shellfish today.”

In summary, there appears to be little supporting data for the role of anthropogenic nutrients in promoting the small coastal blooms in the Gulf of Maine or the growth of HAB species advected into coastal waters of California and the links remains hypothetical. Therefore setting these aside it is clear that of the eight coastal regions of the U.S. reviewed by Anderson et al. (2008), only in three (possibly 5) has the link between the occurrence of HABs and anthropogenic nutrient enrichment been established. If there has been a recent (last 20 – 30 years) increase in the occurrence of HABs in U.S. coastal waters (Hoagland et al. 2002; Heisler et al. 2008) then either the studies have not been undertaken to relate occurrence to nutrient enrichment (as pointed out by Anderson et al. 2008) or other factors such as the spreading of species, increased observation and reporting may be involved. The conclusions of Anderson et al. (2008) are summarised in Table 3.1.

Table 3.1 A summary of the main findings of Anderson et al. (2008).

Region Evidence of a link Comment NE – Gulf of Maine No Conjecture in relation to blooms in bays and harbours, but data lacking. NE and mid Atlantic Indirect Possible importance of DON not DIN Mid Atlantic Yes Chesapeake Bay Gulf of Mexico - west Florida shelf Maybe There is an ongoing debate, but at present insufficient data to resolve the issue - Texas coast Yes California No Conjecture in relation to stimulation of populations once they are transported inshore. Pacific northwest No Heterosigma akashiwo Yes Alexandrium catenella (Puget Sound but not along the open coast)

- 108 - 3.5 Nutrient Ratios, Dissolved Organic and Particulate Nutrients

3.5.1 Introduction

In the preceding sections the focus has largely been on the enrichment of coastal waters with dissolved inorganic nutrients and how this might lead to an increase in the frequency and amplitude of blooms. In this section, consideration is given to if, how, and the evidence that changes in nutrient ratios (N:P [:Si]) select for HAB species (generally or by species/ life-form), or for increased toxicity of toxin producing species. In addition, consideration is given to whether and how the availability of organic forms of nutrients, dissolved and particulate organic nitrogen (DON and PON) and particulate organic phosphorus (POP) and perhaps dissolved and particulate organic carbon (DOC and POC) might favour certain HAB species.

3.5.2 Nutrient ratios

3.5.2.1 Theoretical considerations

Tilman et al. (1982) suggested three broad explanations for the spatial and temporal variation in phytoplankton species composition. Two of these, the physical environment (the capacity of species or lifeforms to grow in environments that differ particularly in terms of vertical mixing) and top down control (variable loss through selective grazing or immunity from grazing) are not considered here. The third, which involves the relationship between the ratios of nutrients required for growth and ambient nutrient ratios, is considered here although it should be noted that no explanation is exclusive, and no explanation precludes consequential effects of floristic changes on the consumers of phytoplankton. Redfield (Redfield, 1934; Redfield, 1958; Redfield et al. 1963) observed that chemical composition of plankton tended towards a ratio of C:N:P atoms of 106:16:1. Redfield saw that since it was mainly the mineralization of plankton-derived organic matter that re-supplied the ambient pools of inorganic N and P, then the molar ratio of nitrate to phosphate should also tend to 16:1. That is, the composition of plankton and these aspects of the composition of seawater were part of a cycle, each determining the other. Modern understanding includes consideration of other elements, such as silicon. Copin-Montegut and Copin-Montegut (1983) found that particulate material from the oceans also had a similar ratio although Geider and La Roche (2002) report variations in particulate N:P ratio with a range from 5 to 34. The microplankton often appear to have ratios that approximate Redfield, but in fact planktonic photoautotrophs display a wide range of cellular composition. For example, in one study the flagellate Pavlova lutheri was grown with C:N:P ratios ranging from 682:66:1 to 88:14:1 (Tett et al. 1985). Similar variability has been shown by euglenoids, dinoflagellates, - 109 - chlorophyceans, cryptomonads, diatoms, pelagophytes, haptophytes, and cyanobacteria (Table 3.2 of Tett et al. 2003b). The relationship between nutrients and the growth of populations of micro-organisms can be described by theories of varying complexity. The simplest ‘Monod’, the intermediate ‘Cell quota’ and complex mechanistic explanations. In the case of the latter, models incorporate realistic accounts of the main biochemical processes and pools within cells (e.g. Flynn & Hipkin 1999; Flynn 2001, 2005). While such mechanistic models include many parameters, there is at present, insufficient information to distinguish between groups or species of phytoplankters. The simplest theory is that developed for organic-carbon-limited bacterial growth by Monod (1942) and applied to phytoplankton by Dugdale (1967). In this, the rate of uptake of dissolved nutrient (per unit biomass) depends on ambient concentration but it does not account for cellular storage of nutrients. The Monod model is therefore too simple to give a good description of laboratory growth of single species populations, but may well be a reasonable approximation for populations in the sea (Davidson 1996). According to ‘cell-quota’ theory (Droop 1968, 1983) algal growth rates are controlled by cellular concentrations (cell quotas) of nutrients. The cellular content Q of nutrient (atomic nutrient element (atom organic carbon (C)) -1), can vary between limits defined by the minimum or subsistence quota (kQ) and the maximum cell quota (Qmax). The quota allows for cellular storage of a nutrient and so buffers against the effects of ambient change. According to Klausmeier et al. (2004) cellular storage of nutrients is one reason for the variation in stoichiometry (ratios of nutrient elements) in phytoplankton but this is additional to changes in structural components (e.g. nucleic acids, proteins and pigments) for which there is a smaller range of N:P (7.1 to 43.3) compared to the overall range.

The ratio of the maximum cell quota: subsistence quota (Qmax/kQ) is lower for N and Si (2- 4) compared to P (5 - 90) and is part of the reason why marine phytoplankton tend to be N rather than P limited. Tett et al. (2003b) give a nitrogen subsistence quota for typical (eukaryote) photo- autotrophs with carotenoid accessory pigments of 0.05 at N: at C (but suggest that large oceanic dinoflagellates have lower nitrogen subsistence quotas (0.02 at N: at C); typical phosphorus subsistence quotas are in the range 0.001 – 0.002 at P: at C. The ratio kQ1/kQ2 determines relative limitation by nutrients 1 and 2. Thus, if the ratio of ambient nutrients 1 and 2 is < kQ1/kQ2 then nutrient 1 may be limiting. Diatoms require silicate for cell wall formation but diatom species differ in their wall thickness, typical silica contents vary widely about a median of 0.11 atoms of silicon per atom of carbon (Brzezinski, 1985). Tett et al. (2003b) give a silicate subsistence quota for species of Thalassiosira and Chaetoceros and Skeletonema costatum of 0.03 at. Si: at. C and (Tett &

- 110 - Droop, 1988) suggest a mean subsistence quota of 0.05 at of Si per at of C. Brzezinski (1985) noted that, although variation exists, many marine diatoms have a relatively balanced N:Si ratio N Si within their biomass. However, if the ratio kQ: kQ is  2, (Tett et al. 2003b) then typical pelagic diatoms require twice as much N as silicate. The silicate requirement of silicoflagellates may be as high as evidenced by the colonial freshwater flagellate Synura petersenii, which has silicified scales (Sandgren et al. 1996) although quantitative studies of marine silicoflagellates have not been reported.

Pan et al. (1996) give ratios of maximum to minimum cellular silicon (Qmax to Qmin) for a range of diatom species. For Pseudo-nitzschia multiseries the ratio is 15.3 compared to 1.1 for Cerataulina pelagica and 8.8 for Coscinodiscus granii. This suggests P. multiseries is able to respond to a wide range of silicate levels and provides one explanation for its ubiquitous distribution although Pan et al. (1996) urge some caution as the highest value of Qmax 214.4 pg Si cell-1 may have been the result of luxury uptake. Note also that the experiments were run with high initial Si concentrations (60.9 – 190.5 µM) that are not representative of typical natural concentrations. The need for evidence to test nutrient ratio hypotheses arising from the above theories presents a number of practical issues of how to measure nutrient element ratios. The ratios of maximum winter dissolved inorganic nutrients may help to assess the general potential for shifts in the balance of organisms in water bodies where there is little summer input of nutrient. This seems less relevant to summer HAB species and in any case is restricted to high latitude waters with an obvious 'winter' or low-growth period (although seasonal monsoons might also result in a low growth period. Summer ratios may be more relevant to HABs, but are more difficult to measure accurately and would be open to the objection that most nutrients are inside the cells at this time and nutrient regeneration and rapid recycling may occur. Finally, the use of ratios of total N and total P (in the appropriate season) as measures of nutrient of particulate form within algal cells is sometimes attempted, but such bulk measurements may be influenced by particulate material from other plankton and seston. Nutrient ratios of input fluxes from human discharges and land runoff have been used to gauge the potential effect of anthropogenic enrichment of coastal waters. Such loadings do not generally take into account the cycling of nutrients in estuaries including: denitrification; the equilibrium dynamics between dissolved available phosphate, bound phosphate and organic phosphorus; the dissolution of Si; the microbial regeneration and rapid utilisation of NH4. That is, nutrient ratios based on riverine loadings may not reflect the ratio of nutrients available to phytoplankton in coastal waters. Other natural inputs (upwelling, diapycnal mixing, and benthic flux) are often hard to measure, so little reliable data exists.

- 111 - Finally, it is possible to measure nutrient uptake ratios from uptake rate measurements or enrichment experiments using bulk chemistry or isotopic measurements although these are not routinely undertaken. It may also be possible to use specific indicators of cellular status/processes/stress - e.g. measurements of particular enzymes - or of cellular toxin content (but this becomes tautological if we seek evidence of nutrient-enrichment link to HABs).

3.5.2.2 Nitrogen to phosphorus ratio

Arguments about perturbations of nutrient ratios in marine waters often begin with the Redfield ratio. The nitrogen to phosphorus atomic ratio of 16:1 (Redfield ratio) is widely used with respect to ambient concentrations of dissolved inorganic N and P to infer which nutrient is likely to be limiting. A ratio < 16:1 is taken to indicate N limitation and a ratio > 16:1 P limitation. However, as acknowledged by Redfield himself the ratio is a general basin wide and seasonal average. Based on nutrient concentrations given in Gowen et al. (2002) the N:P ratios in pristine near surface oceanic waters (salinity  35) in the shelf break region of the Celtic Sea were between 14.5:1 and 17.6: 1, and as discussed above, because of luxury uptake, assimilatory ratios can differ from Redfield. Recently, the assumption that the Redfield ratio can be used to differentiate N and P limiting conditions has been questioned. Maestrini et al. (2000) reviewed values of the cellular N:P ratio at which algae pass from nitrogen to phosphorus limitation. Documented values for five marine and freshwater phytoplanktonic algae range from 14 to 45, with a median at 28. The results from the experiments of Maestrini et al. (2000) suggest a shift to phosphorus limitation at an external N:P ratio of 40:1. More recently, Geider and La Roche (2002) suggested that the critical N:P ratio which marks the transition from N to P limitation is between 20 and 50 mol N : mol P but that based on a typical biochemical composition, the critical N:P ratio for nutrient replete phytoplankters is between 15 and 30. In the Baltic Sea, enrichment with P is generally accepted to have stimulated blooms of cyanobacteria (Larrson et al. 1985; Nehring 1992) because some of these blue-green bacteria are able to fix dissolved nitrogen gas. Changes in N:P ratios have been used as evidence for shifts in floristic composition and HABs in Dutch coastal waters (Riegman 1995, and reviewed by Anderson et al. 2002) and Tolo Harbour in Hong Kong (Hodgkiss & Ho 1997). Here we consider some of the complexities and difficulties in relating shifts in nutrient ratios to increased occurrence, duration and size of algal blooms. Using mesocosms, Riegman et al. (1992) observed that Phaeocystis sp. was a poor competitor for P (Phaeocystis sp. was out competed by Emiliania huxleyi and Chaetoceros socialis) but a good competitor for N (it became the dominant species in N limited conditions).

- 112 - The implication of this is that species of Phaeocystis would be more successful and possibly out compete other species under conditions of N control and that a reduction in N:P ratio would favour Phaeocystis spp. However, Weisse et al. (1986) concluded (on the basis of field observations in the German Wadden Sea of Sylt) that, Phaeocystis sp. had lower inorganic nutrient demands, especially for phosphate compared to diatoms and could grow when the phosphate concentration was 0.2 µM. Moreover, mesocosm experiments by Brussaard et al. (2005) using coastal North Sea water did not display a marked difference in the percentage contribution that Phaeocystis spp. made to the planktonic community under N or P limited conditions. Riegman (1995) used the Phaeocystis sp. N:P ratio hypothesis to argue that in Dutch coastal waters, the shift from P to N limiting conditions (total N: to total P ratios decreased from 38 to 13 during the late 1970s and 1980s in the Marsdiep region of the Dutch Wadden Sea) was the reason that

‘Novel summer blooms of Phaeocystis appeared in the late seventies’.

There are three points which need to be considered in relation to the Phaeocystis spp. - N:P ratio hypothesis. The first is whether as suggested by Riegman (1995) summer blooms first appeared in the late seventies. There is some doubt over this. According to the data in Cadée and Hegeman (1986) summer peaks in Phaeocystis spp. abundance occurred in the early 1970s (1974 and 1975 (see also Savage 1931). This point is also noted by Peperzak (1993):

‘However, summer blooms had already occurred before 1978….’

Although the legend to Figure 1 of Riegman (1995) refers to an increase in summer blooms of Phaeocystis spp. it is clear from the figure that it is an increase in the duration (in days) of the blooms which is related to the N:P ratio. The second point is that Riegman (1995) used ratios of total N and P. He argued that it was not possible to identify the nature of the controlling factor on the basis of nutrient concentrations and ratios but that ratios of total molar N:P might give some indication of the potential controlling factor when light is not limiting. The reason why the former cannot be used but the latter can is not made clear by Riegman (1995) and as discussed above, there are some interpretational difficulties with both. Assuming limitation of a particular inorganic nutrient based on ratios of ambient concentrations presupposes that nutrients will be taken up in the Redfield ratio, and that the Redfield ratio is the critical ratio that denotes the transition from N to P limitation: neither are necessarily the case. Furthermore, total N and P, measurements are likely to include detrital N and P which will distort the algal N:P ratio.

- 113 - This idea of a change in limiting nutrient and shift in floristic composition is also highlighted in the review of Anderson et al (2002). However, as discussed by Philippart et al. (2007) the relationship between nutrients and phytoplankton community composition in the Dutch Wadden Sea is not clear cut. While these authors find some significant relationships between the concentrations and ratios of inorganic nitrogen, phosphorous and silicon to community structure, these relationships were often quite weak. Philippart et al. (2007) therefore concluded that: “the precise responses of biomass and production to changes in nutrient loads is largely unpredictable”

The final point to consider in relation to the Phaeocystis spp. - N:P ratio hypothesis is whether nutrients are in fact controlling summer growth in Dutch coastal waters. In such conditions, ratios may be important in determining floristic composition but this is not the case when nutrients are present in excess. For the inner regions of the Wadden Sea, Postma and Rommets (1970) stated that:

“Even in spring and summer, nutrients are rarely a limiting factor for plankton growth.”

and more recently, Colijn and Cadée (2003) were of the opinion that little attention has been paid to the role of light in controlling phytoplankton production in the Wadden Sea and concluded that:

“During the 1990s the dominant influence of high DIN concentrations implies that underwater irradiance by far exceeds effects of nutrients on the production of phytoplankton biomass.”

For the Marsdiep region in particular, Colijn and Cadée (2003) were of the opinion that slight nutrient limitation only occurred between May and July. For the period 1980 to 1987, Escaravage et al. (1995) give mean June/July concentrations of DIN of 62 µM and 1.16 and 6.6 µM DIP and Si respectively for a station 10 km off the Dutch coast. It seems unlikely that such high concentrations are limiting for phytoplankton growth. These levels are for example, higher than winter concentrations in offshore western Irish Sea waters, where mean March (1993 – 1999) concentrations were 9.5 µM DAIN, 0.9 µM DIP and 7.1 µM Si (Gowen et al. 2002) and are clearly sufficient to support a spring bloom of up to 16 mg chlorophyll m-3 (Gowen & Stewart 2005). In summary, while evidence exists for the role of nutrients in partly controlling the phytoplankton community in Dutch Coastal waters, other factors may be important or dominant. The link between composition and N:P ratios remains tenuous although Lancelot et al. (2009)

- 114 - concluded that for ecosystems such as Belgian coastal waters, a loadings or winter concentration N:P ratio of > 25 was indicative of Phaeocystis spp. colony dominance. Hodgkiss and Ho (1997) analysed data from Tolo Harbour (Hong Kong) collected during the 1980s and showed that a decrease in the annual mean N:P ratio (from ≈ 20:1 in 1982 (1983 is also given in the text) to ≈ 11:1 in 1989) coincided with an increase in the frequency of red tides (Figure 3.20A). Hodgkiss and Ho (1997), cite the study of Ho and Hodgkiss (1995) as supporting their argument that low N:P ratios favoured the growth of Prorocentrum micans, P. sigmoides and P. triestinum. One of the problems interpreting the data presented by Hodgkiss and Ho (1997) is that the data set only covers 8 years. Extending the time series using N:P ratios from the three inner stations in Tolo Harbour, shows that the relationship between the ratio of N:P and HAB occurrence is unclear (Figure 3.20B). During the early 1990s the N:P ratio was less than Redfield but the frequency of red tides was low.

Figure 3.20 Temporal changes in the mean molar N:P ratio and the occurrence of red tides in Tolo Harbour. A, for the period 1982 to 1989 (redrawn from Figure 2 of Hodgkiss and Ho (1997); B, for the period 1982 to 2005 (data from Figure 2 of Hodgkiss and Ho (1997) and the Hong Kong Environmental Protection Department).

40 25 A

r 35 20 30

25 Red tide incidents 15 20 N:P

15 10

ratio molar N:P 10 5

tidesof red yea per Number 5

0 0 1982 1983 1984 1985 1986 1987 1988 1989

40 70

r B 35 60 N:P 30 N:P 50 25 40 20 30 15 N:P molar ratio molar N:P 20 10

Number of red tidesNumber per yea 10 5 0 0

1982 1983 1984 1985 1986 1987 1988 1989 1990 1991 1992 1993 1994 1995 1996 1997 1998 1999 2000 2001 2002 2003 2004 2005

- 115 - 3.5.2.3 Silicate limitation of diatom growth

Officer & Ryther (1980) argued that the Si:N ratio determines the dominant type of phytoplankton. Thus, in situations in which silicate becomes limiting for diatom growth, they are replaced by other lifeforms (dinoflagellates or microflagellates). This ‘silicate limitation’ hypothesis has led to concerns that while rivers historically carried dissolved Si well in excess of dissolved N and P, many rivers are showing signs of a stoichiometric nutrient balance of Si:N:P = 16:16:1, or even Si deficiency (Justic et al. 1995). The concern over the decrease in riverine dissolved silica (or increase in N and P relative to silicate) is that this will lead to an excess of N and P (relative to Si) which is available to flagellates for growth once diatom growth is limited by the depletion of Si. Anthropogenic enrichment of the German Bight of the North Sea has perturbed winter N:Si ratios. Radach et al. (1990) report an increase in the molar N:Si ratio from 1-2 in the late 1960s to 4-8 in the early 1980s. Analysis of the Helgoland time-series led Hickel (1998) to conclude that there was evidence of nutrient enrichment but the expected long-term trends in phytoplankton were not always clearly represented. Recurrent 3 - 5 year cycles of diatom and flagellate biomass were apparent in the data and by separating the phytoplankton data into nanoplankton (< 20 µm) and microplankton (> 20 µm), Hickel (1998) found that the three fold increase in total phytoplankton was largely due to an increase in the winter biomass of nanoplankton. Since light limits phytoplankton growth during the winter in the German Bight, Hickel (1998) was of the opinion that: the flagellates were mostly heterotrophic and mixotrophic species < 5µm in size; their increase was not significantly correlated with inorganic nitrogen but some other compound; the explanation for the increase in flagellates was not known but coincided with other large scale effects. Hickel (1998) concluded that once the nanoplankton component was separated from the autotrophic microplankton:

“It became apparent that neither diatom biomass, nor dinoflagellate biomass without the nanoplankton component showed a clear long-term upward trend, possibly due to the enormous inter-annual variations which might have masked minor trends.”

The silicate limitation hypothesis was considered to be the explanation for the increase in size and duration of spring Phaeocystis spp. blooms along the continental coast of the southern North Sea (Cadée & Hegeman 1986; Lancelot et al. 1987; Lancelot 1990). However, this does not explain a number of observations. For example, during 1988-89, Phaeocystis spp. made up a larger share of the phytoplankton in East Anglian waters than in the German Bight, although N:Si ratios were higher in the latter (Tett et al. 1993; Tett & Walne 1995).

- 116 - A difficulty with conclusions based on observations is that correlation of changes does not prove causation. More rigor is provided by mesocosm experiments. Experiments with deep bags in the eutrophic Seton Akai of Japan showed that dissolved silica depletion led to shifts in dominant species from larger to smaller diatoms or flagellates (Harada et al. 1996). In reviewing results from a number of northern European mesocosm experiments, Tett et al. (2003b) concluded that it was not easy to see an overall pattern. Egge and Aksnes (1992) enriched floating enclosures with nutrients including and excluding silicate and compared these with un- enriched enclosures. Diatoms were found to dominate when the silicate concentration was > 2 µM and Phaeocystis spp. appeared after the bloom of other species but not when silicate concentrations were high. Similarly, Williams & Egge (1998) stimulated diatom growth by the addition of silicate and Gilpin et al. (2004) found that flagellates began to dominate over diatoms in mesocosms in a Norwegian fjord once the N:Si supply ratio exceeded two. These results are generally supportive of the silicate limitation hypothesis and the argument presented above that diatoms typically require N:Si in the ratio of 2:1. It is also possible that rather than the N:Si ratio being the important factor, it is the absolute concentration of Si (perhaps  2 µM,although this is likely to be species specific, Davidson et al. 2007) that controls diatoms growth. One explanation for this might be that Si is required for cell wall formation and an inability to form cell wall material directly influences population growth by preventing cell division. In batch culture experiments to investigate the effect of silicate limitation on domoic acid production by Pseudo- nitzschia multiseries, Pan et al. (1996) observed a cessation of frustule formation at a Si concentration of 3.2 µM (these authors also stated that Si was required for DNA synthesis). There are counter arguments. Escaravage et al. (1995) found that in mesocosms in which the light and nutrient regimes were manipulated to resemble conditions in Dutch coastal waters, Phaeocystis spp. out-competed diatoms in nutrient replete conditions and large blooms of Phaeocystis developed in the mesocosms. In Phaeocystis spp. culture experiments, Peperzak (1993) examined the influence of daily irradiance on growth rate and colony formation and concluded that below a threshold of 100 W h m-2 d-1 cells of Phaeocystis spp. were small and there was no colony formation. Above this threshold increased cell size and colony formation was observed. In a further examination of this light threshold hypothesis, Pepezak et al. (1998) examined data collected from Dutch coastal waters in 1992 and concluded that the timing of Phaeocystis spp. blooms was not related to silicate limitation but a daily light threshold. In conclusion, it is clear that there is a marked difference between the precise control of algal growth in culture by nutrients and the situation in the sea, where in addition to physiological (cellular storage) and ecological (multiple species allowing adaption within a life- form) buffers there are other important factors (hydrodynamics, grazing) regulating the growth

- 117 - rate of particular species. In mesocosms, large shifts in nutrient ratios are needed to bring about (predictable) changes in the balance of organisms. Thus one might expect that nutrient-ratio- driven shifts in the balance, or a move to more HABs, would only occur in semi-enclosed, near shore waters where nutrient ratios are substantially perturbed. Conversely, as data from the Scottish west coast Loch Creran demonstrates, marked changes can take place in the balance of organisms with little change in nutrient loads and ratios (Tett et al. 2008).

3.5.3 Dissolved organic and particulate nutrients

There are significant pools of dissolved organic nitrogen in coastal waters (see review by Bronk 2002) and that during the summer, when concentrations of inorganic nutrients are low, the largest pools of fixed N and P are generally the dissolved organic pools. Sanders et al. (2001) showed that from the inner Thames estuary (UK) to more offshore waters of the southern North Sea, nitrate decreased and dissolved organic nitrogen (DON) assumed greater significance. It is generally accepted that phytoplankton can take up and utilise a range of nitrogenous organic compounds from (see for example, Stolte et al. 2002) but the quantitative importance of these pools of organic nutrients in phytoplankton nutrition is uncertain (Caron et al. 2000). That high levels of DON are measured in surface waters during the summer when phytoplankton biomass is low, suggests that much of the DON pool is relatively poorly exploited for phytoplankton nutrition. This may be because under natural conditions, much of this material is made up of large, refractory, molecules that are largely unavailable to primary producers. Rapid utilisation of labile compounds, perhaps mediated by bacterial uptake and remineralisation (Davidson et al. 2007) may make DON an important source of nutrition.

The nitrogenous compound urea [CO(NH2)2] can be assimilated by many species of micro- algae and also by cyanobacteria, which use urea as a nitrogen source in preference to nitrate. Presumably bacteria can also utilize urea as a nitrogen and energy source, releasing ammonia when energy limited. Mixotrophic protists may also be able to access urea-N by way of bacteria or micro flagellates. Glibert et al. (2006) have recently argued that: urea currently represents > 50 % of nitrogenous fertilizer usage worldwide; unhydrolized urea can be lost to surface runoff; urea concentrations in receiving estuaries and coastal waters can be significantly enhanced by land-based inputs; urea can be a significant fraction of the total DON pool in some coastal waters; urea may represent an important N source for some HAB species. Furthermore Glibert et al. (2006) suggested that:

“….many regions of the world where both total nitrogen use has increased, and where the urea dominates the agricultural applications of nitrogen, are also regions that have experienced increasing frequency and extent of harmful algal

- 118 - blooms (HABs).”

As pointed out by Glibert et al. (2006) however, whether or not urea represents a significant contribution to the nutritional requirements of some HAB species remains open:

“urea may contribute disproportionately to the nitrogen nutrition of some harmful and nuisance phytoplankton groups.”

The evidence of a link between urea and HABs presented by Glibert et al. (2006) is based on a comparison of global maps of urea usage in the 1960s and in 1999 and the occurrence of dinoflagellate species causing PSP or documented cases of PSP. As discussed earlier in this section these maps are difficult to interrogate and while the global maps are suggestive of an increase in PSP species/ incidents in northern Europe, this may be a function of under representation of historical occurrences and increased monitoring effort in western Europe since the 1990s. Given the possibility of a link between urea and enrichment and HABs it is clearly undesirable to conduct deliberate large-scale urea enrichment of the seas in order to sequester carbon, as Glibert et al. (2008) have argued. However, the laboratory data are not clear-cut about HAB species urea preference and in any case natural ecosystems are more complex than single species cultures. Thus, there is a need for small scale experiments in mesocosms to determine whether nitrogen supplied as urea rather than as nitrate or ammonium does lead to more HAB species biomass. Until such experiments have given positive results, the urea - HAB link should be viewed as hypothetical rather than established. Glibert et al. (2004) related urea to the occurrence of Pfiesteria spp. in Chesapeake Bay and Glibert et al. (2007) were of the opinion that the Aureococcus anophagefferens brown tides in coastal bays of Maryland were related to changes in DON. Laroche et al. (1997) were also of the opinion that A. anophagefferens utilized DON as a nitrogen source and Cosper et al. (1990) state that A. anophagefferens can grow on urea as a sole nitrogen source but also has a growth requirement for trace-elements and chelating compounds. Using mesocosm experiments, however, Keller and Rice (1989) found that A. anophagefferens appeared to exhibit an ability to grow at levels of DIN considered to be limiting for diatoms. In culture experiments with the red tide dinoflagellate Heterocapsa circularisquama, Yamaguchi et al. (2001) found that this phytoplankter grew well on inorganic nitrate, nitrate and ammonium but urea and uric acid were not utilized. This species was also able to utilize a wide variety of inorganic and organic compounds of phosphorus as the sole P source. Using cultures of Alexandrium tamarense and Gymnodinium catenatum isolated from Hiroshima Bay (Japan),

- 119 - Oh et al. (2002) found that both phytoplankters were able to use dissolved inorganic and dissolved organic phosphorus and concluded that:

“the DIP-depleted conditions in Hiroshima Bay might have led to the outbreaks of noxious dinoflagellates in recent years.”

The transport of terrestrial-derived, riverine dissolved organic material (DOM) through estuarine systems and into the coastal zone has been reported (Mantoura & Woodward 1983; Seitzinger & Sanders, 1997; Minor et al. 2001). Glibert et al. (2005) also highlighted the importance of terrestrially derived DOM but stated that the chemical composition of DOM from agricultural watersheds is unknown. As noted by Bronk (2002), quantifying the role of DOM in the process of coastal eutrophication is a key challenge. Nishimura (1982) undertook culture experiments using Gymnodinium type-65 and Chattonella antiqua and found that Gymnodinium grew well in cultures using water from the vicinity of fish farms and when extracts of mackerel meat and yellowtail faeces were added to the cultures in low amounts. C. antiqua did not grow in cultures with organic material added. Rather few studies have directly considered the role of particulate organic matter (POM) as a substrate for micro-algal growth. Nevertheless, given that much of the recycled nutrient pool is derived from the remineralisation of algal biomass, the POM pool must represent a source of utilisable organic matter. Isotope based studies indicate that terrestrial POM persists in the coastal zone (Fichez et al. 1993), although this would suggest that it is relatively inaccessible to phytoplankton. Particulate organic matter may be important for mixotrophic and heterotrophic dinoflagellates.

3.5.4 Nutrients and toxin production

Bates et al. (1993) observed that Pseudo-nitzschia pungens required a high external supply of inorganic nitrate to produce domoic acid (DA), consistent with the fact that DA is an amino acid and nitrogen is required for its synthesis. Subsequent studies have found little evidence of DA in balanced growth, but that it is produced under nutrient stress. Pan et al. (1996) studied the effects of silicate limitation on the production of domoic acid in batch cultures of Pseudo-nitzschia multiseries. Domoic acid was produced when population growth was declining and was at a maximum when cells were silicate depleted. These workers suggest their batch culture data were consistent with field observations made during the first domoic acid incident in Prince Edward Island, Nova Scotia (Canada) when peak domoic acid production occurred 10 days after the peak of the bloom and when Si in the water was depleted. These results are supported by Fehling et al. (2004) who found greater toxicity in Pseudo-nitzschia seriata (from Scottish waters) under Si

- 120 - limitation compared to P limitation. In contrast, Marchetti et al. (2004) observed the presence of DA in healthy, growing phytoplankton communities and suggested that there was a need to examine how environmental factors may influence DA production in natural populations of Pseudo-nitzschia species. More recent studies by Fehling et al. (2005) who identified a photoperiod effect on growth and toxicity and Wells et al. (2005) who suggest a linkage between domoic acid, iron and copper are also suggestive of a complex suite of factors influencing DA production. The role of nutrients in promoting PSP toxin production may be species specific. PSP toxins are nitrogenous compounds and hence N is required for their synthesis. This suggests that N stress will be detrimental to PSP toxin synthesis (Flynn & Flynn 1995). Studies have also linked P-stress to increased PSP toxicity (Boyer et al. 1987; Anderson et al. 1990; John & Flynn 2002). Murata et al (2006) report that Alexandrium tamarense becomes more toxic at higher N:P ratios because toxin content is proportional to cellular protein content. Similarly, Granéli et al. (1998) suggested that, with respect to species and strain, toxin production was probably under genetic control but that toxin content per cell was influenced by a variety of abiotic (temperature, light, nutrient concentration) and biotic (competitors, grazers) factors. These workers found that, compared to N deficiency, P deficiency increased toxin level 3 fold in Alexandrium tamarense and was presumed to be a mechanism for storing excess N. Granéli et al. (1998) also reported that for the PSP producing species Gymnodinium catenatum, P deficiency also results in an increase in toxin content per cell. In contrast to the above observations, Flynn and Flynn (1995) found rates of PSP toxin production in Alexandrium minutum to decrease under P-stress, suggesting that for this species, P may be involved in the regulation of toxin synthesis. Furthermore, Flynn and Flynn (1995) suggest a complex relationship with cell-N may exist. When N becomes exhausted, toxin synthesis continues for a few days but then falls to very low levels. They concluded that interpretation of toxin and cell-nutrient relationships in Alexandrium species is complicated. Stolte et al. (2002) showed that A. tamarense is able to utilize organic matter and is consistent with the observations of Granéli et al. (1998) that under N limiting conditions, toxin production decreased but increased again when yeast was added to the culture. Granéli et al. (1998) suggests very little is know about the role of DOM and POM and toxin production, a statement that remains true a decade later. With respect to DSP toxins, both N and P limitation has been shown to produce similar levels of toxicity in Prorocentrum lima. For species of Dinophysis, Granéli et al. (1998) stated that highest toxin content in cells occurred under N limitation. On the basis of results from semi- continuous cultures of Chrysochromulina polylepis, Johansson and Granéli (1999) concluded

- 121 - that the toxicity of C. polylepis was strongly influenced by the physiological state of cells and that this provided one explanation for the large variability in the toxicity of this species. Culture experiments show that high N:P ratios in the medium resulted in increased toxin levels in C. polylepis. In summary, there is an increasing amount of work in algal culture that shows harmful algae becoming more toxic when cells are 'nutrient stressed' i.e when growth slows because one nutrient becomes limiting and nutrient supply ratios are markedly different from Redfield. The general explanation seems to be that toxin is synthesized while biomass synthesis slows. Such findings might imply that blooms are likely to become more toxic towards their end, but do not help to explain any widespread increase in HABs or toxicity - these might be expected to result from changes in nutrient ratios only in semi-enclosed, near-shore, highly loaded, waters.

3.6 Hypotheses Concerning the Occurrence of HABs

One of the arguments for seeking a single general hypothesis for the putative increase in HABs is the apparent global synchronicity of the increase which took place over a period of two to three decades (1960s to the 1980s). According to Smayda (2008) this is suggestive of a disruption of the plankton habitat on a global scale which is driven by anthropogenic activity and for which there are two hypotheses: global climate change and global eutrophication although these two hypotheses are not mutually exclusive. There is evidence for climate driven changes in HAB species abundance at regional and local scales but global scale climate forcing such as that suggested by Hayes et al. (2001) remain hypothetical. With respect to the nutrient enrichment → HAB hypothesis, during a U.S. Environmental Protection Agency workshop in 2003 (see Heisler et al. 2008) a group of US expert scientists concluded that:

“Degraded water quality from increased nutrient pollution promotes the development and persistence of many HABs, and is one of the reasons for their expansion in the U.S. and the world”

In our opinion, HABs may result from a number of natural and anthropogenic causes and there is a need to dissect out the cause-and-effect chain leading from anthropogenic nutrient enrichment to HABs, or to eliminate other causes, in order to provide convincing evidence of the nutrient-HAB link. We therefore hypothesise that:

there is no single general hypothesis for changes in the occurrence of HABS; instead, we must look at interactions between changes in specific pressures,

- 122 - the ecohydrodynamic conditions in particular water bodies and the adaptations of particular HAB species or life-forms.

In Part 4 we investigate the relationships between HABs, and anthropogenic nutrient enrichment in UK and Irish waters to further examine the nutrient enrichment → HAB hypothesis.

- 123 - PART 4

An Evaluation of the Current Distribution of HAB Species in UK and Irish Coastal Waters

4.1 Introduction

The second objective of this study was to investigate the relationship between anthropogenic nutrient enrichment HABs and HAB species in UK and Irish coastal waters. Our starting point was the hypothesis that:

the occurrence of HABs and HAB species abundance increases with anthropogenic nutrient enrichment (proxy: riverine loading and mean winter concentrations of nutrients)

To test this hypothesis, data sets on nutrients and the growing season (April – September) abundance of HAB species were compiled and analysed statistically. In addition, time series analysis was carried out on PSP toxicity data from coastal waters of north east England and selected phytoplankton data from coastal waters of Northern Ireland. The remainder of this section sets out the methods used to compile the data and the analyses performed. The results are presented in tables and as maps showing the geographical distribution of species abundance. In discussing the results we first consider the adequacy of the data and the reliability of the statistical and interpretational analyses used and highlight key findings. A detailed discussion of the results in the context of the findings from the literature review and in relation to the ecophysiology of particular species and the ecohydrodynamic conditions of UK and Irish coastal waters in which these species live, is presented in Part 5.

4.2 Methods

4.2.1 Nutrient data

4.2.1.1 Riverine loadings

The relationship between riverine loadings and HAB species abundance was investigated using UK data only. These data referred to as RID (Riverine Inputs and Direct Discharges) are presented as annual loads per catchment and coastal sea area (Figure 4.1), based on monthly measurements. The data include riverine, domestic and industrial sources of nutrients. Reports on the data are available from 1992 onwards (see for example OSPAR 2001). No loadings data

- 124 - were available for the Orkney and Shetland Islands and therefore the Northern Isles were excluded from this part of the analysis. For the UK annual reports there is a single large sea area Sc2 that covers the Firth of Clyde, Sound of Jura and Firth of Lorne. To better reflect the higher loadings to the Firth of Clyde sea area, we separated this area as Sc1a (Figure 4.1). Loadings to this area are those given for area Sc2 in the annual reports. For loadings to our coastal sea area Sc1b, the loadings to area Sc2a were used on the grounds that both areas have similar land use and population density. The nutrients and ratios derived from the annual reports and used in the analysis are listed in Table 4.1.

Figure 4.1 A map of the UK showing the coastline divided into regions based on riverine catchment. (Redrawn from the UK annual report to OSPAR, 2002).

62

Shetland (Area Sc2d) 60

Orkney (Area Sc2c) Sc2b

58 Sc2a SC3 Scotland Sc4 56 Sc1b Sc5 E1 NI1 Sc1a E2a E3 E2 E5 Northern Sc1 E4

Latitude Ireland E30 E6 NI2 54 E29 E7 E28 E7A E27 E8 E10 England E9 Wales E26 52 E22 E11 E25 E24 E23 E12 E20 E21 E13 Celtic E14 E19 E15 Sea E17 E16 50 E18 English Channel

France 48 10864202

Longitude

- 125 -

Table 4.1 Details of the nutrients and ratios used to test the relationship between nutrient enrichment and HAB species abundance.

Nutrients and ratios UK loadings data Winter concentrations Measured Modelled UK data Irish data Ammonium (NH4)    Nitrate (NO3)     Nitrite (NO2)   

Total oxidisable nitrogen (TOxN as NO3 + NO2)   

DIN (NH4 + NO3 + NO2)   DIP (Phosphate as PO4)     Silicate (Si)    Total N  Total P  Nutrient ratios (NH4+ NO3) : DIP  Total N: Total P  DIN: DIP   DIN:Si   TOxN: PO4   TOxN:Si  

For the UK, unmonitored areas account for ~39% of the landmass and the issue of missing data needs to be addressed to avoid under-estimation of the nutrient loads entering the marine environment. The original monthly data were therefore used to derive modelled loadings (over the same time period). The derivation process involved interpolation to cover missing data points, use of climatology (a standard year based on the available data) to fill in gaps in the data and correction factors to account for un-gauged areas. The modelled loadings data were provided by Cefas. Recent assessments (such as the 2005 OSPAR assessment of the eutrophication status of UK coastal waters, OSPAR 2008) indicate that riverine loadings and concentrations of nutrients are relatively constant (accepting that there is natural inter-annual variability). That is, there seemed to be no trend (1999 to 2005) and it was considered legitimate to extract a multi-year mean from the data showing considerable inter-annual variability to improve the precision of estimate of nutrients.

4.2.1.2 Winter nutrient concentrations

Mean values of maximum winter (January and February) concentrations of dissolved inorganic nutrients collected from UK and Irish coastal waters (Figure 4.2) between 2000 and 2007 were used in the analysis by matching particular years with the phytoplankton species data (see below). The UK data (Table 4.1) were aggregated into the coastal sea areas shown in Figure 4.1 depending on station latitude and longitude. Data from Irish coastal waters (Table 4.1) were

- 126 - aggregated into coastal areas corresponding to the location of phytoplankton sampling stations (Figure 4.2B).

Figure 4.2 The location of winter nutrient and phytoplankton sampling stations UK and Irish coastal waters. A, UK waters; B, Irish coastal waters.

A B 61

59

57

55 Latitude

53

51

49 1198653102 Longitude

4.2.2 Phytoplankton data

Data on the growing season (April to September) abundance of Alexandrium spp., Dinophysis spp., Pseudo-nitzschia spp., Karenia mikimotoi, Prorocentrum lima, P. minimum, Lingulodinium polyedrum and Protoceratium reticulatum were compiled using data collected from UK and Irish coastal waters. In the UK, monitoring programmes have been modified and refined since they were established in the mid 1990s and as a consequence, not all of the species have been continuously monitored. To reduce inter-annual variability and avoid years with unusually high or low HAB occurrence and HAB species abundance, we have taken the mean and maximum abundance for those years that each species was monitored (up to a maximum of 6 years) and which best match the nutrient data. Table 4.2 gives the years from which HAB species data and nutrient data were taken and used in the analysis. The phytoplankton data were aggregated into coastal sea areas corresponding to those used to aggregate the nutrient data.

- 127 -

Table 4.2 Details of the phytoplankton data and nutrient data (loadings and winter concentrations) used in the statistical analysis.

HAB species abundance and nutrient loadings and loading ratios Genus or Species Years Loadings Alexandrium spp. 2002 - 2006 2002 - 2006 Dinophysis spp. 2002 - 2006 2002 - 2006 Pseudo-nitzschia spp. 2002 - 2006 2002 - 2006 Karenia mikimotoi 2000 - 2004 2000 – 2004 Prorocentrum lima 2002 - 2006 2002 - 2006 Prorocentrum minimum 2006 - 2008 2005 - 2006 HAB species abundance and winter nutrient concentrations and ratios Genus or Species Years Winter concentrations Alexandrium spp. 2002 - 2007 2002 - 2007 Dinophysis spp. 2002 - 2007 2002 - 2007 Pseudo-nitzschia spp. 2002 - 2007 2002 - 2007 Karenia mikimotoi 2000 - 2004 2000 – 2004 Prorocentrum lima 2002 - 2007 2002 - 2007 Prorocentrum minimum 2006 - 2008 2005 - 2007

4.3 Statistics

Linear regression analysis was used to test relationships between riverine loadings (measured and modelled), loadings ratios and HAB species abundance and between winter nutrient concentrations, concentration ratios and HAB species abundance. It has been shown that for phytoplankton data the variance of measurements is proportional to the mean (e.g. Tett & Wallis 1978). In such cases, logarithmic transformation is usually undertaken to satisfy the rules for statistical analysis. The phytoplankton data were therefore transformed (log10) before being used in the analysis presented here. For any given genus or species, if more than 40 % of the values for abundance were zero no regression was performed. The threshold used for determining significant relationships was a probability (P) of ≤ 0.050 (to three decimal places). The correlation coefficient was calculated to determine whether winter nutrient concentrations were related to riverine loadings (measured and modelled). Phytoplankton data (FSA(NI) unpubl.) from sampling stations in coastal waters of Northern Ireland at which monitoring has been conducted for a minimum of 10 years were used to determine whether there had been any temporal trends in HAB species abundance. Similarly, toxicity data from the north east of England was used to determine whether there had been any trends in the level of toxicity. Because of changes to the sampling stations, the toxicity data set was divided into two time periods: 1968 – 1992 and 1991 to 2007. The first data set was taken from Joint et al. (1997) and the second data set was provided by Marine Scotland, Marine

- 128 - Laboratory and the Cefas Weymouth Laboratory. The Mann-Kendall non-parametric test for monotonic trends (Hirsch & Slack, 1984; Hirsch et al. 1991) was used for trend analysis. This test was used to compare median values across years, and is less sensitive to extreme values and error distributions than is a trend analysis based on means.

4.4 Results

4.4.1 Statistical analyses

4.4.1.1 Data sets

When the phytoplankton data were compiled it became apparent that Lingulodinium polyedrum and Protoceratium reticulatum only occur infrequently and at low levels of abundance in UK and Irish coastal waters. Between 2005 and 2008, L. polyedrum was only observed at six sites in England and Wales and at these the maximum abundance recorded was 280 cells L-1. For Scottish coastal waters, L. polyedrum was only observed in 3 % of samples analysed between 2006 and 2008. A similar situation was found for coastal waters of the Republic of Ireland. Between 2005 and 2007, L. polyedrum was only recorded in 63 out of 2727 samples and the maximum abundance recorded was 2,440 cells L-1. Protoceratium reticulatum was not recorded at any sites in England and Wales between 2005 and 2008, or in Irish coastal waters between 2005 and 2007. This phytoplankter has not been recorded in samples collected from coastal waters of Northern Ireland. Given the infrequent and low abundance, both of these species were excluded from the statistical analysis and maps showing the geographical distribution have not been prepared for these two species.

4.4.1.2 Nutrient loadings and HAB species abundance

No significant relationships were evident between measured loadings and HAB species abundance. For both mean and maximum abundance, values of P were ≥ 0.173 and ≥ 0.383 for all regressions respectively. In contrast, there were 10 significant regressions between modelled nutrient loadings and the mean and maximum abundance of HAB species (Table 4.3, Figure 4.3).

- 129 -

Table 4.3 Results of regression analysis of HAB species abundance (mean and maximum) against modelled nutrient loading and loading ratios to UK coastal waters. P is probability ≤ 0.050 = significant regressions (in bold); R2, goodness of fit.

Genus or Species Nutrient Mean abundance Maximum abundance P R2 P R2

Alexandrium spp. NO3 0.878 0.001 0.923 0.000

NO2 0.608 0.010 0.929 0.000

NH4 0.587 0.011 0.789 0.003 TOxN 0.897 0.001 0.882 0.001 DIN 0.789 0.003 0.983 0.000

PO4 0.439 0.022 0.720 0.005 Si 0.781 0.003 0.872 0.001

Dinophysis spp. NO3 0.224 0.054 0.856 0.001

NO2 0.085 0.106 0.402 0.026

NH4 0.020 0.185 0.078 0.111 TOxN 0.227 0.056 0.855 0.001 DIN 0.109 0.092 0.587 0.011

PO4 0.039 0.148 0.230 0.053 Si 0.032 0.172 0.054 0.140

Pseudo-nitzschia spp. NO3 0.996 0.000 0.835 0.002

NO2 0.743 0.004 0.980 0.000

NH4 0.363 0.031 0.441 0.022 TOxN 0.984 0.000 0.818 0.002 DIN 0.844 0.002 0.986 0.000

PO4 0.384 0.028 0.683 0.006 Si 0.132 0.089 0.289 0.045

Karenia mikimotoi NO3 0.344 0.033 0.273 0.044

NO2 0.175 0.067 0.166 0.070

NH4 0.213 0.057 0.222 0.055 TOxN 0.336 0.036 0.269 0.047 DIN 0.276 0.044 0.221 0.055

PO4 0.190 0.063 0.242 0.050 Si 0.754 0.004 0.924 0.000

Prorocentrum lima NO3 0.068 0.132 0.068 0.132

NO2 0.096 0.111 0.049 0.153

NH4 0.045 0.158 0.025 0.193 TOxN 0.073 0.134 0.071 0.135 DIN 0.037 0.168 0.032 0.177

PO4 0.092 0.114 0.052 0.148 Si 0.424 0.029 0.302 0.048

Prorocentrum minimum NO3 0.127 0.094 0.142 0.088

NO2 0.488 0.020 0.373 0.033

NH4 0.127 0.094 0.081 0.121 TOxN 0.137 0.094 0.146 0.090 DIN 0.091 0.115 0.089 0.116

PO4 0.151 0.084 0.107 0.104 Si 0.027 0.204 0.048 0.166

- 130 -

Figure 4.3 Plots of significant relationships between modelled nutrient loadings and log10 abundance of HAB species in UK coastal waters.

4 4 4 Mean Dinophysis spp. NH4 Mean Dinophysis PO4 Mean Dinophysis spp. Si abundance abundance 3 3 spp. abundance 3

2 2 2 Abundance Abundance Abundance 1 1 1

0 0 0 0 3000 6000 9000 0 2000 4000 6000 0 10000 20000 30000 40000 50000 60000 Loading Loading Loading

2.0 2.0 4 NH NH Mean P. lima 4 Mean P lima DIN Maximum P. lima 4 abundance abundance 1.5 1.5 3 abundance

1.0 2 1.0 Abundance Abundance Abundancee 0.5 0.5 1

0.0 0.0 0 0 3000 6000 9000 0 7000 14000 21000 28000 35000 0 3000 6000 9000 Loading Loading Loading

4 5 5.0 NO2 DIN Mean P. mimimum Maximum P. lima Maximum P. lima Si 4 4.0 abundance 3 abundance abundance 3 3.0 2 2 2.0 4 Abundance Abundance 1 Abundance 1 1.0

0 0 0.0 0 200 400 600 800 1000 0 7000 14000 21000 28000 35000 0 10000 20000 30000 40000 50000 Loading Loading Loading 4 Si

3

2 Abundance 1 Maximum P. minimum abundance

0 0 10000 20000 30000 40000 50000 Loading

The mean abundance of Dinophysis spp. was negatively related to modelled ammonium

(NH4) and phosphate (DIP) loading and positively related to modelled silicate (Si) loading. The mean abundance of P. lima was negatively related to NH4 and dissolved inorganic nitrogen

(DIN) loading. The maximum abundance of P. lima was negatively related to NH4, nitrite (NO2) and DIN. The mean and maximum abundance of P. minimum was positively related to Si. In all cases values of R2 were < 0.204.

- 131 - 4.4.1.3 Ratios of nutrient loadings and HAB species abundance

There were no significant relationships between ratios of measured nutrient loading and the mean and maximum abundance of HAB species in UK waters. For ratios of modelled nutrient loadings, the data gave 10 significant relationships (Table 4.4, Figure 4.4). The mean and maximum abundance of Dinophysis spp. and mean abundance of Karenia mikimotoi was positively related to DIN:DIP ratio. The mean abundance of Dinophysis spp., Pseudo-nitzschia spp., P. lima and P. minimum and maximum abundance of Dinophysis spp., Karenia mikimotoi, and P. minimum were all negatively related to the DIN:Si loading ratio.

Table 4.4 Results of regression analysis of HAB abundance (mean and maximum) against ratios of modelled nutrient loadings to UK coastal waters. P is probability ≤ 0.050 = significant and significant regressions are given in bold; R2, goodness of fit.

Genus or Species Nutrient ratio Mean abundance Maximum abundance P R2 P R2 Alexandrium spp. DIN:DIP 0.566 0.012 0.972 0.000 DIN:Si 0.756 0.004 0.759 0.004 Dinophysis spp. DIN: DIP 0.000 0.387 0.015 0.200 DIN:Si 0.000 0.505 0.003 0.312 Pseudo-nitzschia spp. DIN: DIP 0.082 0.108 0.412 0.025 DIN:Si 0.014 0.217 0.091 0.110 Karenia mikimotoi DIN: DIP 0.013 0.208 0.051 0.134 DIN:Si 0.051 0.144 0.043 0.154 Prorocentrum lima DIN: DIP 0.146 0.086 0.157 0.082 DIN:Si 0.045 0.170 0.053 0.160 Prorocentrum minimum DIN: DIP 0.439 0.025 0.360 0.035 DIN:Si 0.001 0.416 0.001 0.429

- 132 -

Figure 4.4 Plots of significant relationships between modelled ratios of nutrient loadings and log10 abundance of HAB species in UK coastal waters.

4 4 5 Mean Dinophysis spp. DIN:Si DIN:DIP abundance Mean Dinophysis spp. 4 3 3 abundance 3 2 2 DIN:Si 2 Mean Abundance 1 Abundance 1 Abundance 1 Pseudo nitzschia spp. abundance 0 0 0 0 50 100 150 0123401234 Ratio Ratio Ratio

4 1.0 5 Mean Karenia Mean P. minimum 0.8 Mean P. lima 4 3 mikimotoi abundance abundance abundance 0.6 3 2 0.4 2 DIN:DIP DIN:Si Abundance 1 Abundance Abundance 0.2 DIN:Si 1

0 0.0 0 0 1020304050 01234012345 Ratio Ratio Ratio

5 4 7 DIN:DIP Maximum Dinophysis 6 Maximum Karenia 4 3 spp. abundance 5 mikimotoi abundance 3 4 2 DIN:Si 2 3 2 Abundance Maximum Dinophysis Abundance 1 Abundance 1 spp. abundance DIN:Si 1 0 0 0 0 50 100 150 01234012345 Ratio Ratio Ratio

7 6 Maximum P.mimimum 5 abundance 4 3 DIN:Si 2 Abundance 1 0 012345 Ratio

4.4.1.4 Correlations between loadings and winter concentrations

Measured and modelled nutrient loadings (2000 – 2006 data) were significantly correlated (Table 4.5) and both measured and modelled nutrient loadings were significantly correlated with measured winter concentrations of nutrients (Table 4.5).

- 133 -

Table 4.5 The relationship between measured and modelled nutrient loadings to UK coastal areas and between measured and modelled loadings and mean winter nutrient concentrations.

Measured vs modelled loadings Nutrient Correlation coefficient Degrees of Freedom P NH4 0.844 32 <0.001 NO3 0.937 32 <0.001 PO4 0.915 32 <0.001 Loadings vs winter concentrations Correlation coefficient Degrees of Freedom P measured modelled measured modelled measured modelled NH4 0.217 0.340 36 30 0.191 0.057 NO3 0.048 0.065 35 39 0.780 0.728 PO4 0.634 0.631 36 30 <0.001 <0.001

4.4.1.5 Winter concentrations, ratios and HAB species abundance

The results of the regression analysis of NH4 and DIN concentrations against HAB species abundance and between DIN:DIP and DIN:Si ratios and HAB species abundance in UK waters are given in Table 4.6 and Figure 4.5. Significant negative relationships were evident between the mean abundance of Dinophysis spp. and the winter concentration of NH4 and DIN; the mean abundance of Karenia mikimotoi and the winter concentration of NH4; maximum abundance of Dinophysis spp. and DIN. There were also significant negative relationships between the mean and maximum abundance of P. lima and the winter molar ratio of DIN:Si. In all cases values of R2 were ≤ 0.258.

Table 4.6 Results of regression analysis of HAB species abundance and winter concentrations of NH4, DIN, and molar ratios of DIN:DIP and DIN:Si. Data from UK coastal waters.

Genus or Species Nutrient or ratio Mean abundance Maximum abundance P R2 P R2

Alexandrium spp. NH4 0.752 0.003 0.677 0.006 DIN 0.409 0.023 0.500 0.015 DIN:DIP 0.607 0.010 0.467 0.021 DIN:Si 0.868 0.001 0.738 0.004

Dinophysis spp. NH4 0.049 0.116 0.081 0.092 DIN 0.045 0.127 0.020 0.169 DIN:DIP 0.279 0.045 0.860 0.001 DIN:Si 0.057 0.128 0.055 0.130

Pseudo-nitzschia spp. NH4 0.065 0.102 0.116 0.075 DIN 0.092 0.092 0.203 0.053 DIN:DIP 0.876 0.001 0.643 0.008

- 134 - Table 4.6 continued

Genus or Species Nutrient or ratio Mean abundance Maximum abundance P R2 P R2 Pseudo-nitzschia spp. DIN:Si 0.389 0.028 0.455 0.021

Karenia mikimotoi NH4 0.023 0.157 0.059 0.110 DIN 0.107 0.087 0.291 0.038 DIN:DIP 0.877 0.001 0.675 0.007 DIN:Si 0.362 0.032 0.378 0.030

Prorocentrum lima NH4 0.061 0.112 0.136 0.073 DIN 0.134 0.078 0.215 0.054 DIN:DIP 0.468 0.022 0.415 0.028 DIN:Si 0.019 0.202 0.007 0.258

Prorocentrum minimum NH4 0.105 0.337 0.140 0.325 DIN 0.725 0.034 0.636 0.061 DIN:DIP 0.455 0.146 0.510 0.116 DIN:Si 0.273 0.287 0.306 0.256

Figure 4.5 Plots of the significant relationships between UK winter concentrations of nutrients and nutrient ratios and log10 HAB species abundance.

4 4 4 NH DIN NH 4 Mean Karenia 4 3 3 3 mikimotoi abundance Mean Dinophysis spp. Mean Dinophysis spp. abundance 2 abundance 2 2

Abundance Abundance Abundance 1 1 1

0 0 0 0 5 10 15 20 25 0 50 100 150 200 250 300 0 5 10 15 20 25 Concentration Concentration Concentration

5 2.0 3 Maximum Dinophysis Mean P. lima 4 spp. abundance DIN DIN:Si DIN:Si 1.5 abundance 2 Maximum P. lima 3 abundance 1.0 2 1 Abundance Abundance 0.5 Abundance 1 0 0.0 0 0 100 200 300 0 5 10 15 20 25 30 35 0 5 10 15 20 25 30 35 Concentration Nutrient ratio Nutrient ratio

The results of regression analysis using the combined UK and Irish data set are shown in Table 4.7 and significant regressions are plotted in Figure 4.6. There were significant negative relationships between mean Dinophysis spp. abundance and TOxN, NO2 and DIP; mean Karenia mikimotoi abundance and NO2 and between the mean abundance of P. minimum and Si. For maximum abundance, Dinophysis spp. was negatively related to TOxN, NO3, NO2 and DIP and maximum abundance of P. minimum was negatively related to Si. The maximum abundance of Dinophysis spp. and P. lima was also negatively related to the ratio of winter TOxN:Si.

- 135 -

Table 4.7 Results of regression analysis of HAB species abundance and winter nutrient concentrations and ratios using the combined UK and Irish data set. P is probability ≤ 0.050 = significant and significant regressions are given in bold; R2, goodness of fit.

Genus or Species Nutrient or ratio Mean abundance Maximum abundance P R2 P R2

Alexandrium spp. NO3 0.176 0.039 0.358 0.018

NO2 0.159 0.042 0.102 0.056 TOxN 0.179 0.039 0.395 0.016 DIP 0.091 0.058 0.237 0.029 Si 0.216 0.032 0.104 0.054 TOxN:DIP 0.957 0.000 0.580 0.007 TOxN:Si 0.249 0.031 0.569 0.008

Dinophysis spp. NO3 0.057 0.075 0.014 0.121

NO2 0.016 0.118 0.005 0.156 TOxN 0.041 0.088 0.009 0.140 DIP 0.049 0.078 0.021 0.105 Si 0.657 0.004 0.377 0.016 TOxN:DIP 0.328 0.023 0.840 0.001 TOxN:Si 0.070 0.075 0.037 0.097

Pseudo-nitzschia spp. NO3 0.314 0.022 0.530 0.009

NO2 0.947 0.000 0.793 0.002 TOxN 0.251 0.029 0.443 0.013 DIP 0.804 0.001 0.993 0.000 Si 0.242 0.028 0.260 0.026 TOxN:DIP 0.961 0.000 0.851 0.001 TOxN:Si 0.895 0.000 0.838 0.001

Karenia mikimotoi NO3 0.240 0.036 0.545 0.010

NO2 0.041 0.105 0.162 0.051 TOxN 0.144 0.055 0.359 0.022 DIP 0.065 0.083 0.198 0.041 Si 0.116 0.061 0.239 0.035 TOxN:DIP 0.688 0.005 0.837 0.001 TOxN:Si 0.254 0.037 0.351 0.025

Prorocentrum lima NO3 0.189 0.038 0.232 0.032

NO2 0.911 0.000 0.658 0.004 TOxN 0.140 0.049 0.164 0.044 DIP 0.112 0.054 0.177 0.039 Si 0.677 0.004 0.917 0.000 TOxN:DIP 0.502 0.011 0.565 0.008 TOxN:Si 0.094 0.067 0.038 0.101

Prorocentrum minimum NO3 0.928 0.001 0.725 0.012

NO2 0.274 0.107 0.202 0.143 TOxN 0.980 0.000 0.766 0.008 DIP 0.085 0.197 0.072 0.213 Si 0.027 0.305 0.022 0.322 TOxN:DIP 0.949 0.000 0.797 0.006 TOxN:Si 0.198 0.134 0.290 0.093

- 136 -

Figure 4.6 Plots of the significant relationships between winter concentrations of nutrients and log10 mean abundance of HAB species in UK and Irish coastal waters.

4 4 4

TOxN NO2 DIP 3 3 3 Mean Dinophysis spp. Mean Dinophysis spp. Mean Dinophysis spp. 2 abundance 2 abundance 2 abundance

Abundance Abundance Abundance 1 1 1 0 0 0 0 100 200 300 0246802468 Concentration Concentration Concentration

4 4 5 NO2 Si TOxN 4 3 3 Maximum Dinophysis spp. abundance Mean Karenia mikimotoi 3 2 abundance 2 Mean P. minimum abundance 2 Abundance 1 Abundance 1 Abundance 1

0 0 0 0 0.5 1 1.5 2 2.5 01020300 100 200 300 Concentration Concentration Concentration

5 5 5 NO3 NO2 Maximum Dinophysis 4 4 4 Maximum Dinophysis spp. abundance spp. abundance Maximum Dinophysis 3 3 spp. abundance 3 DIP 2 2 2 Abundance Abundance Abundance 1 1 1

0 0 0 0 100 200 300 0246802468 Concentration Concentration Concentration 5 5 5 Si TOxN: Si TOxN: Si 4 4 4 Maximum P. lima Maximum abundance 3 P. minimum 3 Maximum Dinophysis 3 abundance spp. abundance 2 2 2 Abundance Abundance Abundance 1 1 1

0 0 0 01020300 102030400 10203040 Concentration Nutrient ratio Nutrient ratio

4.4.1.6 Time series analysis

For the phytoplankton data sets from coastal waters of Northern Ireland, only time-series that were at least 10 years in length were analysed. The results of the Mann – Kendall time series analysis are presented in Table 4.8. A significant negative trend was only evident for Dinophysis acuminata and for total Dinophysis spp. at one site in Strangford Lough. No trends were evident in the PSP toxicity data from the north east coast of England (P = 0.292 and 0.925 for the first and second time periods respectively).

- 137 -

Table 4.8 The results of trend analysis for phytoplankton data sets from coastal waters of Northern Ireland. The numbers in parenthesis refer to the number of sites from which data were collected and used in the analysis. nt = no significant trend.

Genus Location or Species Lough Carlingford Strangford Dundrum Larne Foyle (1) Lough (1) Lough (4) Bay (2) Lough (1) Alexandrium spp. nt nt nt nt nt Dinophysis acuminata nt nt negative trend nt nt (P = 0.023) D. acuta nt nt nt nt nt D. norvegica nt nt nt nt nt D rotundata nt nt nt nt nt Dinophysis spp. nt nt negative trend nt nt (P = 0.034) Pseudo-nitzschia spp. nt nt nt nt nt Karenia mikimotoi nt nt nt nt nt Prorocentrum lima nt nt nt nt nt P. minimum nt nt nt nt nt

4.4.2 The distribution of HAB species in UK and Irish coastal waters

Maps showing the geographical distribution of the HABs species under review in this study are presented here (except Lingulodinium polyedrum and Protoceratium reticulatum because their occurrence was so infrequent). In Part 5, we discuss the factors that may be responsible for the distribution of these phytoplankters in UK and Irish coastal waters. During the period 2002 to 2007, species of Alexandrium were generally more abundant in shallow estuaries in the south west of England; western waters of Ireland; the west and east coast of Scotland and in waters around the Orkney and Shetland Islands (Figure 4.7). Abundance of these phytoplankters was low (mean abundance was < 100 cells L-1 at 83 % of sites) although a high abundance (≥ 106 cells L-1) was observed at a few locations. Over the same period of time, species of Dinophysis were more abundant along the south and west coast of Ireland, the west and south east of Scotland and north east and south west of England (Figure 4.7). The abundance of Dinophysis spp. was low in UK and Irish coastal waters with a mean abundance of < 100 cells L-1 at 77 % of locations sampled. Species of Pseudo-nitzschia were widespread throughout UK and Irish coastal waters (Figure 4.7) between 2002 and 2007, although less abundant in waters along the south eastern coast of Ireland and England. These phytoplankters are much more abundant than species of Alexandrium and Dinophysis with a mean and maximum abundance ≥ 1,000 and 106 cells L-1 at 53 % and 2 % of sample sites respectively. Between 2000 and 2004, Karenia mikimotoi was more abundant in coastal waters of the south west of England, along the coast of southern and western Ireland and in coastal waters

- 138 - around Scotland (Figure 4.7). The abundance of Prorocentrum lima was low in coastal waters of the UK and Ireland between 2005 and 2007. For all locations, the mean abundance was ≤ 302 cells L-1 and at 75 % of sites the mean abundance was ≤ one cell L-1. As a consequence it is difficult to obtain a clear picture of the geographical distribution of this species, although the data shown in (Figure 4.7) are suggestive of a higher abundance in coastal waters of the south west of England, west of Ireland and western and northern coastal waters of Scotland. Prorocentrum minimum (2006 to 2008) was more abundant than P. lima, and greater abundance was observed in coastal waters off the south west of England, west coast of Ireland and west coast of Scotland (Figure 4.7). The maximum abundance of this species was recorded in waters to the west and North of Scotland.

- 139 -

Figure 4.7 The mean abundance of HAB species in UK and Irish coastal waters.

61 0 to 1 0 to 1 2 to 99 2 to 99 100 to 999 100 to 999 1000 to 9999 1000 to 28030 59 10000 to 439200

57 ude t

i 55 t La

53

51

Alexandrium spp. Dinophysis spp. 49 61 0 to 1 0 to 1 2 to 99 2 to 99 100 to 999 100 to 999 1000 to 9999 1000 to 9999 59 10000 to 99999 10000 to 99999 100000 to 690000 100000 to 418100

57 ude t

i 55 t La

53

51

Pseudo-nitzschia spp. Karenia mikimotoi 49

61 0 to 1 0 to 1 2 to 99 2 to 99 100 to 999 100 to 302 1000 to 9999 59 10000 to 99999 100000 to 420000

57

55 Latitude

53

51

Prorocentrum lima Prorocentrum minimum 49 10.5 8.7 6.9 5.1 3.4 1.6 0.2 2.010.5 8.7 6.9 5.1 3.4 1.6 0.2 2.0 Longitude Longitude

- 140 -

Figure 4.7 continued The maximum abundance of HAB species in UK and Irish coastal waters.

61 0 to 1 0 to 1 2 to 99 2 to 99 100 to 999 100 to 999 1000 to 9999 1000 to 9999 10000 to 99999 59 10000 to 83870 100000 to 999999 1000000 to 17011000

57

55 Latitude

53

51

Alexandrium spp. Dinophysis spp. 49 61 0 to 1 0 to 1 2 to 99 2 to 99 100 to 999 100 to 999 1000 to 9999 1000 to 9999 59 10000 to 99999 10000 to 99999 100000 to 999999 100000 to 999999 1000000 to 6058000 1000000 to 1786000

57

55 Latitude

53

51

Pseudo-nitzschia spp. Karenia mikimotoi 49 61 0 to 1 0 to 1 2 to 99 2 to 99 100 to 999 100 to 3521 1000 to 9999 59 10000 to 99999 100000 to 999999 1000000 to 4125000

57

55 Latitude

53

51

Prorocentrum lima Prorocentrum minimum 49 10.5 8.9 7.3 5.6 4.0 2.4 0.8 0.9 10.5 8.7 6.9 5.1 3.4 1.6 0.2 2.0 Longitude Longitude

- 141 - 4.5 Discussion

4.5.1 Introduction

The general finding from the statistical analysis is that the abundance of HAB species that occur in UK and Irish coastal waters is not related to anthropogenic nutrient enrichment (as determined by nutrient loading and winter nutrient concentrations). However, before discussing key aspects of the results we first consider the adequacy of the data and the reliability of the statistical and interpretational analyses upon which this conclusion is based.

4.5.2 Data sets and analysis

Attempts to relate changes in the occurrence of HABs to environmental pressure such as nutrient enrichment are frequently confounded by a lack of data or the necessity of using data collected for a different purpose. In relation to this, Smayda (2008) argued that:

“the data invariably are inadequate to analyze the putative eutrophication-HAB relationship, and the evidence adducted in support of the conclusions is often circumstantial, although more quantitative evidence is beginning to emerge…”

There are limits to how the data compiled for this study can be used. The data have been quality assured and are considered to be of high quality and reliable, but it should be noted that these data were not collected for the purpose of relating enrichment to the occurrence of HABs. The nutrient data were collected as part of environmental monitoring programmes and the phytoplankton data were mainly from monitoring programmes designed to protect human health. In the case of the latter, sampling stations have changed, sampling protocols revised and the list of species identified and counted has been refined since the introduction of the programmes in the mid 1990s. To reduce inter-annual variability, multi-year mean values of loadings, winter nutrient concentrations and HAB species abundance were used in the analysis. Mean values were derived from consecutive years (up to a maximum of 6). The years were selected to derive the best geographical coverage for each species and to best match the nutrient data. This meant using different years for some of the species and resulted in some mismatch between the phytoplankton and nutrient data (Table 4.2). The advantages of using multiple years to reduce inter-annual variability and improve the precision of estimates of nutrients and HAB species abundance were considered to outweigh any disadvantages resulting from a mismatch between phytoplankton and nutrient data. The data sets were merged by dividing the coast line into relatively large sections based on river catchment. Therefore our main statistical analysis involved a geographical comparison,

- 142 - between coastal areas with different nutrient loadings, on the scale of the islands of Britain and Ireland. Such an analysis does not rule out the possibility that nutrient – HAB correlations might be found if data were analysed by water bodies rather than coastal sea area, or that we might have found correlated trends in HABs and nutrient time-series from within a particular water body (but see the results of the time-series analysis below). Regression analysis, as employed here, is widely used to model the relationship between two variables by fitting a linear equation (y = a.x + b) to data. The objective is to determine (or predict) the variation in variable-y that results from variation in variable-x (a and b in the equation are the intercept and slope of the regression line respectively). The terms dependent and independent variable are used for the y and x variables respectively and this implies that change in the dependent variable (y) is caused by change in the independent variable (x). In our case the dependent variable is HAB species abundance and the independent variable is nutrient loading and concentration (proxies for enrichment) and the implication of a significant positive regression is that high HAB species abundance can be explained by variation in the level of enrichment. For this reason, regression analysis was used rather than correlation coefficient because the latter is a test of whether two variables co-vary, and our interest was in testing the hypothesis that: the occurrence of HABs and HAB species abundance increases with anthropogenic nutrient enrichment. The generally accepted standard probability (P) value of ≤ 0.05 was used to determine whether a regression was significant i.e. that there was a linear relationship between abundance and enrichment that could have occurred by chance in only 1 analysis in 20. We tested for the existence of relationships amongst 324 pairs of data-sets, and so might expect to find P ≤ 0.05 in about 16 even on the null hypothesis of no true relationship. In situations where a large number of relationships are being tested it is not unusual to use a value of P = 0.01 (and in some situations 0.001). Nevertheless, we decided to use the more widely used threshold of P ≤ 0.05 and evaluate each apparently significant relationship on its merits. Finding a significant regression does not necessarily mean that all of the variation in the dependent variable is the result of variation in the independent variable. The value of R2 (the square of the correlation coefficient) gives the proportion of the variation in the y- variable explained by (a linear function of) variation in the x- variable. The value of R2 can vary between 0 and +1 and values close to zero indicate that only a small amount of the variation in the y- variable can be explained by change in the x- variable using a linear regression model. For some of the significant regressions, the data plots show that some of the individual data appear to be extreme values or ‘outliers’. Such data points can have a large influence on the

- 143 - regression line. Additional analysis could be undertaken to investigate this but was beyond the scope of this study.

4.5.3 Interpretation of results

4.5.3.1 Introduction

Our starting point was the hypothesis that:

the occurrence of HABs and HAB species abundance increases with anthropogenic nutrient enrichment.

The statistical 'null' hypothesis was therefore that:

the values of the HAB indicators varied randomly with respect to the values of the nutrient indicators.

It is this hypothesis that was rejected whenever a relationship between a HAB species indicator and a nutrient indicator was found likely to be due to chance in less than 1 case in 20. Table 4.9 summarises the significant relationships between nutrient indicators and HAB indicators: i.e., those for which the null hypothesis was rejected. In some of these significant cases we attempt a scientific explanation; such explanations are meant to be indicative rather than definitive. Furthermore, it should be kept in mind that some of them might not be truly 'significant' but we don't know which.

4.5.3.2 HAB species abundance, nutrient loadings and winter concentrations

The results of the regression analysis summarised in Table 4.9, show that of the 168 relationships between HAB species abundance and modelled loadings and winter concentrations examined, only 24 were significant. This is based on using P = 0.05 for the level of significance which means that 8 of these significant regressions could be the result of chance (random error) and it is noteworthy that if a probability of 0.01 is used for the level of significance, only 2 of the regressions are significant. All but 3 of the significant regressions were negative. These results show that in general HAB species abundance in UK and Irish waters was not influenced by enrichment with nitrogen and phosphorus. In fact, the significant negative regressions imply that these HAB species were more abundant in un-enriched waters. This does not mean that nutrients suppress the growth of harmful algae. A more likely explanation is that the relevant algae are naturally more abundant in waters to the west and north of our islands, in which there is least anthropogenic nutrient enrichment. That is to say, the apparent negative relationships may be an artefact of non-random distribution of nutrient enrichment.

- 144 -

Table 4.9 A summary of significant regressions of HAB species abundance (mean and maximum) against modelled nutrient loading and loading ratios, winter nutrient concentrations and concentration ratios. Negative and positive regressions are identified as –ve and +ve.

Genus or Species Nutrient Mean abundance Maximum abundance P R2 P R2 Modelled loadings

Dinophysis spp. NH4 0.020 -ve 0.185

PO4 0.039 -ve 0.148 Si 0.032 +ve 0.172

Prorocentrum lima NH4 0.045 -ve 0.158 0.025 -ve 0.193 DIN 0.037 -ve 0.168 0.032 -ve 0.177

NO2 0.049 -ve 0.153 Prorocentrum minimum Si 0.027 +ve 0.204 0.048 +ve 0.166 Winter concentrations

Dinophysis spp. NH4 0.049 -ve 0.116 DIN 0.045 -ve 0.127 0.020 -ve 0.169

NO3 0.014 -ve 0.121

NO2 0.016 -ve 0.118 0.005 -ve 0.156 TOxN 0.041 -ve 0.088 0.009 -ve 0.140 DIP 0.049 -ve 0.078 0.021 -ve 0.105

Karenia mikimotoi NH4 0.023 -ve 0.157

NO2 0.041 -ve 0.105 Prorocentrum minimum Si 0.027 -ve 0.305 0.022 -ve 0.322 Modelled loadings ratios Dinophysis spp. DIN: DIP 0.000 +ve 0.387 0.015 +ve 0.200 DIN:Si 0.000 -ve 0.505 0.003 -ve 0.312 Pseudo-nitzschia spp. DIN:Si 0.014 -ve 0.217 Karenia mikimotoi DIN: DIP 0.013 +ve 0.208 DIN:Si 0.043 -ve 0.154 Prorocentrum lima DIN:Si 0.045 -ve 0.170 Prorocentrum minimum DIN:Si 0.001 -ve 0.416 0.001 -ve 0.429 Winter concentration ratios Dinophysis spp. TOxN:Si 0.037 -ve 0.097 Prorocentrum lima DIN:Si 0.019 -ve 0.202 0.007 -ve 0.258 TOxN:Si 0.038 -ve 0.101

The significant positive relationships between mean Dinophysis spp. abundance and Si loading and between the mean and maximum abundance of Prorocentrum minimum and Si loading are counter intuitive because dinoflagellates do not require silicate for growth. A possible explanation is that the high Si waters are those that are little enriched with anthropogenic N and P. In enriched waters, extra diatom growth might remove dissolved silica before giving way to other algae; and it is the un-enriched waters where Dinophysis and Prorocentrum are naturally found. If that is true, then we are again seeing an artefact of the non- random distribution of nutrient enrichment.

- 145 - For the 24 significant regressions, values of R2 ranged from 0.078 to 0.322 (Table 4.9). Thus, the linear model explained only 8 to 32 % of the variation in HAB abundance as a result of variation in loading and winter concentrations of nutrients. One implication of this is that factors other than loadings and winter concentrations were also influencing abundance.

4.5.3.3 HAB species abundance and nutrient ratios

Of the 72 analyses for potential relationships between HAB species abundance and nutrient ratios derived from modelled loadings and winter concentrations, only 14 were significant at P = 0.050 (6 at P = 0.01). The mean and maximum abundance of Dinophysis spp. and maximum abundance of Karenia mikimotoi was positively related to the DIN:DIP loading ratio. Variation in the ratio explained between 20 and 39 % of the variation in HAB species abundance (Table 4.9). But whereas greater abundance of these two HAB species was associated with high N:P ratios, they were not, as the previous section discussed, associated with enriched waters. Several published papers (see Part 3) have argued that perturbations of nutrient ratios relative to their natural, or Redfield ratio, value of 16:1 (atoms N: atom P) will favour harmful algae, and these significant relationships would seem to support that argument. Nevertheless, if such effects were strong, they should have led to a higher proportion of significant relationships than Table 4.9 reports. All of the other significant regressions (Table 4.9) were between HAB species abundance and ratios of N (as TOxN and DIN):Si loadings and ratios of winter concentrations and were negative. That is, higher abundance was associated with low N:Si ratios. This would appear contrary to the arguments usually presented in the literature (see Part 3), which have increases in N:Si changing the balance of organisms in favour of harmful algae. However, an alternative explanation is possible: waters with increased N:Si are, typically, those suffering anthropogenic enrichment, and are in coastal waters where hydrodynamic conditions are unsuitable for most harmful lifeforms. Thus, the negative relationships with N:Si ratios could be an artefact of the significant negative relationships between abundance and loadings/ concentrations. It should also be remembered that our data are for winter ratios, or loading ratios, which may not correspond to actual nutrient ratios during summer, when harmful algae are most likely to be abundant.

4.5.3.4 Time-series analysis

The analysis of time-series is fraught with difficulties, and a suite of methods has been developed to deal with these (see e.g. Chatfield 1989). Trends are commonly extracted from time-series by the fitting of a linear regression (y = a.x + b) but a key assumption of the standard

- 146 - regression model is that each pair of x-y data is independent of each other pair. This is often not the case with time-series, when events in one year can influence events in the next year. In such cases (exemplified by Alexandrium spp. cysts) a year-to-year temporal autocorrelation might be falsely interpreted as a trend. For this reason, in the study presented here the Mann-Kendall non- parametric test for monotonic trends (Hirsch & Slack, 1984; Hirsch et al. 1991) was used for trend analysis. A major concern with time series analysis is the length of the time-series. The difficulty with short time series such as ours is that some of the factors which influence phytoplankton composition and the occurrence of HABs have decadal (e.g. the North Atlantic Oscillation and El Niño Southern Oscillation) or longer time scales. White (1987) and see also Martin et al. (2009) linked periodicity in the intensity of toxicity events in the Bay of Fundy (Canada) to an 18.6 year tidal cycle). Ideally 2 periods are required to begin to determine that there is a periodic variation which has implication for the length of time series required. Dale et al. (2006) were of the opinion that the ideal basic requirement would be a time-series of at least 30 consecutive years of regular monitoring. Only the PSP toxicity data from the north east of England meets this requirement but even this might not be sufficiently long since Borkman et al. (2009) suggest that even the longest phytoplankton time series ( 5 decades), may not be sufficiently long to resolve questions about the long term effects of climate change and anthropogenic nutrients. Time-series of ≤ 10 years are generally considered too short to identify trends. There are however, a number of examples in the scientific literature where short time series (< 10 years) have been used to link an apparent trend in the occurrence of HABs to a particular environmental pressure but where additional data show that the original conclusions were premature. We illustrate this point with two examples from the literature review in Part 3. The first is taken from the study by Liu and Wang (2004) who present data on changes in the frequency of red tides in coastal waters of Guangdong province (China) in relation to changes in population and industrial development (Figure 3.11A). If only the first part of the time series is considered, then between 1980 and 1990, the number of red tides in coastal waters of middle Guangdong province increased from  one to 15 and over the same period GDP increased (from  54 to 300 x108 Yuan). However, when the second part of the time-series is considered, it is apparent that while GDP continued to increase between 1990 and 2001, there was no corresponding increase in red tides. In fact, the 1990 peak in red tide occurrence was followed by a marked decrease. Unless there was a decrease in monitoring or major reduction in nutrient input to coastal waters (but Qi et al. (2004) state that in 1997, 2.8 billion tonnes of sewage was discharged into the Pearl River estuary) the data suggest that some other pressure was overriding the effect of enrichment. Liu and Wang (2004)

- 147 - were of the opinion that red tide initiation was due to natural causes but that nutrient enrichment enhanced red tides. The second example is taken from Hodgkiss and Ho (1997) who report on an eight year time series of data from Tolo Harbour (Hong Kong) and suggest that as the N:P ratio declined, there was a corresponding increase in red tides (Figure 3.20A). However, with the benefit of additional data (Figure 3.20B), it is clear that between 1990 and 1995 when the N:P ratio was approximately Redfield (16:1), there was a low occurrence of red tides. The analysis of the rather short time series of phytoplankton data from coastal waters of Northern Ireland (Table 4.8) therefore needs to be interpreted with some caution. With the exception of the negative trend in Dinophysis acuta (and total Dinophysis spp.) at one site, the analysis suggests that there has been little change in the abundance of HAB species over the last 10 years. Bresnan et al. (2008) report a decreasing trend in PSP in Scottish shellfish since the 1990s although there was no trend in the PSP toxicity data from the north east coast of England. The latter finding is consistent with the analysis of the 23 year time-series (1968 – 1990) of the same data by Wyatt and Saborido-Rey (1993) who concluded that no obvious trend was apparent in the time series.

4.6 Conclusions

On the basis of the discussion presented above we are of the opinion that the data sets are suitable for the analysis undertaken. We therefore reject our hypothesis that: the occurrence of HABs and HAB species abundance increases with anthropogenic nutrient enrichment (proxy: riverine loading and mean winter concentrations of nutrients and conclude that the abundance of the HAB species that occur in UK and Irish coastal waters is not related to anthropogenic nutrient enrichment.

- 148 - Part 5

Discussion and Synthesis

5.1 Introduction

In Part 3 some of the scientific arguments linking harmful algal blooms and anthropogenic nutrient enrichment were reviewed and exemplified in case studies from several parts of the world. Smayda (2008) has proposed two, non-exclusive global hypotheses (climate change [meaning the long-term trend in global warming] and eutrophication), and Heisler et al. (2008) argued that increased nutrient pollution promotes the development and persistence of many HABs, and was one of the reasons for their expansion in the U.S. and the world. We concluded that these global hypotheses could not be supported from the available data. In Part 4, the nutrient enrichment → HAB hypothesis was tested using data sets from the UK and Ireland. Our hypothesis was:

The occurrence of HABs and HAB species abundance increases with anthropogenic nutrient enrichment.

Riverine loading and mean winter concentrations of nutrients were used as proxies for enrichment and regression analysis used to determine whether HAB species abundance in UK and Irish coastal waters was related to enrichment. The results of the regression analysis (summarised in Table 4.9) show that only 24 of the 168 relationships examined between HAB species abundance and modelled loadings and winter concentrations were significant. These results show that in general HAB species abundance in UK and Irish waters was not influenced by enrichment with nitrogen and phosphorus. In fact, all but 3 of the significant regressions were negative implying that these HAB species were more abundant in un-enriched waters. This does not mean that nutrients suppress the growth of harmful algae and a more likely explanation is that the relevant algae are naturally more abundant in waters in which there is least anthropogenic nutrient enrichment. In this part of the report we examine an alternative hypothesis, which is:

there is no single causal mechanism for all the phenomena labelled HABs. Instead there are a number of explanatory hypotheses, suggesting a range of ecohydrodynamically mediated relationships (including none) between nutrient enrichment and HABs.

- 149 - The manner by which key ecohydrodynamic processes influence phytoplankton growth and the accumulation of biomass is discussed and consideration given to how these processes intersect with the ecophysiology of particular HAB species to determine their response to nutrient enrichment and govern their geographical distribution in UK and Irish coastal waters. The relationship between aquaculture and the abundance of HAB species is also briefly considered. The final part of this section is a synthesis of our findings, drawing on information presented and conclusions drawn from previous sections and from which some overall conclusions are drawn.

5.2 Ecohydrodynamics: Some General Principles

According to Tett et al. (2007), the application of the ecosystem approach to the sea, and the use of the concept of ecosystem health to assess disturbance, requires the spatial extent of marine ecosystems to be defined in functional and management terms. Once delimited, an ecohydrodynamic unit can be characterized by: (i) its physical conditions; (ii) its typical primary producers (in the absence of anthropogenic interference); (iii) significant ecosystem features emerging from such primary producer dominance and from biogeography. In the present context we need to consider water bodies or sea areas that can be treated as a functional unit for the purposes of assessing the cause and nature of HABs. Such units should be large enough for ecosystem structure and function to be controlled more by internal processes than by external conditions, should each be subject to one dominant set of hydrodynamic processes, and if possible defined by hydrographic features. The latter might be static seabed topography such as depth (with consequences for mixing or water column illumination) or dynamic water properties exemplified by the use of salinities to distinguish coastal from offshore waters by the OSPAR Comprehensive Procedure (OSPAR, 2009). These ideas can be illustrated by reference to a coastal Region of Restricted Exchange (RRE) as presented in Figure 5.1. A region of restricted exchange is a water body partly surrounded by land, so that water exchange with the adjacent sea is restricted and can be defined by an exchange rate that gives the daily proportion of RRE water replaced by sea water. Tett et al. (2003a) define a region of restricted exchange as:

“a water that is enclosed on three sides, so having restricted exchange with the sea; and in which the ratio of daily freshwater inflow to mean volume is less than 0.1.”

- 150 -

Figure 5.1 The ecohydrodynamics of regions of restricted exchange (RREs).

This definition excludes estuaries that are strongly flushed by river discharge and Tett et al (2003a) argue that while the definition specifies restricted exchange, it does not require the exchange volume to be small. Exchange takes place as a result of freshwater entering the RRE and tidal inflow and outflow although wind driven water movement may also play a role. There

- 151 - is therefore a continual replacement of water and the exchange rate (dilution or flushing rate) for a RRE with a volume V and from which a small volume dV is removed and replaced by new water in each time interval dT, has a dilution rate (D) equivalent to (dV/dT) / (1/V). The units of D are reciprocal time and the residence or flushing time is therefore 1/D. Thus, if an RRE has an exchange rate of 0.2 d-1, the water within it would have a mean residence time of 5 days, assuming that it can be treated as a well mixed box. The greater the exchange rate, the more the contents of the RRE resembles those in the external sea (Gowen et al. 1983). Conversely, if an RRE of low exchange rate is enriched with nutrients, and other conditions are favourable, algal blooms are likely to occur. Tett et al (2003a) explore this matter with data from a number of RREs in western Europe. Rates of lateral exchange, mixing, or dispersion within and between water bodies are, in our view, one of the key determinants of algal blooms. A second crucial set of hydrodynamic characteristics involve the strength of vertical mixing and its consequences for the illumination experienced by primary producers. Solar warming of the surface of the sea, or the input of freshwater, potentially creates superficial layers of lower-density water. Phytoplankters within such layers are retained close to the sea-surface and are well-illuminated throughout the year in tropical and subtropical waters and during spring and summer in temperate latitudes. Nutrient inputs to such layers (either natural or anthropogenic, in urban waste water or enriched river discharges) are likely to stimulate algal blooms unless planktonic animals or benthic filter- feeders consume the increased algal production. Conversely, strong vertical mixing due to wind, tidal currents, or surface cooling, carries phytoplankters away from the light and can resuspend large quantities sediment from the sea-bed and reduce the depth to which light can penetrate.

5.3 Ecophysiology: Phytoplankton Lifeforms and Species Succession

Hypotheses about lifeforms and the succession of species in coastal and shelf seas stem from the work of Margalef (1978) who suggested that variations in external energy in the form of nutrients and turbulence was the main factor controlling the temporal succession of phytoplankton. Margalef (1978) illustrated his ideas by plotting lifeforms on a surface defined by nutrient availability and turbulence (Figure 5.2). Diatoms are generally more abundant in waters of low vertical stability and high nutrients and dinoflagellates dominate stable water columns. Red tides of dinoflagellates occur when there is a nutrient supply to stratified waters. Figure 5.2 refers to k and r strategy. In ecology, r-strategists have a high potential growth rate and in general, the smaller diatoms fall into this category. These phytoplankters are able to succeed in situations where there is a transient supply of nutrients or improvement in the light climate that might be found in tidally stirred waters which intermittently stratify over a spring

- 152 - neap tidal cycle. Large, relatively slow growing diatoms and dinoflagellates that have a capacity to store nutrients are k-strategists.

Figure 5.2 Diatom and dinoflagellate lifeforms in state space defined by nutrient supply and turbulence and showing the succession from diatoms (r- strategy species: rapidly growing) to dinoflagellates (k- strategy species: slow growing able to store nutrients). Redrawn from Margalef (1978).

A number of studies have shown that Margalef’s general model is broadly applicable to the succession of phytoplankton species in shelf seas (Pingree et al. 1976, 1978; Holligan & Harbour 1977; Bowman et al. 1981; Jones et al. 1984) and in an extension to this general model, Smayda and Reynolds (2001) distinguished 9 different dinoflagellate lifeforms ordinated along a resource (nutrients and light) and energy (turbulent mixing) template. Jones and Gowen (1990) investigated the distribution of lifeforms in relation to turbulent mixing and irradiance regimes in shelf seas around the British Isles and found that diatoms were generally more abundant in waters of low vertical stability and steep irradiance gradients and dinoflagellates dominated stable water columns where irradiance gradients were small. Interestingly, Jones and Gowen (1990) found that unlike diatoms and dinoflagellates, microflagellates were not associated with a particular irradiance turbulence regime. In shelf seas that seasonally stratify there is therefore an expectation that diatoms as one lifeform will be replaced later in the year by dinoflagellates. This succession is not fixed and variations in mixing can retard and alter the pattern of

- 153 - succession (Smayda 1980). The dominance of dinoflagellates as a lifeform has been attributed to the ability of dinoflagellates to migrate vertically and remain in a stratified water column although Smayda (1997a) also considered the adverse effects of turbulence and sheer stress on dinoflagellate physiology and growth. In contrast, being non- motile, diatoms are more likely to sink out of the euphotic zone of a stratified water column but turbulent mixing will tend to keep non motile cells in suspension. Interestingly, recent studies have observed high abundance of diatoms in thin layers (see below) within the picnocline of stratified waters although whether the occurrence of such populations is the result of growth or the passive accumulation of cells as a result of reduced sinking is unclear (Velo-Suárez et al. 2008). In seasonally stratifying coastal and shelf seas therefore, the seasonal evolution of stratification provides a mechanism that selects for the dinoflagellate lifeform, many species of which are HAB species. This does not mean that the dominant species will necessarily be a HAB species, or that HABs will always occur. Dominance of one or more phytoplankters (harmful or benign) may require that their unique niche requirements are met by environmental conditions although the selection process may be stochastic, ‘a case of being in the right place at the right time’ (Smayda & Reynolds (2001).

5.4 The Interaction between Ecohydrodynamics and Ecophysiology

5.4.1 Introduction

We hypothesise that blooms of pelagic micro-algae (benign and harmful) result from the interaction between the ecophysiology of individual species and the ecohydrodynamic conditions in which the algae find themselves. Some coastal environments are more sensitive to the effects of nutrient enrichment and hence more at risk of eutrophication. Within such environments some of the species of harmful algae are more likely to form HABs. Other HAB species grow naturally and may be seeded into sensitive areas. In the following section we use the differences in ecohydrodynamic conditions that characterise different water bodies to explain why the nutrient enrichment → HAB hypothesis is supported in some water bodies at the spatial scales of Tolo Harbour (Hong Kong) and the Seto Inland Sea of Japan but not in other water bodies with similar spatial scales.

- 154 - 5.4.2 Small regions of restricted exchange

5.4.2.1 Introduction

As noted above, the residence time of water in coastal areas has an important bearing on the degree of enrichment and RREs whose dynamics depend on the rate of water exchange with the sea may be particularly sensitive to enrichment. Regions of restricted exchange in UK and Irish waters includes fjords, rias (flooded river valleys), other types of estuary, and coastal embayments and straits. These range in size from the small Scottish west coast sea lochs (Ross et al. 1994; Tett 1986), sea loughs in Northern Ireland (Ferreira et al. 2007) and Killary Harbour in Ireland to the larger Scottish firths (Clyde and Forth), river estuaries (such as Cork Harbour, the Thames and Humber estuaries), and more open coastal bays (Liverpool Bay, Dublin Bay) and large sheltered bays on the west coast of Ireland (e.g. Bantry Bay). In small RREs that have a residence time of a few days (a timescale similar to the typical doubling time of phytoplankton populations) phytoplankton do not remain within the RRE for a sufficient time for biomass to accumulate or for compositional changes to take place by way of seasonal succession. It is likely that in such RREs, the dynamics of phytoplankton populations simply reflect the situation in the adjacent coastal water. For example, Gowen et al. (1983) studied the small (2.39 x 106 m3 mid tide volume) sea loch Ardbhair on the west of Scotland over two years and observed that in 1981 summer biomass was low (≤ 1.0 mg m-1) but was up to 3.9 mg m-3 in 1982. The difference was attributed to ecohydrodynamic conditions in the source water. The water column was vertically mixed in 1981 and phytoplankton growth was assumed to be constrained by light (near surface nitrate 1.68 µM). In contrast, the water column was thermally stratified in 1982, near surface nitrate was ≤ 0.16 µM and there was a pronounced sub surface chlorophyll maximum. Jones and Gowen (1985) hypothesised that in moderately flushed (exchange rate ≈ 0.1 d-1) Scottish west coast sea lochs, phytoplankton may remain for several generation times but that the summer composition and biomass is dependent in part on the preconditioning of phytoplankton by ecohydrodynamic conditions in adjacent coastal water. In cases where phytoplankters entering an RRE are replete in nutrients, their growth may be uncoupled from ambient nutrient concentrations at least in the short term (Jones & Gowen 1985). This may render such moderately flushed RREs less sensitive to nutrient enrichment. With increasing residence time, ecohydrodynamic conditions within the RRE itself become increasingly important in determining phytoplankton dynamics. With a residence time in excess of ≈ 15 days there is the potential for biomass to accumulate as a result of in situ growth and compositional changes to occur in response to ecohydrodynamic conditions such as

- 155 - the formation of stratified waters (Tett et al. 1986; Rippeth et al. 1995). Although longer residence times generally means that phytoplankton dynamics within the RRE is largely independent of short term changes in the ecohydrodynamic conditions in adjacent coastal waters, this is not always the case. An example of a large scale event influencing phytoplankton within an RRE is provided by Belgrano et al. (1999) who found that the abundance of Dinophysis (acuminata, acuta and norvegica) in the Swedish Gullmar fjord was significantly related to the North Atlantic Oscillation Index.

5.4.2.2 Tolo and Victoria Harbour (Hong Kong)

The way in which flushing modifies the response of phytoplankton to anthropogenic nutrient enrichment and ultimately whether enrichment leads to an increase in HABs is exemplified by recent studies in coastal waters of Hong Kong (China) and in particular comparisons between Victoria and Tolo Harbours (Figure 3.6), two enriched RREs (Xu 2007). Tolo Harbour has been previously described in Part 3. It is a long (15 km) narrow sea inlet, 1 km wide at its entrance (Lam & Ho 1989). The surface area of the harbour is approximately 50 km2; depth ranges from 2-3 m in the inner region to 20 m in the outer part giving an overall mean depth of  12 m (Li et al. 2004); the volume is ≈ 0.6 km3. Lee et al. (2006) cite Choi and Lee (2004) as the source for estimates of residence times of between 14.4 and 38 days during the wet and dry seasons respectively. Victoria Harbour is a 12 km long tidal channel with a surface area of ≈ 50 km2 and depth of between 9 and 50 m (Yung et al. 1999). Assuming an average depth of 15 m would give a volume similar to that of Tolo Harbour. Flushing times are 1.5 – 2.5 days during the wet season and 5 – 7 days during the dry season (Kuang & Lee 2005). It is evident that compared to Tolo Harbour, red tides are much less frequent in Victoria Harbour. Between 1983 and 1998, a total of 288 red tides were recorded in Tolo Harbour compared to only 21 in Victoria Harbour (Yin 2003). Furthermore, a comparison between surface nutrient and chlorophyll data collected from the two harbours between 1991 and 2000 (Yin 2003), shows that compared to Tolo Harbour, the monthly mean concentration of near surface DAIN (NH4 + NO3 + NO2) was on average 29 % higher in Victoria Harbour but monthly mean chlorophyll concentrations were on average 35 % lower (Figure 5.3). Coastal waters to the west of Hong Kong are influenced by outflow from the Pearl River in the wet season (May to September) and the water column becomes stratified with surface to bottom differences in salinity (∆S) of up to 17 (Lee et al. 2006). Salinity stratification is weaker in Victoria Harbour (∆S < 7, Lee et al. 2006) but is probably sufficient to retain phytoplankton in the surface illuminated layer despite this being shallow. For Victoria harbour, Yung et al. (1999) give Secchi disc measurements of 1.72 – 1.78 m for the summer which gives a euphotic

- 156 - zone of ≈ 5 m37) and at this time, the mean daily total irradiance of the surface mixed layer exceeds 311 W m-2 (Xu, pers. comm.). This is similar to the mean irradiance of 0.03 gcal cm-2 min-1 = 21 W m-2 (assuming 1 W m-2 ≡ 0.0014 g cal cm-2 min-1) or 252 W m-2 for a day length

Figure 5.3 Monthly mean concentrations (µM) of dissolved inorganic nitrogen (DAIN as HN4, -3 NO3 and NO2) and chlorophyll (mg m ) in near surface waters of Tolo and Victoria Harbours between 1991 and 2000. A, DAIN; B, Chlorophyll. Redrawn from Figure 10 of Yin 2003).

35 A 30 25 M) 20 15

DAIN ( 10 5 Tolo Harbour Victoria Harbour 0 123456789101112 Month

25 B Tolo Harbour )

-3 20 Victoria Harbour

15

10

Chlorophyll (mg m Chlorophyll (mg 5

0 123456789101112 Month

of say 12 hours given by (Riley (1957) for the start of the spring bloom in coastal waters of the U.S. and the threshold (183 to 245 W m-2  12 to 16 W m-2 for a 15 hour day) daily light exposure required for the start of the phytoplankton production season in the Irish Sea derived by Gowen et al. (1995). During the summer therefore, conditions favour phytoplankton growth

37 For oceanic and optically clear coastal waters, Parsons et al. (1977) use a factor of 1.7 to estimate the attenuation coefficient of downwelling irradiance (Kd) from Secchi depth (Kd = 1.7/Secchi depth). In the absence of data on suspended particulate and coloured dissolved organic matter, we have used this e-kdh factor to estimate Kd in order to calculate euphotic zone depth using the equation Iz = I0 , assuming the bottom of the euphotic zone corresponds to the depth at which irradiance (Iz) is 1 % of surface irradiance (I0).

- 157 - (Figure 5.3B) but not all of the available DAIN is utilised and the monthly mean biomass is low (relative to Tolo Harbour) suggesting that the accumulation of a high biomass is constrained by rapid flushing of the harbour. During the dry season (October to April) the retention time is longer (5 – 7 days) but the water column is vertically mixed and the water column appears to be completely isohaline (Figure 2 of Yin & Harrison 2007) and isothermal (Figure 2 of Xu et al. 2008) as a result of low Pearl River outflow coupled with wind driven and tidal mixing. It is likely therefore that phytoplankton growth is constrained by light. The mean daily total surface mixed layer irradiance is low (< 153 W m-2, Xu pers comm.): near surface concentrations of DAIN are high and near surface concentrations of chlorophyll are low (Figure 5.3).

5.4.2.3 A comparison between Tolo Harbour and small enriched RREs in UK coastal waters

The comparison between Tolo and Victoria Harbours can be extended to consider small enriched RREs in UK waters. As examples we consider Carlingford Lough on the border between Northern Ireland and the Republic of Ireland and Loch Striven on the west coast of Scotland. Both of these water bodies have similar dimensions and flushing times as Tolo Harbour (Table 5.1) although Loch Striven is longer (21 km), narrower (on average 1 km) and deeper with a maximum depth of 69 m. All three RREs are enriched: Lam and Ho (1989) give - -1 the annual median nitrogen (as nitrate, NO3 ) for Tolo Harbour as 0.135 mg l (9.6 µM) in 1986 (see also Figure 5.3A); Ball et al. (1997) give a maximum (winter) concentration of up to 36 µM nitrate for Carlingford; Tett et al. (1986) give a winter maximum of 23.4 µM N (nitrate + nitrite) for Striven. Furthermore, based on estimates of the Equilibrium Concentration Enhancement38 (ECE), the 1997 level of enrichment in Carlingford Lough is comparable to that in Tolo Harbour in 1986 (Table 5.2). No estimate has been made of the ECE for Loch Striven because the freshwater catchment of Striven (98 km2) is only 6 times the surface area of the loch and most of the freshwater (and hence nutrients) entering the loch is derived from the enriched waters of the inner Firth of Clyde. Waters of Tolo Harbour thermally stratify and the water column in Loch Striven exhibits thermohaline stratification during the spring and summer period (Tett et al. 1986; Tett et al. 2001). Marshall and Orr (1930) report salinity layering at the time of the spring bloom (March) and Tett et al. (1986) give a salinity difference of between 5 – 7 for water at depths of 2 and 20 m and suggest that in Striven, vertical gradients in temperature can account for up to 83 % of the

38 ECE in units of µM was calculated from: N loading / (V · 1/D) where V is volume of the RRE and D is residence time in days. - 158 - density difference between 2 and 20 m, particularly during late spring and summer. In contrast, the waters within Carlingford Lough are generally more dynamic. On the basis of monthly sampling in 1992 and 1993, Ball et al. (1997) found that throughout the year, vertical gradients in salinity in the inner region of the lough were generally small (larger in winter associated with greater freshwater inflow) and concluded that the lough was vertically well mixed.

Table 5.1 Some physical characteristics of Tolo Harbour, Carlingford Lough, Loch Striven, The Seto Inland Sea of Japan and the eastern Irish Sea.

Location Volume Mean depth Residence time Temperature Water column (km3) (m) (d) (ºC) stability Tolo Harbour 0.6 12 15-30 13 – 32.0 thermal stratification (mean 24.3) Carlingford 0.5 8 14 - 26 6 – 16.6 Intermittent salinity Lough Mean (12.0)1 stratification Loch Striven 0.6 37 8 - 50 7 - 14 Seasonal thermo- haline Seto Inland 816 37 438 8 – 26 seasonal thermal Sea stratification Eastern Irish 480 15 40 – 480 5 - 17 intermittent salinity Sea stratification

Note: data sources are given in the text except for the mean annual temperature for Carlingford Lough which is from a study by E. Capuzzo.

Table 5.2 Estimates of the equilibrium concentration enhancement for selected water bodies.

Location Annual Residence (d) used to Equilibrium Loading data loading (t) calculate ECE time concentration source Enhancement (µM) Tolo Harbour 750 20 4.8 Lam & Ho (1989) (total N) Carlingford 1,311 10 5.6 Taylor et al. (1999) Lough (NO3 + NH4) Seto Inland Sea 181,000 438 19.0 Takeoka (1997) (total N) Eastern Irish 30,684 300 4.4 UK RID surveys Sea (total N)

Phytoplankton growth occurs throughout the year in Tolo Harbour and there is little evidence of periods of time when nutrients constrain growth (Figure 5.3). In contrast, there is a distinct seasonal cycle of phytoplankton growth and biomass in both Carlingford Lough and Loch Striven (Figure 5.4) with the production season restricted to the spring and summer

- 159 - months (March to October). The beginning of the production season is marked by a pronounced spring bloom in both of these RREs. Up to 8 mg chlorophyll m-3 was measured in the inner region of Carlingford Lough during May 1992 by Ball et al. (1997) although a spring bloom biomass of up to 20 mg m-3 has been recorded (E. Capuzzo, pers comm.), perhaps as a result of higher frequency (weekly) sampling and therefore better resolution of the spring bloom. Loch -3 Striven supports a larger spring bloom (89 mg chlorophyll m in 1980, Tett et al. 1986).

Figure 5.4 The seasonal cycle of phytoplankton biomass as chlorophyll (mg m-3) in Carlingford Lough and Loch Striven. A, the inner region of Carlingford Lough, redrawn from Douglas (1992) and Ball et al. (1997); B, Loch Striven, redrawn from Tett et al. (1986). Note the difference in the scale of the Y axis. Carlingford Lough data from 1990 and 1992; Loch Striven data from 1980.

14 A Data from Douglas 12 Data from Ball et al

) -3 10

8

6

4 Chlorophyll (mg m (mg Chlorophyll 2 0 01-Jan 22-Feb 15-Apr 06-Jun 28-Jul 18-Sep 09-Nov 31-Dec

90 B 23rd September 80 181 mg m-3

) 70 -3 60 50 40 30 Chlorophyll (mgm 20

10 0 01-Jan 20-Feb 10-Apr 30-May 19-Jul 07-Sep 27-Oct 16-Dec

Summer biomass in the inner region of Carlingford is variable with short term peaks of up to 12 mg m-3. The occurrence of these peaks is suggestive of a pulsed supply of nutrients although the source is somewhat unclear. Resupply from bottom water is unlikely because unlike some of the larger Scottish sea lochs in which nutrient rich bottom water can be entrained - 160 - into surface waters (see for example, Tett et al. 1986), Carlingford is relatively shallow (maximum depth 35 m) and the water column only intermittently stratified. Douglas (1992) does not discuss the nutrient supply which fuelled the peaks in July and August 1990, but of several possible sources (resupply from coastal water, anthropogenic sources and remineralisation), Ball et al. (1997) discounted the natural input of nutrients from coastal waters and anthropogenic nutrients (riverine and domestic) as being insufficient and were of the opinion that nitrogen remineralisation within the lough was a possible source. More recently, a peak in summer chlorophyll was shown to follow an increase in river flow suggesting that anthropogenic nutrients play a role (E. Capuzzo pers com). During the summer, phytoplankton biomass in Loch Striven appears to be equally variable. Tett et al (1986) reported a sequence of blooms in the loch beginning with the spring bloom and followed by three other blooms (see also Tett et al. 2001) and concluded that the supply of nitrogen from the inner Firth of Clyde fuelled summer blooms, although the periodic resupply of nutrients to near surface waters from nutrient rich ( 16 µM N, Tett et al. 1986) bottom water cannot be ruled out. The phytoplankton in Carlingford Lough is dominated by diatoms throughout much of the year (Ball et al. 1997; Douglas 1992) as is the case in Loch Striven (Marshall & Orr, 1927) and in Tolo Harbour (Yung et al. 1997) although the latter experiences regular harmful blooms. Lam and Ho (1989) reported blooms of the dinoflagellates Noctiluca scintillans, Prorocentrum triestinum, P. dentatum and P. sigmoides, an unidentified Gymnodinioid and a wide range of microflagellates and Yin (2003) included the dinoflagellates Gonyaulax polygramma and Prorocentrum minimum together with the diatom Skeletonema costatum and photosynthetic ciliate Myrionecta rubra amongst the six most frequently occurring red tide species in the harbour. Most of the problems caused by blooms of these species have been associated with deoxygenation rather than biotoxins (Holmes & Lam 1985) indicating that the problems are associated with large biomass blooms. Summer blooms in Carlingford appear to be dominated by diatoms. In 1992, the dominant species were Leptocylindrus danicus and Chaetoceros socialis in June and Asterionella japonica and Rhizosolenia hebetata in July (Ball et al. 1997). There is no indication that these summer blooms were HABs i.e. had a negative impact on ecosystem goods and services. In 1980, summer blooms in Loch Striven were composed of mixtures of diatoms and dinoflagellates (Tett et al. 1986) although it is evident that in previous years, some summer blooms (comprised of microflagellates, see Part 2) were associated with mortalities of farmed fish (Tett 1980) and were therefore HABs.

- 161 - One obvious difference between Tolo Harbour and the two UK RREs is water temperature. The annual mean surface water temperature in the inner region of Tolo Harbour (2000 – 2006) was 24.3º C with a range from 13 to – 32.0º C and compares with a 2007 mean of 12.0 º C and range from 6 to 16.6 º C for Carlingford (E. Capuzzo, pers. com). Striven probably has a similar seasonal temperature range. In March the near surface water temperature is  7° C (Marshall & Orr 1930) and during the summer, near surface temperature is ≈ 14° C (Tett et al. 1986). While higher temperature might be expected to influence the growth rate of individual phytoplankters, it is unlikely to be the reason for the higher occurrence of red tides in Tolo Harbour. Maximum solar radiation occurs at the equator and decreases with increasing latitude and in general this will result in higher levels of subsurface irradiance in tropical and subtropical coastal waters compared to cool temperate waters. However this is balanced by longer day length during the spring and summer in northern latitudes. As a consequence, the differences in mean sub-surface daily irradiance in Tolo Harbour and the two UK RREs are likely to be small and insufficient to account for the greater frequency of HABs in Tolo Harbour. Furthermore, the light climate is clearly adequate to support phytoplankton growth in both Carlingford and Striven as evidenced by the recurrent annual cycle of production and biomass. Ecohydrodynamically, Tolo Harbour and Loch Striven appear to share similar features that contrast with those of Carlingford Lough. In our opinion, the key difference between these two RREs and Carlingford is the persistence of water column stratification. The water column in Tolo Harbour thermally stratifies and Lam and Ho (1989) suggested that the north easterly winds aid the accumulation of biomass in the inner harbour and that slow currents and calm conditions provide a stable environment for blooms to develop. In Loch Striven, thermohaline stratification retains phytoplankton in the near surface illuminated layer, and a supply of nutrients from the inner Firth of Clyde (or perhaps from nutrient rich bottom water) supports summer blooms. In contrast, stratification in Carlingford is intermittent and short lived and the lough appears to be physically more dynamic. Intermittent stratification would be expected to favour diatoms as the dominant lifeform and it is likely that the periods of stratification are too short for succession to a dinoflagellate lifeform. Furthermore, that the summer biomass peaks were dominated by diatoms is consistent with the suggestion of Smayda and Reynolds (2001), that once physical conditions have enabled a particular lifeform to become dominant, it is species of this lifeform that are best placed to respond to favourable conditions: in this case an increase in nutrient supply. We therefore conclude that it is the dynamic physical conditions within Carlingford Lough which prevent the development of dinoflagellate and flagellate HABs. This does not, however, explain the occurrence of harmful blooms caused by the diatom

- 162 - Skeletonema costatum in Tolo Harbour and the fact that no harmful effects were attributed to the summer diatom blooms in Carlingford reported by Douglas (1992) and Ball et al. (1997). It is unlikely that the S. costatum that causes HABs in coastal waters of China is the same species that forms an important component of the spring bloom in coastal waters of north west Europe (e.g. Marshall & Orr 1930; McKinney et al. 1997). More importantly, the size of the summer blooms in Carlingford may be limited by the level of nutrient loading. Although Carlingford is enriched, there is a clear difference between winter and summer loading. Ball et al. (1997) report average winter and summer flows as 2.85 and 0.46 m3 s-1 respectively. According to Taylor et al. (1999) during the summer the mean daily input of nitrogen is 68.4 x 103 mol d-1 (compared to a winter input of 398.4 x 103 mol d-1) and would only add ≈ 1.0 µM assuming an upper region volume of 57 x 106 m-3 and a residence time of 1 day (based on the ratio of lough volume to tidal prism). This would support the contention of Ball et al. (1997) that there is insufficient anthropogenic N to fuel summer blooms. However, such rapid flushing assumes that none of the water leaving the upper region of the lough during the ebb tide returns on the following flood tide and this is unlikely to hold true. More recently, Ferreira et al. (2007) estimated the residence time of the whole lough as 14 – 26 days (based on a hydrodynamic model [Delft3]) which is longer than the 3 days suggested by Taylor et al. (1999). Furthermore, using river inflow and salinity39 to estimate the exchange time gives a value of ≈ 8 days and an ECE concentration of 7 µM N. These simple calculations suggest that during the summer, average land based sources of nitrogen could fuel summer blooms of up to ≈ 10 mg chlorophyll m-3 assuming a yield of chlorophyll from nitrogen of 1.05 (Gowen et al. 1992). Such blooms are unlikely to cause oxygen depletion in the lough given the dynamic nature of the water column.

5.4.3 Regional Seas

The second of two examples presented in Part 3 and which supports the nutrient enrichment → HAB hypothesis was the Seto Inland Sea of Japan, a large semi enclosed regional sea (Table 5.1). There is no direct equivalent in UK and Irish waters, although the Irish Sea (Figure 5.5) is a partially enclosed regional sea but has a volume of 2540 km3, some three times that of the Seto Inland Sea. Furthermore, the deep (100 m) seasonally stratifying waters of the western Irish Sea are only moderately enriched by ≈ 2 – 3 µM N relative to historical concentrations (Gowen et al. 2008). The eastern Irish Sea is enriched (Gowen et al. 2008), relatively shallow and has a volume that is approximately half that of the Seto Inland Sea. Furthermore, the distribution of near surface salinity in the Irish Sea suggested that there is limited exchange between its eastern

3 Dilution (D) = (R · So)/ (V · (So – S)), where R is river inflow, V, volume of RRE and So and S are the mean salinities of sea water flowing into and out of the RRE respectively. - 163 - and western regions (Gowen et al. 2002) and the two can be considered as hydrographically distinct regions at least during the phytoplankton growing season. For these reasons we have compared the Seto Inland Sea with the eastern Irish Sea.

Figure 5.5 A map of the Irish Sea showing locations mentioned in the text. The dashed line shows the approximate position of the western Irish Sea tidal mixing front. IOM, Isle of Man. The numbers 1 to 6 show the positions of stations used by Bowden and Sharef El Din (1966) and LB is the station used by Gowen et al. 2000.

56.0 Malin Shelf Islay

55.5 Firth Scotland of Clyde 55.0

North Solway Firth Belfast Channel 54.5 North eastern Cumbria Irish Sea IOM 54.0 Western South eastern Irish Irish Sea Sea 6 R Ribble Liverpool 2 Bay 51 53.5 4LB 3

Latitude R Mersey R Dee 53.0

St George's Channel 52.5 Wales

52.0

51.5 Celtic Sea Bristol Channel

51.0 7.0 6.5 6.0 5.5 5.0 4.5 4.0 3.5 3.0 2.5 2.0

Longitude

Details of the Seto Inland Sea were presented in Part 3 (and see Table 5.1). In brief, the Seto Inland Sea has a surface area of 21,827 km2, mean depth of 37 m and volume of 816 km3. Residual currents are generally weak in the large bays causing closed circulation within each and restricting exchange between them. Freshwater inflow is considered low (44 km3 year-1) and estuarine circulation correspondingly weak except in Osaka and Hiroshima Bays where larger rivers discharge. Frontal boundaries between water bodies are a key feature in the Seto Inland - 164 - Sea (see for example, Yanagi & Yoshikawa 1987). In his review, Takeoka (2002) discussed the occurrence of different types of front. Thermohaline fronts form mostly during winter between cold less saline water and warmer more saline oceanic water. Tidal mixing fronts develop in summer between deep vertically mixed water of the narrow channels between the islands and shallower stratified bay waters and between vertically mixed inshore water and more offshore stratified water. Estuarine and shelf fronts also develop within the Seto Inland Sea. These fronts play an important role in biological processes. For example, Takeoka et al. (1993) observed a pronounced chlorophyll maximum (≈ 7.5 mg m-3) at the tidal front formed in Iua-Nada around the Hayasui Strait. It is interesting however, that Takeoka et al. (1993) also observed what they considered to be the influence of freshwater at the front:

“The Chl a peak in Fig.5 appears not in the frontal region but in the surface layer in the stratified region. This is probably due to the nutrients brought by the river.”

These observations would support our earlier contention that nutrient inputs to near surface illuminated layers (either natural, during upwelling or anthropogenic) are likely to stimulate algal blooms unless planktonic animals or benthic filter-feeders consume the increased algal production. The physical accumulation of biomass may also be an important mechanism for red tide formation in the Inland Sea and on the basis of an observational and modelling study Yanagi et al. (1995) concluded that:

“these results suggest that the night intake of ammonium by G. mikimotoi [Karenia mikimotoi] and the physical accumulation of cells by current play very important role in the formation of red tides of G mikimotoi at Suo- Nada.”

The eastern Irish Sea has a volume of approximately 480 km3 and a residence time of 40 to 480 days (Dickson & Boelens 1988). The mean depth is  15 m and of the total freshwater inflow into the Irish Sea (31 km3) some 80 % (24.9 km3 year-1) flows into the eastern Irish Sea (Bowden 1955). Flow through the Irish Sea is generally considered to be northwards (Bassett 1909) with the bulk of the outflow along the Scottish side of the North Channel. The eastern Irish Sea can be divided into a southern and northern area (Figure 5.5). The southern area which includes Liverpool Bay is much influenced by freshwater inflow and is therefore a ‘Region Of Freshwater Influence’ or ROFI, meaning that there is tidal straining of the horizontal salinity gradient; sporadic lenses of fresher water that are moved by wind and mixed away when stirring increases. The area is physically dynamic: horizontal dispersion occurs by wind driven, residual and tidal flows. Abdullah and Royle (1973) estimated horizontal dispersion rates of 175 - 204 m-2 S-1 - 165 - in December 1970 and 89 – 100 m-2 S-1 for March 1971. Bowden and Sharef El Din (1966) given current speeds of between 38 and 51 cm S-1, (see Figure 5.5 for station positions). Vertical stirring is brought about by the tide, wind and secondary circulations. However, there is also considerable input by freshwater. As a consequence there is a rapidly changing vertical state that can be mixed or stratified depending on local outcome of opposing tendencies. Bowden and Sharef El Din (1966) give salinity gradients of 0.08 to 0.21 (0.78 at one sampling station). At a sampling station in Liverpool Bay (LB in Figure 5.5), Gowen et al. (2000) calculated surface to bottom differences in temperature and salinity of 0.0 – 0.2º C and 0.1 –0.14 respectively in spring and 0.0 – 0.2º C for temperature and 0.01 – 0.22 for salinity in summer. A salinity front marks the boundary between more saline offshore eastern Irish Sea water and lower salinity inshore water but this does not appear to be a persistent feature (Foster et al. 1984). Further offshore, the vertically mixed waters of the eastern Irish Sea are separated from deeper thermally stratified waters of the western Irish Sea by the western Irish Sea front (Simpson & Hunter 1974). In contrast to other tidal mixing fronts that support high summer biomass and surface blooms of dinoflagellates (e.g. Karenia mikimotoi at the Ushant front in the western English Channel, Pingree et al. 1975) this front appears to have little influence on phytoplankton biomass (Richardson et al. 1985) and we are unaware of any reports of large surface blooms at the western Irish Sea front. The northern region of the eastern Irish Sea is characterised by weak thermohaline stratification in summer and occasional haline stratification in the vicinity of the Solway Firth in winter (Kennington et al. 1999). Weak, possibly intermittent, frontal regions may separate the more tidally stirred waters of the southern region from those to the north. Loadings to the Seto Inland Sea and the eastern Irish Sea (Table 5.2) show that both are influenced by anthropogenic nutrient enrichment, although estimates of the equilibrium enhancement concentration suggest that the level of enrichment in the Seto Inland Sea was substantially greater than in the eastern Irish Sea. A number of studies show that in waters of Liverpool Bay (see Gowen et al. 2002 and references cited therein) and inshore waters of the north eastern Irish Sea (Kennington et al. 2002) winter concentrations of nitrate (+ nitrite) typically reach 30 µM. Offshore eastern Irish Sea waters are only moderately enriched (≈ 10 µM DAIN, Kennington et al. 2002) compared to oceanic waters (≈ 7 µM nitrate + nitrite) and historical concentrations of 5-6 µM in the western Irish Sea (Gowen et al. 2002). The phytoplankton in both regional seas exhibit seasonality in growth and it has been reported that nutrient enrichment has elevated primary production in the Seto Inland Sea (Hashimoto et al. 1997) and Liverpool Bay (Gowen et al. 2000). In contrast to the Seto Inland Sea (where there are about 100 red tides per year, Figure 3.16), the eastern Irish Sea does not

- 166 - suffer from the same intensity of HABs as the Seto Inland Sea. As discussed in Part 2, HABs occur infrequently in the eastern Irish Sea and we are unaware of any recent reports of large (high biomass and/or geographical distribution) HABs or toxic events in the region. Earlier in this part of the report, we suggested that the introduction of nutrients (natural or from anthropogenic sources) to the surface layers of stratified waters is likely to promote the formation of phytoplankton blooms. This would appear to be the case in the Seto Inland Sea, within which, the dominant ecohydrodynamic characteristics are the seasonal development of thermal stratification in the large bays; the formation of frontal boundaries; currents that can cause biomass to accumulate. The south eastern part of the eastern Irish Sea is physically more dynamic and this may restrict the development of HABs despite anthropogenic nutrient enrichment. Gowen et al. (2000) considered that low amounts of sediment phyto-pigments and low concentrations of sediment pore water Si were suggestive of a low input of phytoplankton carbon to the sediments in Liverpool Bay and concluded that much of the phytoplankton biomass was lost from the area by advection. Summer growth of phytoplankton may be influenced by the effect of the spring – neap tidal cycle on vertical mixing and the sub-surface light climate (Cefas, unpubl. data) as appears to be the case in the Bay of Brest in France (Le Pape et al. 1996) and Southampton water (Crawford et al. 1997). Finally, Gowen et al. (2000) concluded that enrichment and shifts in nutrient ratios in Liverpool Bay did not favour flagellate growth over that of diatoms, suggesting that the intermittent stratification favours diatoms as the dominant lifeform of pelagic primary producer and limits the seasonal succession to dinoflagellates. The weak but persistent seasonal stratification in the north eastern Irish Sea might be expected to provide more favourable conditions for phytoplankton growth and a seasonal succession from diatoms to dinoflagellates. Phytoplankton data from this area of the Irish Sea are limited although Kennington et al. (1999) noted that during July 1996, on average, the phytoplankton was dominated by diatoms. It seems likely that offshore in the north eastern Irish Sea, the low level of anthropogenic nutrient enrichment constrains the development of HABs.

5.4.4 Summary

In our discussion of ecohydrodynamics in Section 5.2, we argued that rates of lateral exchange, mixing, or dispersion within and between water bodies and the strength of vertical mixing and its consequences for the illumination experienced by primary producers are key factors in determining whether anthropogenic nutrient enrichment of a water body is likely to stimulate HABs. In our opinion, the characteristic ecohydrodynamic features of the water bodies discussed above and differences in the occurrence of HABs in them (summarised in Table 5.3)

- 167 - support this view at the scale of both small RREs (Tolo Harbour and to a lesser extent Loch Striven) and regional seas (the Seto Inland Sea). Tolo Harbour, Loch Striven and the Seto Inland Sea show general symptoms of eutrophication including increased biomass and increased frequency of high-biomass blooms with sometimes harmful consequences. Victoria Harbour, Carlingford Lough and the eastern Irish Sea do not exhibit the symptoms of eutrophication and this is more because their hydrodynamic characteristics (i.e. rapid flushing in Victoria Harbour and tidal stirring in Carlingford Lough and the south eastern Irish Sea) rather than nutrient loading, do not favour the development of large biomass HABs. With respect to low biomass blooms of toxin producing species, the analysis of UK and Irish data shows that the occurrence and abundance of these species is not determined by nutrient enrichment.

Table 5.3 The ecohydrodynamic characteristics of selected water bodies discussed in the text and the occurrence of HABs and benign micro-algal blooms.

Regions of Restricted Characteristic ecohydrodynamic Harmful and benign blooms Exchange features Tolo Harbour Persistent thermal stratification and High frequency of HABs slowly flushed. Victoria Harbour Short seasonal stratification and Low frequency of HABs rapidly flushed. Carlingford Lough Tidally stirred, with infrequent Low frequency of benign summer thermo-haline stratification and blooms moderately flushed.

Loch Striven Seasonal thermo-haline stratification Benign summer blooms and and moderately flushed. occasional HABs Regional Seas Seto Inland Sea Thermal stratification and numerous High frequency of HABs frontal boundaries Eastern Irish Sea Region of freshwater influence, Infrequent occurrence of HABs tidally stirred and intermittent thermohaline stratification

5.5 The Distribution of HAB Species in UK and Irish Coastal Waters

5.5.1 Introduction

There are clear patterns in the geographical distribution of some of the species under examination in this study (Figure 4.7) and in this section we attempt to explain the observed distributions in terms of the intersection of ecophysiology and ecohydrodynamics. It is apparent, however, that the highest abundance of some of the species under discussion is found in those

- 168 - coastal waters most extensively used for aquaculture and we briefly consider this issue at the end of this section.

5.5.2 Ecohydrodynamic conditions in UK and Irish coastal waters

Coastal waters around the British Isles and Ireland provide a range of ecohydrodynamic conditions and hence a variety of environments for phytoplankton growth. In addition to being rapidly flushed, many of the larger estuaries in the southeast of the UK such as the Thames, Wash and Humber are turbid and summer growth of phytoplankton is constrained by light. The Humber for example, is the largest estuarine system in England (catchment 24,240 km2). It is shallow (< 5 to ≈ 20 m), has a tidal range of 3.5 – 6 m and the high (suspended load 0.2 – 2 g L-1 in the turbidity maximum) limits primary production within the estuary (Jickels et al. 2000). Sanders et al. (2001) studied the effect of nutrient enrichment on phytoplankton at an inshore site within the Thames estuary and concluded that low summer biomass despite high levels of N and P (10 µM nitrate and 1 µM phosphate) was due to high turbidity which reduced sub-surface irradiance to a level sufficient for the growth of diatoms but not flagellates and that growth of diatoms was limited by silicate limitation. Other coastal regions such as the eastern Irish Sea, discussed above are less turbid but are tidally stirred regions of freshwater influence and many of the RREs on the west coast of Ireland and Scotland (exemplified by Loch Striven described above) are deep, sheltered and exhibit strong thermo-haline stratification (Wood et al. 1973; Tett et al. 1986) Deeper and more open coastal waters, in which tidal flows are weak, undergo seasonal thermal or thermo-haline stratification. The Celtic Sea (Fasham et al. 1983), Northern North Sea (Tett et al. 1993; Mills et al. 1994; Lee et al. 2002), western Irish Sea (Gowen et al. 1995; Horsburgh et al. 2000) and Sound of Jura (Jones et al 1984) provide examples. A key physical feature of such regions are the tidal mixing fronts (e.g. Simpson & Hunter 1974; Pingree et al. 1975) that develop at the interface between vertically mixed seasonally stratifying waters (Figure 5.6). Some of these frontal regions are important sites of enhanced biological production and are often regions where dinoflagellate populations develop. For example, Pingree et al. (1975) measured chlorophyll concentrations of up to 100 mg m-3 at the Ushant front in the western English Channel, and Karenia mikimotoi (1.7 106 cells L-1) associated with a 34 mg m-3 patch of chlorophyll. This suggests that ecohydrodynamic conditions at these fronts are particularly suitable for growth or that frontal conditions cause aggregation. Within some of these stratified regions such as the western Irish Sea, the development of bottom density fronts results in the formation of a cyclonic gyre of near surface water (Hill et al. 1994) that may act as retention areas for plankton (White et al. 1988).

- 169 -

Figure 5.6 A map showing the generalised flow of water around British Isles and Ireland and in the North Sea. The approximate locations of tidal mixing fronts are shown by dashed red lines. ICC, Irish coastal current; SCC, Scottish coastal current; NCC, Norwegian coastal current. The continuous arrows show the main persistent flows, with green colour indicating significant freshwater content. Dashed lines indicate flows that are likely stronger in winter (i.e. in absence of fronts). The dotted line is the subsurface inflow from the Atlantic which combines with the other oceanic inflows and the outflow from the Baltic to make the NCC.

There is evidence of a salinity front, the Irish Shelf front, that separates Irish coastal water from water of more oceanic characteristics to the south west (see Raine & McMahon 1998 and references cited therein) and north west (Bowyer & Ward 1996) of Ireland. Ellett (1979) noted that to the west of Islay there was a pronounced oceanic front (where Atlantic water flowing onto the shelf to the north of Ireland meets northward flowing Irish Sea/ Clyde water) and suggested that there was an oceanic front to the west of the Outer Hebrides. There is therefore the possibility of a shelf front extending from the south west of Ireland north along the western shelf of Ireland and Scotland.

- 170 - According to Pingree and Le Cann (1989) there is a near surface north westerly flow of coastal water along the Amorican shelf (northern Bay of Biscay) which moves across the mouth of the English Channel to the Isles of Scilly and along the south west of Ireland. The Irish coastal current (Figure 5.6) that flows along the south coast of Ireland and northwards around the west coast of Ireland could aid the transport of plankton. Raine and McMahon (1998) suggested that dinoflagellate populations could be transported along the southern coast of Ireland from the Celtic Sea at ≈ 10 – 15 km d-1 and the occurrence of warm temperate phytoplankters (e.g. Ceratium azoricum, C. arietinum and C. hexacanthum) off the north west of Ireland (Gowen et al. 1998) is also suggestive of transport in a coastal current. Ellett (1979) and see (Ellett & Edwards 1983; Simpson & Hill 1986; Hill et al. 1997) suggested that there is a general northward transport (coastal current) of water in coastal waters to the west of Scotland which is diverted around the Outer Hebrides with some flow between the Outer Hebrides and the mainland. Whether the Irish coastal current is continuous with the Scottish coastal current is unclear but such a coastal current could aid the distribution of HAB species along the western seaboard of Ireland and Scotland.

5.5.3 Species of Alexandrium

The current distribution of Alexandrium spp. (Figure 4.7) shows these species are generally more abundant in: shallow estuaries of south west England; western waters of Ireland; the west and east coast of Scotland and in waters around the Orkney and Shetland Islands. Species of Alexandrium are least abundant in coastal regions like the Thames estuary and coastal waters of the Irish Sea. As noted in Part 2, both toxic and non toxic strains of Alexandrium occur in waters along the SW coast of England and provides an explanation for the low level of toxicity associated with these blooms. In Scotland the presence of both the PSP toxin producing North American (group I) and non toxic Western European (group III) (Lilly et al. 2007) has been demonstrated, along with the presence of A. ostenfeldii (Medlin et al. 1998; Higman et al. 2001; John et al. 2003; Collins et al. 2009). In a study of the occurrence of PSP toxicity in coastal waters of the north east of England, Joint et al. (1997) considered whether the 1990 A. tamarense bloom that caused widespread toxicity, originated from the Firth of Forth where there is a region of high cyst abundance. On the basis of observations and a transport model, Joint et al. (1997) concluded that there was little evidence to support this hypothesis and that it was likely that the 1990 bloom had several offshore sources.

- 171 - More recently, Brown et al. (2001) suggested that there is a strong and persistent seasonal baroclinic (density driven) southward coastal transport with typical flows of 0.07 m s-1 that could transport cysts and vegetative cells from the Firth of Forth. Nevertheless, Brown et al. (2001) also discounted the Firth of Forth as the source of A. tamarense blooms but argued that each year, the transport of cysts and vegetative cells from the Firth of Forth could maintain cyst populations in sediments along the north east coast of England. It is interesting that although the north east of England has historically been a region where PSP events have occurred (Ayres et al. 1975) recent data (1998 to 2007) show that the abundance of Alexandrium spp. is lower in this region compared to western Scottish waters. A key aspect of the life cycle of A. tamarense and other species of Alexandrium is the formation of cysts that ‘over winter’ in bottom sediments providing a mechanism for maintaining the population of vegetative cells from one year to the next. A depositional environment which allows cysts to settle to the sea bed rather than being dispersed is considered a key ecohydrodynamic feature of areas where these species occur. Anderson et al. (2005) estimated that in the Gulf of Maine, the region of coastal sediment containing cysts of Alexandrium fundyense extended over ≈ 500 km and the abundance of cysts in sediments was between 2 and 20 x 106 m-2. The relatively high abundance of Alexandrium spp. in RREs such as the Fal estuary (south west coast of England), Cork harbour (south west coast of Ireland) and Belfast Lough (Northern Ireland) may reflect the presence of a depositional area which allows the accumulation of cysts in the sediment. Cysts of Alexandrium tamarense and A. minutum have been identified in sediments from Belfast Lough (Tylor et al. 1995) and the Fal estuary (Blanco et al. 2009) respectively.

5.5.4 Species of Dinophysis

These species are widely distributed throughout UK and Irish coastal waters and Figure 4.7 shows that between 2002 and 2007 these phytoplankters were generally more abundant along the south and west coast of Ireland, the west and south east of Scotland and the north east and south west of England. Much remains unknown about the ecophysiology of these species and it is only recently that Dinophysis has been successfully grown in culture (Park et al. 2006) and aspects of its complex nutrition elucidated. These phytoplankters are not known to produce cysts. There is currently much interest in the ecological importance of the micro scale features in stratified waters and associated patches of plankton referred to as ‘thin layers’ (Dekshenieks et al. 2001). According to Velo-Suárez et al. (2008), thin layers are structures that have characteristic physical, chemical and biological features that are different from the water

- 172 - immediately above and below. The thickness of thin layers varies from a few centimetres to meters; the layer can extend over several kilometres and persist for periods of days. Thin layers are a property of stratified waters in which the vertical variation in horizontal flow (i.e. shear) results in the horizontal distribution of any property including plankton (Gentian et al. 2005). Layer thickness is dependant on the degree of turbulence with very thin layers observed in strongly stratified waters, thicker layers occur in more weakly stratified waters and it follows that thin layers are absent from tidally stirred waters. With respect to the dynamics of populations of HAB species, interest in thin layers stems from observations that a number of HAB species have been observed to reach a much higher abundance in thin layers compared to the surrounding water. There is also the potential for such populations to be transported from offshore stratified waters into near-shore waters. For example, Gentian et al. (2005) reported that during a study off the south west of Ireland in 1992, a thin layer of phytoplankton was dominated by Dinophysis acuminata that reached an abundance of up to 0.124 x 106 cells L-1. Different species may dominate different layers (McManus et al. 2003) resulting in the fine scale vertical distribution of species down the water column. It is evident that much remains unknown regarding the processes controlling the formation of harmful populations of HAB species in thin layers and there are likely to be both ecological advantages and disadvantages for a HAB species population in a thin layer (Gentian et al. 2005). Nevertheless, the adaptation of dinoflagellates such as Dinophysis spp. to stratified waters and the occurrence of large populations of these phytoplankters in thin layers of stratified waters, provides one explanation for the greater abundance of Dinophysis spp. in seasonally stratifying coastal waters and the larger RREs (in which thermo-haline stratification is a key ecohydrodynamic feature) of western coastal waters of Ireland and Scotland. It is important to note however that species of Dinophysis are found in low abundance (typically a few hundreds of cells L-1) in tidally stirred waters such as the eastern Irish Sea (Figure 4.7). At the present time it is unclear whether the occurrence of such populations is due to growth in these mixed waters or the transport of cells from populations in stratified waters.

5.5.5 The genus Pseudo-nitzschia

Species of Pseudo-nitzschia were widespread throughout UK and Irish coastal waters between 2002 and 2007 (Figure 4.7) and on occasion exceeded 106 cells L-1. There is no obvious pattern to the geographical distribution of these phytoplankters although mean and maximum abundance was lower in south eastern coastal waters of Ireland and England. Being diatoms, species of Pseudo-nitzschia have a requirement for silicate although the fact that these species

- 173 - grow well in microcosms (see for example Jones et al. 1978) suggest they have no specific nutritional requirements although the presence of physical surfaces (walls of the culture vessel) may also play a role. The utilisation of organic N by Pseudo-nitzschia has recently been demonstrated (Loureiro et al. 2009) and Pan et al. (1996) suggest that species of Pseudo- nitzschia may be adapted to a wide range of silicate concentrations giving then a competitive advantage over other diatoms although this is based on a limited number of observations. Examination of Pseudo-nitzschia isolated from Scottish waters shows that the growth rate of different species is influenced by photoperiod (Fehling et al. 2005). Regional diversity can be seen in the distribution of Pseudo-nitzschia species with P. australis and P. seriata dominating Psuedo-nitzschia blooms at sites along the Atlantic and Northern North sea coasts (Fehling et al, 2006, Marine Scotland, unpubl data) while P. pungens and P. multiseries dominate populations in the Southern North Sea (Evans et al. 2005, Casteleyn et al. 2008). Interestingly, of all of the coastal waters of the UK and Ireland, species of Pseudo- nitzschia were less abundant in the Thames estuary region. Data in Sanders et al. (2001) show this region to be enriched (winter nitrate ≈ 34 µM) but summer biomass is low (≈ 2 mg chlorophyll m-3) and Sanders et al. (2001) discuss whether low concentrations of silicate (1 – 2 µM) limits the growth of diatoms during the summer. Recent studies have shown species of Pseudo-nitzschia to occur in thin layers (Rines et al. 2002; Velo-Suárez et al. 2008). Rines et al. (2002) suggest that the complex hydrographic regimes and physical forcing that brought about the formation of thin layers of Pseudo-nitzschia fraudulenta in East Sound (a fjord on the coast of Orcas Island, Washington U.S.) are likely to occur in other fjordic coastlines (e.g. Chile, Norway, Scotland) and that species of Pseudo- nitzschia may well occur in undetected thin layers in such waters. Whether this would result in a greater abundance of Pseudo-nitzschia spp. in stratified waters compared to tidally stirred waters is unclear. The current distribution of Pseudo-nitzschia spp. in these waters suggests that these species are not more abundant in stratified western waters but as pointed out by Rines et al. (2002) the traditional sampling methods (on which the current distribution is based) may not detect Pseudo-nitzschia spp. present in thin layers. The processes controlling the current distribution and abundance of Pseudo-nitzschia spp. in UK and Irish waters are unclear. The potential for large populations of Pseudo-nitzschia spp. to occur in thin layers has implications for monitoring for the presence of these species and relating episodes of ASP toxicity to the occurrence of Pseudo-nitzschia spp. populations. In addition to the likelihood that thin layers of Pseudo-nitzschia spp. would be undetected using traditional sampling methods, Rines et al. (2002) suggest that: large populations of toxin producing species of Pseudo- nitzschia in thin layers, have the potential to directly affect the toxicity of shellfish; and there is

- 174 - the possibility of such thin layers being advected at depth (and undetected) from one coastal area to another. The Irish and Scottish coastal currents could provide a mechanism for the transport of populations of Pseudo nitzschia spp. in thin layers.

5.5.6 Karenia mikimotoi

Between 2000 and 2004, K. mikimotoi was found in low abundance throughout coastal waters of the UK and Ireland (Figure 4.7) but was more abundance in waters off the south west of England, the southern and western coasts of Ireland and in coastal waters around Scotland. The first recorded occurrence of Karenia mikimotoi (an ichthyotoxic dinoflagellate) in European waters was in 1966 (Braarud & Heimdal 1970) and there have been a number of blooms throughout northern European waters since that time. Smayda (1990) plotted the distribution of K. mikimotoi in NW Europe (Figure 5.7) and observed that it was largely absent from the southern North Sea and eastern English Channel. Although the apparent absence of Karenia mikimotoi from some coastal regions (e.g. south east England) may have been due to limited monitoring at that time, Smayda (1990) was of the opinion that the absence of K. mikimotoi from continental coastal waters of the southern North Sea was not due to a lack of observation but possibly the result of chemical modification of riverine inputs which prevented the development of K. mikimotoi blooms. A counter argument to this is that these waters are too turbulent and the ability of K. mikimotoi to migrate vertically has no competitive advantage over other species. In our opinion the geographical distribution of K. mikimotoi results from suitable ecohydrodynamic conditions (seasonally stratifying waters) and the adaptation of this phytoplankter to these conditions. There is considerable evidence to support this view. Dahl (1989) and see Dahl and Tangen (1993) suggested that summer populations of K. mikimotoi develop in the North Sea and Skagerrak within the pycnocline of stratified waters and that large blooms develop in coastal waters of the Skagerrak following the horizontal advection of these offshore populations. Once inshore, this phytoplankter is transported to the west coast of Norway by the Norwegian coastal current. Similar mechanisms have been suggested for the occurrence of large K. mikimotoi blooms in coastal waters of Ireland and Scotland. Raine et al. (1993) concluded that a K. mikimotoi bloom in Bantry Bay (south west Ireland) during the summer of 1991, had been advected towards the coast from the shelf (see also Raine et al. 2001). Jones et al. (1982) discussed whether the large bloom that occurred in sea lochs of the Firth of Clyde in 1980 was seeded from populations growing at fronts at the entrance to the Firth and note that Pingree et al. (1978) recorded the presence of K. mikimotoi at the Islay front (located between Malin head and the island of Islay, Figure 5.5). Gowen et al. (1998) also

- 175 - observed a population of 0.3 x 106 cells L-1 at the Islay front in August 1996. Davidson et al. (2009) were of the opinion that the prolonged bloom of K. mikimotoi that occurred in Scottish coastal waters in 2006 originated in shelf waters.

Figure 5.7 The distribution of Karenia mikimotoi in north western European waters. Redrawn from Smayda (1990).

K. mikimotoi < 106 cells L-1

K. mikimotoi > 106 cells L-1 and coloured water

A comparison between the distribution given by Smayda (1990) and the current distribution (Figure 4.7) shows an apparent absence from the Irish Sea in recent years. Although K. mikimotoi formed several large harmful blooms in the eastern Irish Sea during the early 1970s (see Part 2) there are no published accounts of blooms since that time. The Irish Sea is isolated from more open shelf waters to the south and north and this may restrict the transport of cells from populations present in more open shelf waters. For example, at the time of the large K. mikimotoi on the west coast of Ireland in July 2005 (Silke et al. 2005) K. mikimotoi was

- 176 - observed in Loughs Foyle (31,100 cells L-1) Larne (17,260 cells L-1) and Belfast (400 cells L-1) on the north coast of Northern Ireland (FSA(NI) unpublished data) and in waters to the north of the North Channel (4,551 cells L-1, but not in the Irish Sea (AFBI unpublished data). Within the Irish Sea, offshore western waters seasonally stratify (Gowen et al. 1995; Horsburgh et al. 2000) and during the summer are separated from vertically mixed waters to the east by the western Irish Sea front (Simpson & Hunter 1974). The stratified waters and frontal region might therefore be expected to provide suitable conditions for dinoflagellates and especially Karenia mikimotoi but the importance of the western Irish Sea front in providing a suitable environment for dinoflagellates is unclear. Detailed observations of phytoplankton composition in the frontal region are lacking but high phytoplankton biomass does not appear to be associated with the front (Richardson et al. 1985). Furthermore, we are unaware of any reports of large surface K. mikimotoi blooms occurring at the western Irish Sea front (despite the region being regularly traversed by a variety of ships including research vessels). Finally, the occurrence of K. mikimotoi blooms in the Seto Inland Sea where there are numerous tidal mixing fronts and seasonally stratifying waters is consistent with the hypothesis that this phytoplankter is adapted to conditions at fronts and stratified waters. Imai (2006) were of the opinion that Karenia mikimotoi formed red tides before anthropogenic nutrient enrichment of the Seto Inland Sea and referred to this phytoplankter as ‘an inherent red-tide species’.

5.5.7 Prorocentrum minimum and P. lima

In their review, Heil et al. (2005) considered P. minimum to be:

“potentially harmful to humans via shellfish poisoning; it has a detrimental effect at both the organismal and environmental levels; blooms appear to be undergoing a geographical expansion over the past several decades; and, a relationship appears to exist between blooms of this species and increasing coastal eutrophication.”

Prorocentrum minimum appears to be capable of: high growth rates (up to 3.54 d-1); photosynthesising under a range of irradiance regimes; utilising a variety of nutrient sources including nitrate, ammonium, urea and both inorganic and organic phosphorus and is mixotrophic (Heil et al. 2005 and references cited therein). According to Heil et al. (2005) all of these ecophysiological characteristics make P. minimum responsive to eutrophication. In considering the toxicity however, Heil et al. (2005) note that evidence linking this species to toxicity in shellfish is equivocal. Data from the UK is similarly inconclusive. Toxicity in farmed mussels (Mytilus edulis) from Belfast Lough in August 1999 was associated with an

- 177 - extensive bloom (> 5 x 106 cells L-1) of this phytoplankter (FSA(NI) unpubl. data.) but there was no reported shellfish toxicity during a bloom (2.4 x 106 cells L-1) in Shetland in 2007 (Bresnan et al. 2007). The circumstances which allowed this phytoplankter to bloom in Belfast Lough in August 1999 and in Shetland in 2007 are unclear and we do not know why this species has not bloomed in these waters since. Between 2006 and 2008, P. minimum was found in low abundance throughout UK and Irish coastal waters (Figure 4.7) with generally higher abundance in coastal waters of the south west of England, west coast of Ireland and west coast of Scotland. The maximum abundance of this species was recorded in waters to the west and North of Scotland.The data do not support the contention that P. minimum has responded to anthropogenic nutrient enrichment and the current geographical distribution suggests that the abundance of this species is determined by the more favourable ecohydrodynamics conditions that occur in western coastal waters of Ireland and Scotland. Unlike the other dinoflagellates discussed in this study, Prorocentrum lima is a benthic/ epiphytic species and can be considered as a different lifeform. The distribution of this species might therefore be expected to be governed by the intersection between a different combination of ecophysiological characteristics and ecohydrodynamic conditions compared to the pelagic dinoflagellates. According to Taylor et al. (2003), P. lima has a worldwide distribution from tropical to subarctic waters. In his book on dinoflagellates of the British Isles, Dodge (1982) states that this dinoflagellate has been found in almost every sandy and muddy shore that has been sampled, is never found in the plankton and appears to be absent from the north east coast of England. Studies of the seasonal distribution of P. lima have been undertaken in the Canadian Gulf of St. Lawrence (Levasseur et al. 2003), UK Fleet Lagoon (Foden et al. 2005) and by Maranda et al. (2000, 2007) in coastal waters of the Gulf of Maine (U.S.). In these studies, P. lima was found growing attached (by a coating of mucilage) to a variety of macrophytes and sea grasses and was only rarely found in the water column. For example, Maranda et al. (2007) only recorded P. lima in the water column in 4 out of 353 samples over a two year period. It has been suggested (Levasseur et al. 2003; Maranda et al. 2007) that the seasonal variation in the abundance of P. lima is due to changes in macroalgal substrate and it would appear that P. lima is mostly found in low or moderately low energy environments. Maranda et al. (2007) found P. lima to be almost absent from two open water sites, present at low abundance at two sites with low but persistent turbulence and found significant populations in semi enclosed bays that were protected from breaking waves and strong currents.

- 178 - Recent data (2005 – 2007) show that P. lima occurs throughout coastal waters of the UK and Ireland (Figure 4.7) but at low abundance. The low abundance makes it difficult to obtain a clear picture of its distribution, although the data in (Figure 4.7) are indicative of higher abundance in waters to the south west of England, the west of Ireland and to the west and north of Scotland. Such a distribution is consistent with the hypothesis that low turbulence is a key ecohydrodynamic feature governing the distribution of this species. Furthermore, many of the turbid and tidally stirred estuaries of southern England such as the Humber and Wash provide poor environments for the growth of macrophytes and seas grasses and their absence/ low abundance in these waters may also influence the abundance of P. lima.

5.5.8 Lingulodinium polyedrum and Protoceratium reticulatum

The recent data (2005 – 2008) show that L. polyedrum only occurs infrequently in UK and Irish coastal waters. The data also show that the maximum abundance is generally low (e.g. 280 cells L-1 in England and Wales between 2005 and 2008 and 2,440 cells L-1 in Irish coastal waters between 2005 and 2007) although as detailed in Part 2, a small bloom (0.148 x 106 cells L-1) was observed in the Scottish Loch Creran in 1983 by Lewis (1985). There is little indication that Loch Creran was enriched at that time (Tett & Wallis 1978, give a winter nitrate concentration of 10 µM for 1972-1973) indicating that L. polyedrum can bloom in natural (unenriched conditions). The factors controlling the abundance of L. polyedrum are largely unknown but the infrequent occurrence and generally low abundance argues against nutrient enrichment being an important factor. Dodge (1982) stated that P. reticulatum is found all round the British Isles and is common in the North Sea. However, the data compiled for this study (2006 to 2008) show this phytoplankter to be largely absent from UK and Irish waters. The available data do not support the contention that the abundance of P. reticulatum in these waters is influenced by anthropogenic nutrient enrichment.

5.5.9 Aquaculture and HABs

We have attempted to explain the distribution of some of the HAB species and their abundance in UK and Irish waters on the basis of the intersection between their ecophysiology and the ecohydrodynamic conditions in these coastal waters. It is evident that some HAB species are generally more abundant in western coastal waters and this is where most of the aquaculture (finfish and shellfish) is located. One argument might therefore be that the greater abundance of HAB species is related to aquaculture, especially the intensive cultivation of fish.

- 179 - Much has been written on the interaction between aquaculture and the environment especially the potential for intensive fish farming to bring about nutrient enrichment (Gowen & Bradbury 1987; Rosenthal et al. 1988; GESAMP 1991; Gowen 1994). As noted in the introduction to Part 4, however, the way nutrient and phytoplankton data were compiled and analysed involved a geographical comparison on the scale of the British Isles and coastal waters of the Republic of Ireland and that such an analysis does not rule out the possibility of nutrient – HAB correlations for individual water bodies. We cannot therefore address this question directly but present the key findings of relevant studies carried out in Scotland. Gowen and Ezzi (1992) studied the effects of a large fish farm on nutrient levels and phytoplankton in Loch Hourn, a sea loch on the west coast over a two year period and concluded that although there was an increase in the concentration of ammonium during the period that the fish farm was in operation, there was no evidence of a change in the phytoplankton. In reviewing the occurrence of HABs in Scottish coastal waters and the suggestion that intensive fish farming has led to an increase in HABs, Tett and Edwards (2002) concluded that:

“Concerning the allegation that increases in marine salmonid farming have been responsible for apparent increases in HABs in Scottish waters, we conclude that neither the increase in HABs nor the link to fish farming is supported by the available direct evidence - which is, however, incomplete. There seems little doubt that human influenced nutrient enrichment is having some effect on some Scottish waters, but the mathematical logic involved in the calculation of equilibrium concentration enhancements suggests that fish farms are likely to have a detectable effect only in enclosed basins in which water exchange is slow in relation to nutrient loading.”

A further review was undertaken by Rydberg et al. (2003) who considered the potential for eutrophication, for fish farms to affect algal communities and the available scientific evidence on the linkage between aquaculture and algal bloom development. With respect to the first issue Rydberg et al. (2003) concluded that:

“The issue is well clarified in Tett and Edwards (2002) and in SE (2002). Excessive nutrient loads may have effects on algal communities and algal blooms and also cause oxygen deficit in deep waters and bottoms. The fish farming nutrient wastes are not at present levels large enough to cause negative effects except in a few fjords with restricted exchange. The Scottish waters are with a few exceptions (Clyde, Forth, Moray) neither hypernutrified nor eutrophic. The west coast fjords are not eutrophic, but may still be sensitive to additional nutrient loads.” and in relation to the second issue that:

- 180 - “Available evidence indicate that it is unlikely that fish farming should have an impact on the occurrence of harmful algal blooms, and particularly on such blooms, which are related to shellfish poisoning (ASP, DSP, PSP). Large scale mapping of biotoxic blooms shows that these blooms are more frequent in open areas, away from eutrophic waters, and appear independent of fish farming. Tett and Edwards (2002) provides a good background to this issue, indicating that there is a coupling between higher nutrient loads and more comprehensive blooms (of all kinds), but that the size of the fish farms are not enough to cause such notable effects. The potential for changing nutrient ratios in a way that favours specific algal communities is also small. However, this might happen in some enclosed fjords.”

A third review was undertaken by Smayda (2006) who drew a similar conclusion to the authors of the previous two reviews:

“Using the year 1985, when fish farming accelerated, as a branch point, differences in regional bloom patterns and frequencies during the pre- and post-1985 period are not evident. Similarly, the patterns and trends in harmful species of Alexandrium, Dinophysis, Pseudo-nitzschia, phytoflagellates, diatoms and ichthyotoxic Karenia mikimotoi do not show a detectable relationship with increasing delivery of fish farm nutrients. This conclusion agrees with that reached by Tett and Edwards, and by Rydberg and co-workers who applied different, but related analytical approaches.” and with respect to the central question of his review i.e. whether aquaculture development has resulted in an increase in the occurrence of HAB events or whether such events are a part of the natural dynamics of phytoplankton in Scottish coastal waters, Smayda (2006) concluded:

“● There is no evidence of a significant increase in nutrient levels, altered phytoplankton behavior, or an increase in harmful algal blooms in Scottish waters. ● While blooms at fish farm sites are known from other regions, and there is experimental evidence that fish wastes can both stimulate and inhibit the growth of harmful species, there is no evidence for such impacts in Scottish waters. ● Blooms of the harmful species present in Scottish waters are not dependent on aquacultural stimulation; all harmful species bloom in habitats not influenced by fish farm wastes or shellfish cultivation. ● The differences in harmful blooms that occur between Scotland and elsewhere in Europe can not be related to differences in aquaculture intensity or, within this, whether fish farming or shellfish cultivation is the more prominent. ● The current level of shellfish aquaculture in Scottish coastal waters is not a factor in harmful bloom stimulation. ● Based on the data available, the observed phytoplankton behavior in Scottish coastal waters does not appear to differ significantly from the natural and variable behavior expected of an indigenous phytoplankton flora exposed to the "open system" features of boreal

- 181 - waters. If anything, there is a surprising lack of anomalous bloom behavior in contrast to that recorded in Scandinavian waters, where salmonid fish farming is also extensive, and an absence of HAB induced farmed fish kills in contrast to those occurring in Scandinavian waters and at Pacific salmonid fish farms.”

The overall conclusion from these studies is that the occurrence of HABs in Scottish coastal waters is not related to fish farming and shellfish cultivation. On the basis of our analysis we conclude that the greater abundance of some HAB species in south west England, Scottish coastal waters and souther and western waters of the Republic of Ireland is a consequence of the intersection between the characteristic ecohydrodynamic conditions found in these waters and the adaptation of particular HAB species to these conditions.

5.6 Synthesis

5.6.1 Introduction

In this study we set out to investigate relationships between anthropogenic nutrient enrichment and harmful algal blooms. The subject matter is complex and anthropogenic nutrient enrichment is one of several pressures that influence phytoplankton species and population dynamics. For these reasons, the approach taken was to review the literature, focussing on four geographical areas in particular and undertake an analysis of phytoplankton and nutrient data from coastal waters of the UK and the Republic of Ireland. By way of drawing together information in earlier sections of the report, in this section we consider the nutrient enrichment → HAB hypothesis by asking a series of questions.

5.6.2 Does the occurrence of HABs imply eutrophication and is eutrophication always accompanied by HABs?

The answer to the first part of this question is clearly no because some HABs occur naturally. Examples of such HABs include the early records of PSP in British Columbia (Canada); the large Karenia mikimotoi blooms in coastal waters of western Ireland and Scotland and Alexandrium fundyense blooms in the Gulf of Maine (U.S.). For this reason (and because other pressures such as climate change can influence the occurrence of HABs) the occurrence of HABs does not diagnose eutrophication. However, Tett et al. (2007) argued that an increase in the frequency or spatial/ temporal extent of HABs related to anthropogenic nutrient enrichment would represent an undesirable disturbance and hence be indicative of eutrophication.

- 182 - Whether eutrophication is always accompanied by HABs is more equivocal. We cannot provide a definite answer to this question but consider that an increase in HABs or in their spatial and temporal extent is one potential outcome of anthropogenically driven eutrophication.

5.6.3 Has an increase in HABs been reported and is this increase real?

The literature review in Part 3 has identified a number of peer review publications the authors of which have argued that on a global scale there has been an increase in harmful algal blooms. The answer to the first part of this question is therefore yes; an increase in HABs has been reported. On a global scale, it is not possible at present to provide a definitive answer to whether the reported increase is real because there are a number of confounding issues. Any answer depends on the definition of what a harmful algal bloom is. In some cases phenomena not previously called HABs have been redefined. As discussed, the term 'bloom' has several meanings but we conclude that a bloom is a discrete event:

an increase in abundance of one or more species, which stands out from what has happened before and the state to which the phytoplankton returns after a bloom.

An increase in abundance is relative to a background level which may be 0, low or high, depending on the organism and following from this definition, we define a 'HAB' as:

a discrete event associated with a 'bloom' of micro-algae or cyanobacteria that damages human use of ecosystem goods and services

It is recognised that there is a need to distinguish large-biomass HABs and HABs associated with low abundances of toxin producing algae. It is also necessary to associate different causal models to large and low biomass HABs and this argues against a single causal pressure. Furthermore, in some regions of the world there has been an unprecedented increase in monitoring over the last 15 – 20 years (e.g. northwest Europe) and it is difficult to discriminate between the increased reporting of HABs associated with this and what might be a real increase in their occurrence.

5.6.4 Does nutrient enrichment lead to more large-biomass HABs?

It is evident from the case studies that anthropogenic nutrient enrichment has caused an increase in large biomass HABs in some regions of the world. For the small enclosed Tolo Harbour in Hong Kong and the larger semi- enclosed Seto Inland Sea of Japan, the data available provide

- 183 - evidence for a nutrient driven increase in HABs. In the case of Dutch coastal waters, there is evidence of a temporal increase in the size and duration of spring Phaeocystis spp. blooms linked to enrichment. It is also the case that in many publications, evidence is more equivocal or lacking. A number of studies allude to an increase in HABs as a result of nutrient enrichment but provide little or no detail. For some coastal waters of China, increasing trends in the occurrence of HABs are evident, but relationships with riverine nutrient loadings are complex. In some of the examples discussed in Part 3, the temporal trend in HABs is not consistent with the trend in riverine nutrient load. In other examples, increased environmental monitoring confounds attempts to identify trends in HABs and it is evident that the inter-annual variability in the occurrence and scale of some HABs has been related to climate change.

5.6.5 Does nutrient enrichment lead to greater abundance of toxin producing species and hence an increase in low biomass HABs?

This question was the subject of the analysis of UK and Irish data presented in Part 4 and discussed in Part 5. Results of the analysis carried out in Part 4 show that the summer abundance (taken as the mean and maximum abundance between April and September) of HAB species in UK and Irish coastal waters is not determined by anthropogenic nutrient enrichment. These results support the suggestion made during a workshop during the 4th international conference on Toxic marine phytoplankton in 1990 (see Smayda & White 1990) where it was concluded that: “The causes and mechanisms of blooms may differ as there seem to be two major types of blooms, those in which nutrient additions to coastal systems are obviously implicated (for example, in Tolo Harbour (Hong Kong), in the Seto Inland Sea (Japan), and in the Aegean Sea in the vicinity of sewage outfalls) and those blooms that are not obviously associated with coastal enrichment (for example, Alexandrium, Pyrodinium, Dinophysis, etc.).”

Few other studies have looked in detail at the relationship between anthropogenic nutrient enrichment and low biomass HABs of toxin producing species. Of these, the study by Trainer et al. (2003) is important because it demonstrated an increase in toxic episodes associated with Alexandrium catenatum in the fjordic Puget Sound (Washington State, U.S.) and found this increase to be correlated with changes in human population in the catchment area around the Sound (see Part 3). These findings are in contrast to our results but in our opinion reflect differences in the ecohydrodynamic characteristics of Puget Sound and enriched coastal waters of the UK. According to Trainer et al. (2003), all of the waters of Puget Sound are influenced by freshwater inflow that results in density dependent stratification and earlier in this section we

- 184 - argued that the introduction of nutrients into the illuminated surface mixed layer of stratified waters is likely to stimulate algal blooms (some of which may be HABs). In contrast, enriched UK coastal waters such as the estuaries of the south east of England (see also discussion of the eastern Irish Sea above) are turbid and tidally stirred: conditions more likely to constrain phytoplankton growth (light limitation) and favour diatoms as the dominant lifeform.

5.6.6 Do shifts in nutrient ratios lead to more HABs

Evidence of a link between anthropogenically driven changes in N:P and N:Si ratios and HABs has been reported in the scientific literature (Part 3). It has been suggested that in continental coastal waters of Europe, Phaeocystis spp. blooms follow the diatom spring bloom because diatom growth becomes limited by silicate before nitrogen and phosphate are depleted. However this hypothesis is not supported by results from mesocosm experiments. We suggest that for continental coastal waters of Europe the relationship between the occurrence of Phaeocystis spp. blooms and nutrient ratios (N:P and N:Si) remains unresolved. With respect to Tolo Harbour, the suggestion that a reduction in N:P ratio favoured the occurrence of HABs (Hodgkis & Ho 1997) is not supported by more recent data. For UK and Irish coastal waters there were three significant positive relationships between the N:P loading ratio and the mean and maximum abundance of Dinophysis spp. and the mean abundance Karenia mikimotoi that would seem to support that argument. Nevertheless, if such effects were strong, they should have led to a higher proportion of significant relationships With respect to ratios of N (as TOxN and DIN):Si loadings and winter concentrations and HAB species abundance, there were 11 significant negative regressions. That is, higher abundance was associated with low N:Si ratios. This would appear contrary to the arguments usually presented in the literature (see Part 3), which have increases in N:Si changing the balance of organisms in favour of harmful algae. An alternative explanation is that waters with elevated N:Si ratios are enrichment, and are in coastal waters where hydrodynamic conditions are unsuitable for most harmful lifeforms. Thus, the negative relationships with N:Si ratios could be an artefact of the significant negative relationships between abundance and loadings/ concentrations. It should also be remembered that our data are for winter ratios, or loading ratios, which may not correspond to actual nutrient ratios during summer, when harmful algae are most likely to be abundant. We conclude that based on theoretical grounds and the data available, large shifts in nutrient ratios are necessary to bring about changes in HAB species and an increase in the frequency of HABs.

- 185 - 5.6.7 Are toxin producing algae more toxic when nutrient ratios are perturbed in the sea?

It is evident that there is a complex relationship between cellular growth, toxin production and nutrient availability and supply ratios. Silicate limitation may promote domoic acid (DA) production by Pseudo-nitzschia spp. (Fehling et al. 2004; Pan et al. 1996) although a more complex environmental control seems likely (Fehling et al. 2005; Marchetti et al. 2004; Wells et al. 2005). PSP toxins are nitrogenous compounds and N stress will be detrimental to PSP toxin synthesis (Flynn & Flynn 1995). The role of nutrients in promoting PSP toxin production may be species specific although toxin production may also be influenced by a variety of abiotic (temperature, light, nutrient concentration) and biotic (competitors, grazers) factors (Granéli et al. 1998). Increased toxicity has been linked to P stress (Anderson et al. 1990; Boyer et al. 1987; John & Flynn 2002) and P deficiency increased toxin content per cell in Alexandrium tamarense and Gymnodinium catenatum (Granéli et al. 1998, but see Flynn & Flynn 1995). The role of dissolved and particulate matter in toxin production and toxicity is largely unknown (Granéli et al. 1998). Both N and P limitation has been shown to produce similar levels of DSP toxicity in Prorocentrum lima and for Dinophysis, the highest toxin content in cells occurred under N limitation (Granéli et al. 1998). Under semi-continuous culture conditions the toxicity of Chrysochromulina polylepis was strongly influenced by the physiological state of cells and may explain the large variability in the toxicity of this species (Johansson & Granéli 1999). There is an increasing amount of work involving algal cultures that shows harmful algae becoming more toxic when cells are 'nutrient stressed' i.e when growth slows because one nutrient becomes limiting and nutrient supply ratios are markedly different from Redfield. The general explanation seems to be that toxin is synthesized while biomass synthesis slows. Such findings might imply that blooms are likely to become more toxic towards their end, but do not help to explain any widespread increase in HABs or toxicity. Perturbations of nutrient ratios induced by anthropogenic nutrient enrichment could influence toxicity, however evidence from field studies are generally lacking and whether there is a causal link remains hypothetical. Such an effect might be expected to result from changes in nutrient ratios only in semi-enclosed, near- shore, highly loaded, waters.

5.6.8 Is the distribution of HAB species in UK and Irish waters related to niche requirements and ecohydrodynamics?

For some of the HAB species that occur in UK and Irish waters their distribution can be interpreted as being the result of ecophysiological adaptation to particular ecohydrodynamic - 186 - conditions. Furthermore, the distribution of HAB species (Figure 4.7) is suggestive of a connectivity between western coastal regions of the UK and Ireland brought about by the Irish and Scottish coastal currents (see Figure 5.6). Karenia mikimotoi is adapted to stratified waters and perhaps tidal mixing fronts in particular. We hypothesise that in comparison to enriched but tidally stirred waters of the eastern Irish Sea and south east England, the greater abundance of K mikimotoi in western coastal waters is due to many of these waters undergoing seasonal thermo- haline stratification and their proximity to stratified waters and frontal boundaries of western shelf waters. Advective processes such as downwelling at the coast serve to connect open shelf and coastal waters thereby promoting the transport of cells from offshore populations and the Irish and Scottish coastal currents provide mechanisms for transporting populations along the coast. A key part of the life cycle of Alexandrium spp. is the formation of cysts that maintain populations of vegetative cells from one year to the next. Depositional areas where particulate material settles to the sea bed also allows cysts to settle out of the water column and remain in the sediment over winter. The presence of Alexandrium cysts in sediments in some RREs (e.g. the estuaries of south west England, the Firth of Forth, Belfast Lough) and off the Rivers Tyne and Tees, provide an explanation for these regions being coastal ‘hotspots’ for the occurrence of Alexandrium spp. Details of the ecophysiology (particularly its complex nutrition) of Dinophysis spp. are lacking. Nevertheless, the greater abundance of these phytoplankters in seasonally stratifying waters to the south and west of Ireland, and in Scottish coastal waters is consistent with dinoflagellates being adapted to stratified conditions and the ability of these species to form dense populations (several thousands of cells L-1) in thin layers within the pycnocline of seasonally stratified waters. Species of Pseudo nitzschia are ubiquitous throughout UK and Irish waters but there does not appear to be any obvious pattern to their spatial distribution. Whether the ability of these species to form thin layers in stratified waters is a key ecophysiological characteristic that would result in higher abundance in seasonally stratifying coastal waters is unclear. Interestingly, of all of the coastal waters of the UK and Ireland, species of Pseudo-nitzschia were less abundant in the enriched Thames estuary region and this may be due to silicate limitation of diatom growth (Sanders et al. (2001). For UK and Irish coastal waters, the data do not support the view of Heil et al. (2005) that P. minimum is associated with enriched coastal areas. The current geographical distribution suggests that the abundance of this species is determined by the more favourable ecohydrodynamics conditions that occur in western coastal waters of Ireland and Scotland.

- 187 - The low abundance of P. lima makes it difficult to obtain a detailed picture of the distribution of this phytoplankter. The data are suggestive of a greater abundance in seasonally stratifying western coastal waters of Ireland and Scotland and lower abundance in tidally stirred waters of the east coast of England and coastal waters of the Irish Sea. Such a distribution is consistent with the niche requirements (benthic and epibenthic habit with suitable substrate in low turbulence coastal areas) of P. lima. The factors controlling the abundance of L. polyedrum and P. reticulatum are largely unknown but the infrequent occurrence and generally low abundance argues against nutrient enrichment being an important factor.

5.7 General Conclusion

Evidence is presented that shows: HABs (as defined for the purposes of this report) occur naturally; anthropogenic nutrient enrichment has increased the occurrence of large biomass HABs in some water bodies but not in others; the global evidence for enrichment having brought about an increase in low biomass HABs of toxin producing species is more equivocal and the UK and Irish data do not support hypothesized relationships. The influence of climate change on the occurrence of HABs together with increased environmental monitoring and reporting of HABs and the limited time-series of data currently available confounds attempts to link nutrient enrichment to the occurrence of HABs. We therefore conclude that there is no single general hypothesis for changes in the occurrence of HABs but hypothesise that: their occurrence is the result of interactions between changes in specific pressures (including nutrient enrichment), the ecohydrodynamic conditions in particular water bodies and the adaptations of particular harmful algal species or life-forms. As a consequence, we are of the opinion that the occurrence of HABs and the abundance of HAB species should not be used to diagnose eutrophication unless a link to anthropogenic nutrient enrichment can be demonstrated. Furthermore, evidence of a link in one coastal region should not be taken as evidence of a general linkage in other coastal regions.

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PAPERS cited by other authors and not reviewed during this study

Azanza RV (1999) Seafood poisoning from harmful algal blooms in coastal areas (abstract only). In: IOC-SOA International Workshop on Coastal Megacities: Challenges of Growing Urbanisation of the World’s Coastal Areas Intergovernmental Oceanographic Commission Workshop Report No 166 (Hangzhou, People’s Republic of China 27–30 September 1999). UNESCO Choi KW, Lee JHW (2004) Numerical determination of flushing time for stratified waterbodies. J Mar Syst 50:263-281 Evans D (1976) The occurrence of Gyrodinium aureolum in the Eastern Irish Sea, 1975. (Provided by) P.A. Driver at a meeting of Liverpool Bay Working Group (Standing Committee on the disposal of sewage sludge) LBWG (76)16 Gainey LF, Shumway SE (1989) Effects of Aureococcus anophagefferens (brown tide) on the lateral ciliary activity of bivalve mollusks. Am Zool 29:A72-A72 - Han XR, Wang XL, Sun X, Shi XY, Zhu CJ, Zhang CS, Lu R (2003) Nutrient distribution and its relationship with occurrence of red tide in coastal area of East China Sea. Chinese Journal of Applied Ecology 14:1097-1101 (in Chinese) Hardy AC (1925) Part II. - Report on Trials with the Plankton Indicator, Ministry of Agriculture and Fisheries. Fishery Investigations, Series II, Vol VIII, No.7 Ho KC, Hodgkiss IJ (1995) A study of red tides caused by Prorocentrum micans Ehrenberg, P.sigmoides Bohm and P.triestinum Schiller in Hong Kong. In: Morton B, Xu G, Zhou R, Pan J, Cai G (eds) The Marine Biology of the South China Sea II. World Publishing Corporation, Beijing, PRC, p 111-118

- 216 - Jackson D, Doyle J, Moran M (1991) Algal blooms around the Irish Coast in 1990 (Toxic and Nuisance Species). In: Aquaculture and the Environment. EAS special publication No. 14 p163 Li RX, Zhu MY, Wang ZL, Shi XY, Chen BZ (2003) Mesocosm experiment on competition between two HAB species in East China Sea. Chinese Journal of Applied Ecology 14:1049-1054 (in Chinese) Nakanishi H (1993) In: Environmental Characteristics, Tokyo Bay - Its Environmental Change in a Hundred Years. Ogura N, Koseisha Koseikaku (eds) (in Japanese) p159-162 Reinecke P (1967) Gonyaulax grindleyi sp. nov. A dinoflagellate causing red tide at Elands Bay, Cape Province, in December 1966. J S Afr Bot 33:157-160 Wang J, Huang X (2003) Ecological characteristics of Prorocentrum dentatum and the cause of harmful algal bloom formation in China Sea. Yingyong Shengtai Xuebao 14:1065-1069

- 217 - Annex I

Project partners and their affiliations

Dr Richard Gowen Mr Alan Gordon1 Mrs April McKinney Mrs Ann Marie Crooks

Fisheries and Aquatic Ecosystems Branch Agriculture Food and Environmental Sciences Division Agri Food and Biosciences Institute Newforge Lane Belfast, BT9 5PX

1 Biometrics Branch, Applied Plant Science and Biometrics Division

Professor Paul Tett Dr Keith Davidson

Scottish Association for Marine Science Dunstaffnage Marine Laboratory Oban Argyll, PA37 1QA

Dr David Mills Mr Steve Milligan

Centre for Environment, Fisheries & Aquaculture Science Pakefield Road Lowestoft Suffolk, NR33 0HT

Dr Eileen Bresnan

Marine Scotland Marine Laboratory P.O. Box 101 Victoria Road, Aberdeen AB11 9DB

Mr Joe Silke

The Marine Institute Rinville Oranmore Galway

- 218 - Annex II

Pictures of selected species of phytoplankton

Dinoflagellates

Alexandrium tamarense (AFBI) Alexandrium tamarense (SEM40) (AFBI) (22 – 44 x 20 - 36 µm, l x w)

Dinophysis acuminata (AFBI) Dinophysis acuta (L Naustvoll) (38 – 58 µm long) (54 – 85 µm long)

40 Scanning Electron Micrograph showing the thecal plates

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Prorocentrum lima (SAMS) Karenia mikimotoi (Cefas) (32 – 50 x 20 – 28 µm, l x w) (24 – 40 x 17 – 32 µm, l x w)

Prorocentrum minimum (AFBI) P. minimum (SEM) (AFBI) (14 – 22 x 10 – 15 µm, l x w)

Noctiluca scintillans (A Alazri) (diameter 200 – 2,000 µm)

- 220 - Diatoms

Pseudo-nitzschia spp. (AFBI) (33 – 160 µm long)

Chaetoceros brevis (E Capuzzo) Leptocylindrus danicus (E Capuzzo) (7 – 40 µm diameter) (5 – 16 µm wide)

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Microflagellates

Chattonella sp. (L Naustvoll) Chrysochromulina polylepis (L Naustvoll) (30 – 50 µm long) (6 12 µm long)

Photographs A Alazri, Sultan Qaboos University, College of Agriculture and Marine Sciences P.O.Box 34 Al-KhodhPC123 Muscat Oman E Capuzzo, Cefas, Pakefield Road Lowestoft, Suffolk NR33 OHT L Naustvoll, Institute for Marine Research, Research Group Plankton Institute of Marine Research, Flødevigen Nye Flødevigveien 20 N-4817 His, Norway

Cell dimensions Dodge (1982) Thomas (1993, 1996, 1997) (For Chaetoceros brevis: www.10-warnemaende.de/gallery-ofobaltic-microalgae.htmi.)

- 222 - Annex III Acknowledgements Funding This research project was funded by the UK department of Environment Fisheries and Rural Affairs (Research grant ME2208) A number of individuals and organisations provided data, scientific input to the study and critical comment on the report and we would like to thank the following: Dr D Anderson, Woods Hole Oceanographic Institute (Massachusetts, USA) for helpful comments on the draft report. Dr M Best, Environment Agency (England and Wales) for coastal nutrient data. Dr S Boyd Food Standards Agency Northern Ireland (Belfast, UK) phytoplankton data from coastal waters of Northern Ireland. Dr P Harrison, The Hong Kong University of Science and Technology (Hong Kong, China) for helpful comments on the draft report, in particular the occurrence of HABs in coastal waters of Hong Kong. Dr K Hargin Foods Standard Agency (UK) (London, UK) phytoplankton data from England and Wales. Dr W Higman, Cefas (Weymouth, UK) for providing data on PSP toxicity in shellfish from the NE of England. Professor Ichiro Imai, Plankton Laboratory, Division of Marine Biology and Environmental Science, Graduate School of Fisheries Sciences, Hokkaido University (Minatomachi 3-1- 1, Hakodate, Hokkaido 041-8611, Japan) for helpful comments on red tides in Japan. ICES/IOC Working group on Harmful Algal Bloom Dynamics for helpful comments on the draft report. Ms E Joyce Marine Institute (Galway, Ireland) for coastal nutrient data. Dr. Adam Mellor AFBI (Belfast, UK) for providing winter nutrient data from coastal waters of Northern Ireland. Ms L Murry Foods Standards Agency Scotland (Aberdeen) phytoplankton data from Scottish coastal waters. Ms Judy Dobson, Scottish Environmental Protection Agency (Edinburgh, UK) for Scottish coastal nutrient data. Dr Sonja Van Leeuwen, Cefas (Lowestoft, UK) for providing data on modelled UK riverine nutrient loadings. Professor Zhongyuan Chen and Dr Maotian Li, State Key Laboratory of Estuarine and Coastal Research, East China Normal University (Shanghai 200062, China) for nutrient loading times-series data for the Yangzi River. Dr Kedong Yin Australian Rivers Institute (Griffith University, Australia) for provision of nutrient data from Tolo harbour, Hong Kong.

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