Download Final report Anthropogenic nutrient enrichment and blooms of

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
Fisheries 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 Scotland 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.
5
Centre for Environment, Fisheries & Aquaculture Science, Pakefield Road, Lowestoft, Suffolk,
NR33 0HT.
6
The Marine Institute, Rinville, Oranmore, Galway, Ireland.
Contents
Summary
Part 1 Introduction
Page
iv
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 tides, 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 United Kingdom 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 pressures
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 Current 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 concentrations
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 upwelling, 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 health 1 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 OSPAR 2 .
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 ecohydrodynamic 3
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
1
Ecosystem 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 microalgal bloom’. 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 services 4 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., Pseudonitzschia 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
-9-
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 mixotrophic 5 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.
- 10 -
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
-3
Chlorophyll (mg m )
15
12
9
6
3
0
01-Jan
10
9
22-Feb
14-Apr
05-Jun
28-Jul
18-Sep
09-Nov
31-Dec
22-Feb
15-Apr
06-Jun
28-Jul
18-Sep
09-Nov
31-Dec
B
Concentration ( M)
8
7
6
5
4
3
2
1
0
01-Jan
- 11 -
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
- 12 -
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
environments such as the Baltic Sea, phosphate (PO43-) 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).
- 13 -
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 mikimotoi 6 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 Ombai 7 , 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
7
Formally know as Gyrodinium aureolum
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 tide = ‘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, WGHABD 8 ).
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 Oceanography 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 subjective 9 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 nonharmful 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 costatum 10 . 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
pressure e.g. climate change or anthropogenic nutrient enrichment.
- 18 -
Table 1.1 Mortality modes and impact mechanisms of harmful species and their blooms (from
Smayda 1997b).
Effect
Starvation
Nutritional mismatch
Size mismatch
Excess prey density
Mechanical
“Bumping”
Particle irritation
Physical
Viscosity barrier
Gelatinous barrier
Mucoid layer reduction
Anoxia
Ammonium toxicity
Phycotoxins
Direct vs. vectored toxicity
Saxitoxin,
Brevetoxin
Hemolysins
Cytotoxins
Allelopathic 11
Ambush predation
Unresolved
11
Species associated with effect
Chaetoceros spp.
Karenia mikimotoi
Cerataulina pelagica
Chattonella marina
Ceratium blooms
Noctiluca blooms
Pfiesteria piscicida
Aureococcus anophagefferens
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 dopaminelike 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 oxygen 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
oxygen was 1.3 mg O2 L-1 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 tamarense 12
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 fundyense 13 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
13
Previously called Gonyaulax tamarensis.
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 Alexandrium 14 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 polyedrum 15 and Protoceratium reticulatum 16 (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 pungens 17 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
Previously named Gonyaulax grindleyi
17
Previously named Nitzschia pungens
16
- 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).
- 26 -
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 fishing nets and interfere with commercial
fishing (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 fishery 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 reef 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.
- 28 -
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 opinion 18 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 Scotland, 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).
- 29 -
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.
- 30 -
Reports of paralytic shellfish poisoning are rare in the United Kingdom. Based on a reevaluation 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, Eyemouth;
2, Farne Islands; 3, Northumberland; 4, Trow Rocks; 5, Staithes; 6, Southampton; 7,
Plymouth; 8, Penzance; 9, Belfast 10, Loch Long; 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
- 31 -
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
-1
6
exceeds 200 cells L . The exception to this was a bloom (~ 1 x 10 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
-1
spp. (below 2,000 cells L ) 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).
- 32 -
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
- 33 -
(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 Pseudonitzschia 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).
- 34 -
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 Pseudonitzschia 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).
- 35 -
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 Scapa Flow 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
- 36 -
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.
- 37 -
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.
- 38 -
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 speculum 19 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 rubra 20 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 L1
) 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
20
Previously named Distephanous speculum
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).
- 40 -
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
Year/Week
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
Jan
1
2
3
4
Feb Mar Apr
5
6
7
8
9
May
Jun July Aug Sep Oct
Nov
De
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
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 landlocked 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 Tg 21 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
Nitrate ( M)
600
500
400
300
200
100
140
1934
1939
1945
1951
1957
1962
1968
1974
1985
1.8
B
1.6
120
Nitrate and silicate ( M)
1979
1.4
100
1.2
80
1.0
60
0.8
0.6
40
Si
NO3
20
0
DIP
Phosphate ( M)
0
1928
0.4
0.2
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 denitrified 22 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 LOICZ 23 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
as 0.3 mol N m-2 y-1 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-NSP 24 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 fish
farming (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
(PO43-) 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
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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:
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“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
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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.”
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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
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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.
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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
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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 cysts 25 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 period 26 ; 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 temperature and storminess) is the main
driving force 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.
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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
5
In 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
70
1200
1000
80
800
60
600
40
400
20
200
B
60
HAB occurrence
100
Samples per year
Number of stations
A
50
40
30
20
10
0
0
0
0
1980 1983 1986 1989 1992 1994 1997 2000 2003
Ratio of HABs to monitoring
0.40
200
400
600
800
Monitoring frequency
C
0.30
0.20
0.10
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:
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1000
1200
“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 brevis 28 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 nearshore 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 blooms 29 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). Temperatures 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 Odontella 30 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
nitrate, NO3-) for example, increased from 0.005 mg l-1 (0.36 µM) in 1977 to 0.135 mg l-1 (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
0.5
Population
14
12
0.4
10
0.3
8
6
0.2
4
0.1
2
0
1976
Population
Red tide incidents
16
1978
1980
1982
1984
0
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
Red tide incidents
30
25
20
15
10
5
1999
1998
1997
1996
1995
1994
1993
1992
1991
1990
1989
1988
1987
1986
1985
1984
1983
1982
1981
1980
0
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
DAIN
25
20
20
15
15
10
10
5
5
0
0
1976
1979
1982
1985
1988
1991
1994
1997
2000
2003
HABs
B
DAIP
25
2.5
2.0
20
1.5
15
1.0
10
0.5
5
0
1976
Annual mean DAIP `
30
2006
3.0
35
Red tide Incidents `
25
Annual mean DAIN `
30
Red tide incidents
HABs
A
1979
1982
1985
1988
1991
1994
1997
2000
2003
0.0
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
insufficient to support the high biomass (15 x 106 cells L-1 and 56 mg chlorophyll m-3). 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 aureolum 31
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
Both species are considered to be Karenia mikimotoi.
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).
31
32
- 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)
respectively 33 . 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 Yaqian 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
A
Red tide incidents
14
12
10
8
6
4
2
0
1980
2500
1982
1984
1986
1988
1990
1992
1994
1996
1998
2000
1994
1996
1998
2000
B
GDP
2000
1500
1000
500
1982
1984
1986
1988
1990
1992
18
C
16
14
12
10
8
6
4
Red tide incidents
0
1980
2
-15.0
-10.0
-5.0
0.0
5.0
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
Num ber of HABs per year
90
80
70
60
50
40
30
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. costatum 34 . 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
40
Northern region
30
20
10
50
40
20
02
20
00
19
98
19
96
19
94
19
92
19
90
19
88
19
86
19
84
19
82
19
80
0
Eastern region
30
20
10
20
02
20
00
19
98
19
96
19
94
19
92
19
90
19
88
19
86
19
84
19
82
19
80
0
50
40
Southern region
30
20
10
19
84
19
86
19
88
19
90
19
92
19
94
19
96
19
98
20
00
20
02
19
84
19
86
19
88
19
90
19
92
19
94
19
96
19
98
20
00
20
02
70
60
50
40
19
82
19
80
0
Total
19
82
19
80
30
20
10
0
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:
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“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
chlorophyll concentrations > 1 mg m-3 relative to typical concentrations of ≤ 0.5 mg m-3) 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 IndoWest 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 s1
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
westerly monsoon and results in a total volume transport of 3.3 x 1010 m-3 ( 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).
Number of red tides per year
350
A
300
250
200
150
100
50
0
1968
1971
1974
1977
1980
1983
1986
60
Num ber of incidents
50
B
40
30
20
10
1997
1995
1993
1991
1989
1987
1985
1983
1981
1979
1977
1975
1973
1971
0
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
600
200
500
150
400
N load
Red tide frequency
700
N-load
250
300
100
200
50
100
2004
2002
2000
1998
1996
1994
1992
1990
1988
1986
1984
1982
1980
1976
1978
1974
1972
1970
1968
1966
1964
1962
1960
1958
1956
1954
1952
0
1950
0
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
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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
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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:
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“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.
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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 1632 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)
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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
concentrations.”
inorganic
phosphate
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had
dropped
to
minimal
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.”
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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
B
A
54
Latitide
53
53
52
52
51
0.3
1.2
2.1
2.9
3.8
4.7 0.3
55
1.2
2.1
2.9
3.8
4.7
Longitude
Longitude
C
55
Latitude
54
54
53
53
52
0.3
1.2
2.1
2.9
3.8
4.7
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)
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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
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(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
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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
Duration of bloom (days)
180
160
140
120
100
80
60
40
20
2000
1998
1996
1994
1992
1990
1988
1986
1984
1982
1980
1978
1976
1974
0
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 &
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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 longterm 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 humaninduced 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 NANI 36 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.
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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
NE – Gulf of Maine
No
NE and mid Atlantic
Mid Atlantic
Gulf of Mexico
- west Florida shelf
Indirect
Yes
- Texas coast
California
Yes
No
Pacific northwest
No
Yes
Comment
Conjecture in relation to blooms in bays and harbours,
but data lacking.
Possible importance of DON not DIN
Chesapeake Bay
Maybe
There is an ongoing debate, but at present insufficient
data to resolve the issue
Conjecture in relation to stimulation of populations
once they are transported inshore.
Heterosigma akashiwo
Alexandrium catenella (Puget Sound but not along the
open coast)
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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,
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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 (24) 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) photoautotrophs 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 &
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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
within their biomass. However, if the ratio NkQ:SikQ 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.
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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)
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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).
25
A
35
25
Red tide incidents
20
N:P
15
10
15
10
5
5
0
0
1982
1983
1984
1985
1986
1987
1988
1989
70
40
B
35
60
N:P
30
N:P
50
25
40
20
30
15
20
10
- 115 -
2005
2004
2003
2002
2001
2000
1999
1998
1997
1996
1995
1994
1993
1992
1991
1990
1989
1988
1987
1986
0
1985
0
1984
10
1983
5
1982
Number of red tides per year
N:P molar ratio
20
30
N:P molar ratio
Number of red tides per year
40
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).
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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 unenriched 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 Pseudonitzschia 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 lifeform) 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-ratiodriven 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 solution (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 microalgae 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 semicontinuous 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
Sc5
Latitude
Sc1b
E1
NI1
Sc1a
Northern
Ireland
E2
E4
Sc1
E2a
E3 E5
E6
E30
NI2
54
E27
E29
E28
E7
E7A
E8
England
Wales
E26
E10
E9
52
E25
Celtic
Sea
E21
E19
E12
E13
E17
E18
50
E11
E23 E22
E24
E20
E15
E16
E14
English
Channel
France
48
10
8
6
4
Longitude
- 125 -
2
0
2
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
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
Measured



Winter concentrations
Modelled







UK data













Irish data











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
Latitude
57
55
53
51
49
11
9
8
6
5
3
1
0
2
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
2000 - 2004
2000 – 2004
Karenia mikimotoi
2002 - 2006
2002 - 2006
Prorocentrum lima
2006 - 2008
2005 - 2006
Prorocentrum minimum
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
2000 - 2004
2000 – 2004
Karenia mikimotoi
2002 - 2007
2002 - 2007
Prorocentrum lima
2006 - 2008
2005 - 2007
Prorocentrum minimum
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
Alexandrium spp.
Dinophysis spp.
Pseudo-nitzschia spp.
Karenia mikimotoi
Prorocentrum lima
Prorocentrum minimum
Nutrient
NO3
NO2
NH4
TOxN
DIN
PO4
Si
NO3
NO2
NH4
TOxN
DIN
PO4
Si
NO3
NO2
NH4
TOxN
DIN
PO4
Si
NO3
NO2
NH4
TOxN
DIN
PO4
Si
NO3
NO2
NH4
TOxN
DIN
PO4
Si
NO3
NO2
NH4
TOxN
DIN
PO4
Si
Mean abundance
P
R2
0.878
0.608
0.587
0.897
0.789
0.439
0.781
0.224
0.085
0.020
0.227
0.109
0.039
0.032
0.996
0.743
0.363
0.984
0.844
0.384
0.132
0.344
0.175
0.213
0.336
0.276
0.190
0.754
0.068
0.096
0.045
0.073
0.037
0.092
0.424
0.127
0.488
0.127
0.137
0.091
0.151
0.027
- 130 -
0.001
0.010
0.011
0.001
0.003
0.022
0.003
0.054
0.106
0.185
0.056
0.092
0.148
0.172
0.000
0.004
0.031
0.000
0.002
0.028
0.089
0.033
0.067
0.057
0.036
0.044
0.063
0.004
0.132
0.111
0.158
0.134
0.168
0.114
0.029
0.094
0.020
0.094
0.094
0.115
0.084
0.204
Maximum abundance
P
R2
0.923
0.929
0.789
0.882
0.983
0.720
0.872
0.856
0.402
0.078
0.855
0.587
0.230
0.054
0.835
0.980
0.441
0.818
0.986
0.683
0.289
0.273
0.166
0.222
0.269
0.221
0.242
0.924
0.068
0.049
0.025
0.071
0.032
0.052
0.302
0.142
0.373
0.081
0.146
0.089
0.107
0.048
0.000
0.000
0.003
0.001
0.000
0.005
0.001
0.001
0.026
0.111
0.001
0.011
0.053
0.140
0.002
0.000
0.022
0.002
0.000
0.006
0.045
0.044
0.070
0.055
0.047
0.055
0.050
0.000
0.132
0.153
0.193
0.135
0.177
0.148
0.048
0.088
0.033
0.121
0.090
0.116
0.104
0.166
Figure 4.3 Plots of significant relationships between modelled nutrient loadings and log10
abundance of HAB species in UK coastal waters.
4
4
Mean Dinophysis spp.
abundance
2
1
1
0
3000
6000
2000
Loading
6000
0
4
Mean P lima
abundance
0.5
3
1.0
0.5
3000
6000
0
9000
7000
14000
Loading
21000
28000
3
2
800
1000
3.0
2.0
0.0
0
7000
14000
Loading
21000
Loading
4
Si
1.0
0
600
9000
Mean P. mimimum
abundance
4.0
1
0
6000
5.0
Abundance
Abundance
4
1
400
3000
DIN
Maximum P. lima
abundance
4
2
200
0
35000
Loading
5
NO2
Maximum P. lima
abundance
0
1
Loading
4
3
2
0
0.0
0.0
NH4
Maximum P. lima
abundance
DIN
Abundance
1.5
1.0
0
10000 20000 30000 40000 50000 60000
Loading
2.0
NH4
Mean P. lima
abundance
Abundancee
Abundance
4000
Loading
2.0
1.5
2
0
0
9000
Si
1
0
0
Abundance
3
Abundance
2
Mean Dinophysis spp.
abundance
PO4
Mean Dinophysis
spp. abundance
3
Abundance
Abundance
3
4
NH4
28000
35000
0
10000
20000
30000
40000
50000
Loading
Si
Abundance
3
2
Maximum P. minimum
abundance
1
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
Alexandrium spp.
Dinophysis spp.
Pseudo-nitzschia spp.
Karenia mikimotoi
Prorocentrum lima
Prorocentrum minimum
Nutrient ratio
Mean abundance
P
R2
0.566
0.756
0.000
0.000
0.082
0.014
0.013
0.051
0.146
0.045
0.439
0.001
DIN:DIP
DIN:Si
DIN: DIP
DIN:Si
DIN: DIP
DIN:Si
DIN: DIP
DIN:Si
DIN: DIP
DIN:Si
DIN: DIP
DIN:Si
- 132 -
0.012
0.004
0.387
0.505
0.108
0.217
0.208
0.144
0.086
0.170
0.025
0.416
Maximum abundance
P
R2
0.972
0.759
0.015
0.003
0.412
0.091
0.051
0.043
0.157
0.053
0.360
0.001
0.000
0.004
0.200
0.312
0.025
0.110
0.134
0.154
0.082
0.160
0.035
0.429
Figure 4.4 Plots of significant relationships between modelled ratios of nutrient loadings and
log10 abundance of HAB species in UK coastal waters.
4
2
1
0
5
Mean Dinophysis spp.
abundance
3
2
DIN:Si
1
50
100
150
1
3
4
0
0.6
0.4
DIN:Si
0.2
0
30
40
1
2
Ratio
3
4
0
Abundance
Abundance
3
2
Maximum Dinophysis
spp. abundance
3
100
150
Ratio
3
4
5
Maximum Karenia
mik imotoi abundance
6
2
1
5
4
DIN:Si
3
2
1
0
50
2
7
Maximum Dinophysis
spp. abundance
DIN:Si
0
0
1
Ratio
4
DIN:DIP
1
DIN:Si
Ratio
5
4
2
0
0
50
4
3
1
0.0
20
3
Mean P. minimum
abundance
4
Abundance
DIN:DIP
10
2
5
Mean P. lima
abundance
0.8
2
0
1
Ratio
1.0
Mean Karenia
mik imotoi
abundance
Abundance
Abundance
2
Mean
Pseudo nitzschia
spp. abundance
Ratio
4
1
2
0
0
Ratio
3
3
1
0
0
DIN:Si
4
Abundance
DIN:DIP
Abundance
3
Mean Dinophysis spp.
abundance
Abundance
Abundance
4
0
0
1
2
3
4
0
Ratio
1
2
3
4
5
Ratio
7
Maximum P.mimimum
abundance
Abundance
6
5
4
DIN:Si
3
2
1
0
0
1
2
3
4
5
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
NH4
NO3
PO4
Loadings vs winter concentrations
Correlation coefficient
NH4
NO3
PO4
Degrees of Freedom
P
32
32
32
<0.001
<0.001
<0.001
0.844
0.937
0.915
measured
0.217
0.048
0.634
modelled
0.340
0.065
0.631
Degrees of Freedom
measured
36
35
36
modelled
30
39
30
P
measured
0.191
0.780
<0.001
modelled
0.057
0.728
<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
P
R2
Alexandrium spp.
NH4
DIN
DIN:DIP
DIN:Si
NH4
DIN
DIN:DIP
DIN:Si
NH4
DIN
DIN:DIP
0.752
0.409
0.607
0.868
0.049
0.045
0.279
0.057
0.065
0.092
0.876
Dinophysis spp.
Pseudo-nitzschia spp.
- 134 -
0.003
0.023
0.010
0.001
0.116
0.127
0.045
0.128
0.102
0.092
0.001
Maximum abundance
P
R2
0.677
0.500
0.467
0.738
0.081
0.020
0.860
0.055
0.116
0.203
0.643
0.006
0.015
0.021
0.004
0.092
0.169
0.001
0.130
0.075
0.053
0.008
Table 4.6 continued
Genus or Species
Nutrient or ratio
Mean abundance
P
R2
Pseudo-nitzschia spp.
Karenia mikimotoi
DIN:Si
NH4
DIN
DIN:DIP
DIN:Si
NH4
DIN
DIN:DIP
DIN:Si
NH4
DIN
DIN:DIP
DIN:Si
0.389
0.023
0.107
0.877
0.362
0.061
0.134
0.468
0.019
0.105
0.725
0.455
0.273
Prorocentrum lima
Prorocentrum minimum
Maximum abundance
P
R2
0.028
0.157
0.087
0.001
0.032
0.112
0.078
0.022
0.202
0.337
0.034
0.146
0.287
0.455
0.059
0.291
0.675
0.378
0.136
0.215
0.415
0.007
0.140
0.636
0.510
0.306
0.021
0.110
0.038
0.007
0.030
0.073
0.054
0.028
0.258
0.325
0.061
0.116
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
NH4
DIN
3
Mean Dinophysis spp.
abundance
2
1
Mean Dinophysis spp.
abundance
2
1
0
5
10
15
20
25
0
0
50
Concentration
150
200
250
300
0
2
1
0
200
300
15
20
1.5
DIN:Si
25
DIN:Si
1.0
0.5
0.0
100
10
3
Mean P. lima
abundance
Abundance
DIN
3
0
5
Concentration
2.0
Maximum Dinophysis
spp. abundance
Abundance
Abundance
100
Concentration
5
4
2
1
0
0
NH4
Mean Karenia
mik imotoi abundance
3
Abundance
Abundance
Abundance
3
2
Maximum P. lima
abundance
1
0
0
5
10
Concentration
15
20
Nutrient ratio
25
30
35
0
5
10
15
20
25
30
35
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
P
R2
Alexandrium spp.
NO3
NO2
TOxN
DIP
Si
TOxN:DIP
TOxN:Si
NO3
NO2
TOxN
DIP
Si
TOxN:DIP
TOxN:Si
NO3
NO2
TOxN
DIP
Si
TOxN:DIP
TOxN:Si
NO3
NO2
TOxN
DIP
Si
TOxN:DIP
TOxN:Si
NO3
NO2
TOxN
DIP
Si
TOxN:DIP
TOxN:Si
NO3
NO2
TOxN
DIP
Si
TOxN:DIP
TOxN:Si
0.176
0.159
0.179
0.091
0.216
0.957
0.249
0.057
0.016
0.041
0.049
0.657
0.328
0.070
0.314
0.947
0.251
0.804
0.242
0.961
0.895
0.240
0.041
0.144
0.065
0.116
0.688
0.254
0.189
0.911
0.140
0.112
0.677
0.502
0.094
0.928
0.274
0.980
0.085
0.027
0.949
0.198
Dinophysis spp.
Pseudo-nitzschia spp.
Karenia mikimotoi
Prorocentrum lima
Prorocentrum minimum
- 136 -
0.039
0.042
0.039
0.058
0.032
0.000
0.031
0.075
0.118
0.088
0.078
0.004
0.023
0.075
0.022
0.000
0.029
0.001
0.028
0.000
0.000
0.036
0.105
0.055
0.083
0.061
0.005
0.037
0.038
0.000
0.049
0.054
0.004
0.011
0.067
0.001
0.107
0.000
0.197
0.305
0.000
0.134
Maximum abundance
P
R2
0.358
0.102
0.395
0.237
0.104
0.580
0.569
0.014
0.005
0.009
0.021
0.377
0.840
0.037
0.530
0.793
0.443
0.993
0.260
0.851
0.838
0.545
0.162
0.359
0.198
0.239
0.837
0.351
0.232
0.658
0.164
0.177
0.917
0.565
0.038
0.725
0.202
0.766
0.072
0.022
0.797
0.290
0.018
0.056
0.016
0.029
0.054
0.007
0.008
0.121
0.156
0.140
0.105
0.016
0.001
0.097
0.009
0.002
0.013
0.000
0.026
0.001
0.001
0.010
0.051
0.022
0.041
0.035
0.001
0.025
0.032
0.004
0.044
0.039
0.000
0.008
0.101
0.012
0.143
0.008
0.213
0.322
0.006
0.093
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
3
Mean Dinophysis spp.
abundance
2
1
0
100
200
300
0
0
2
Concentration
4
6
0
4
1
Mean P. minimum
abundance
2
1
0
2
2.5
10
20
30
0
2
1
0
0
200
Maximum Dinophysis
s pp. abundance
3
1
300
4
6
8
0
3
1
0
0
20
30
Maximum Dinophysis
spp. abundance
2
1
8
TOxN: Si
4
Abundance
Abundance
Maximum
P. minimum
abundance
6
5
TOxN: Si
4
10
4
Concentration
5
Si
0
2
Concentration
4
2
DIP
2
0
2
Concentration
3
3
1
0
5
300
Maximum Dinophysis
spp. abundance
4
Abundance
Abundance
2
200
5
NO2
4
Maximum Dinophysis
spp. abundance
100
100
Concentration
5
NO3
0
2
Concentration
5
3
3
0
0
Concentration
4
Maximum Dinophysis
s pp. abundance
1
0
1.5
8
TOxN
4
Abundance
Abundance
Mean Karenia mik imotoi
abundance
1
6
5
3
0.5
4
Si
3
0
2
Concentration
NO2
Abundance
8
Concentration
4
2
Mean Dinophysis spp.
abundance
2
1
0
0
Abundance
Abundance
Mean Dinophysis spp.
abundance
2
1
Abundance
DIP
3
Abundance
Abundance
3
Maximum P. lima
abundance
3
2
1
0
0
Concentration
10
20
Nutrient ratio
30
40
0
10
20
30
40
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
or Species
Lough
Foyle (1)
Carlingford
Lough (1)
Alexandrium spp.
Dinophysis acuminata
nt
nt
nt
nt
D. acuta
D. norvegica
D rotundata
Dinophysis spp.
nt
nt
nt
nt
nt
nt
nt
nt
Pseudo-nitzschia spp.
Karenia mikimotoi
Prorocentrum lima
P. minimum
nt
nt
nt
nt
nt
nt
nt
nt
Location
Strangford
Lough (4)
nt
negative trend
(P = 0.023)
nt
nt
nt
negative trend
(P = 0.034)
nt
nt
nt
nt
Dundrum
Bay (2)
Larne
Lough (1)
nt
nt
nt
nt
nt
nt
nt
nt
nt
nt
nt
nt
nt
nt
nt
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
59
0 to 1
2 to 99
100 to 999
1000 to 9999
10000 to 439200
0 to 1
2 to 99
100 to 999
1000 to 28030
Latitude
57
55
53
51
49
61
59
Alexandrium spp.
Dinophysis spp.
0 to 1
2 to 99
100 to 999
1000 to 9999
10000 to 99999
100000 to 690000
0 to 1
2 to 99
100 to 999
1000 to 9999
10000 to 99999
100000 to 418100
Latitude
57
55
53
51
49
Karenia mikimotoi
Pseudo-nitzschia spp.
61
0 to 1
2 to 99
100 to 999
1000 to 9999
10000 to 99999
100000 to 420000
0 to 1
2 to 99
100 to 302
59
Latitude
57
55
53
51
Prorocentrum lima
49
10.5
8.7
6.9
5.1
Prorocentrum minimum
3.4
1.6
0.2
Longitude
2.010.5
8.7
6.9
5.1
3.4
Longitude
- 140 -
1.6
0.2
2.0
Figure 4.7 continued The maximum abundance of HAB species in UK and Irish coastal waters.
61
59
0 to 1
2 to 99
100 to 999
1000 to 9999
10000 to 99999
100000 to 999999
1000000 to 17011000
0 to 1
2 to 99
100 to 999
1000 to 9999
10000 to 83870
Latitude
57
55
53
51
Dinophysis spp.
Alexandrium spp.
49
61
59
0 to 1
2 to 99
100 to 999
1000 to 9999
10000 to 99999
100000 to 999999
1000000 to 1786000
0 to 1
2 to 99
100 to 999
1000 to 9999
10000 to 99999
100000 to 999999
1000000 to 6058000
Latitude
57
55
53
51
49
61
59
Karenia mikimotoi
Pseudo-nitzschia spp.
0 to 1
2 to 99
100 to 999
1000 to 9999
10000 to 99999
100000 to 999999
1000000 to 4125000
0 to 1
2 to 99
100 to 3521
Latitude
57
55
53
51
Prorocentrum lima
49
10.5 8.9
7.3
5.6
Prorocentrum minimum
4.0
2.4
0.8
0.9
10.5
8.7
6.9
5.1
3.4
Longitude
Longitude
- 141 -
1.6
0.2
2.0
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 yvariable 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
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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.
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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
Modelled loadings
Dinophysis spp.
Prorocentrum lima
Prorocentrum minimum
Winter concentrations
Dinophysis spp.
Karenia mikimotoi
Prorocentrum minimum
Modelled loadings ratios
Dinophysis spp.
Pseudo-nitzschia spp.
Karenia mikimotoi
Nutrient
Mean abundance
P
R2
NH4
PO4
Si
NH4
DIN
NO2
Si
0.020 -ve
0.039 -ve
0.032 +ve
0.045 -ve
0.037 -ve
0.185
0.148
0.172
0.158
0.168
0.027 +ve
0.204
NH4
DIN
NO3
NO2
TOxN
DIP
NH4
NO2
Si
0.049 -ve
0.045 -ve
0.116
0.127
0.016 -ve
0.041 -ve
0.049 -ve
0.023 -ve
0.041 -ve
0.027 -ve
0.118
0.088
0.078
0.157
0.105
0.305
DIN: DIP
DIN:Si
DIN:Si
DIN: DIP
DIN:Si
DIN:Si
DIN:Si
0.000 +ve
0.000 -ve
0.014 -ve
0.013 +ve
0.387
0.505
0.217
0.208
0.045 -ve
0.001 -ve
0.170
0.416
Prorocentrum lima
Prorocentrum minimum
Winter concentration ratios
Dinophysis spp.
TOxN:Si
Prorocentrum lima
DIN:Si
TOxN:Si
0.019 -ve
0.202
Maximum abundance
P
R2
0.025 -ve
0.032 -ve
0.049 -ve
0.048 +ve
0.193
0.177
0.153
0.166
0.020 -ve
0.014 -ve
0.005 -ve
0.009 -ve
0.021 -ve
0.169
0.121
0.156
0.140
0.105
0.022 -ve
0.322
0.015 +ve
0.003 -ve
0.200
0.312
0.043 -ve
0.154
0.001 -ve
0.429
0.037 -ve
0.007 -ve
0.038 -ve
0.097
0.258
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 nonrandom distribution of nutrient enrichment.
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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
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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 nonparametric 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)
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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.
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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.
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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.”
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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
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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 filterfeeders 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
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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
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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.
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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
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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
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zone of ≈ 5 m 37 ) 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,
NO3 and NO2) and chlorophyll (mg m-3) 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
DAIN ( M)
25
20
15
10
Tolo Harbour
5
Victoria Harbour
0
1
2
3
4
5
6
7
Month
8
9
10
11
12
25
-3
Chlorophyll (mg m )
B
Tolo Harbour
Victoria Harbour
20
15
10
5
0
1
2
3
4
5
6
7
8
9
10
11
12
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
factor to estimate Kd in order to calculate euphotic zone depth using the equation Iz = I0e-kdh, 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
the annual median nitrogen (as nitrate, NO3-) for Tolo Harbour as 0.135 mg l-1 (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 Enhancement 38
(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.
Volume
(km3)
0.6
Mean depth
(m)
12
Residence time
(d)
15-30
Carlingford
Lough
Loch Striven
0.5
8
14 - 26
0.6
37
8 - 50
Temperature
(ºC)
13 – 32.0
(mean 24.3)
6 – 16.6
Mean (12.0)1
7 - 14
Seto Inland
Sea
Eastern Irish
Sea
816
37
438
8 – 26
480
15
40 – 480
5 - 17
Location
Tolo Harbour
Water column
stability
thermal stratification
Intermittent salinity
stratification
Seasonal thermohaline
seasonal thermal
stratification
intermittent salinity
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
loading (t)
Residence (d) used to
calculate ECE time
Tolo Harbour
750
(total N)
1,311
(NO3 + NH4)
181,000
(total N)
30,684
(total N)
20
4.8
Lam & Ho (1989)
10
5.6
Taylor et al. (1999)
438
19.0
Takeoka (1997)
300
4.4
UK RID surveys
Carlingford
Lough
Seto Inland Sea
Eastern Irish
Sea
Equilibrium
Loading data
concentration
source
Enhancement (µM)
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
Striven supports a larger spring bloom (89 mg chlorophyll m-3 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
-3
Chlorophyll (mg m )
12
Data from Douglas
A
Data from Ball et al
10
8
6
4
2
0
01-Jan
90
-3
Chlorophyll (mg m )
80
22-Feb
15-Apr
06-Jun
28-Jul
B
18-Sep
09-Nov
31-Dec
23rd September
181 mg m -3
70
60
50
40
30
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 salinity 39 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
Scotland
Firth
of Clyde
55.0
Belfast
54.5
North
Channel
Solway Firth
IOM
Western
Irish
Sea
Latitude
54.0
Cumbria
North eastern
Irish Sea
South eastern
Irish Sea
R Ribble
6
Liverpool 2
Bay
51
53.5
4LB 3
R Mersey
R Dee
53.0
St George's
Channel
52.5
Wales
52.0
51.5
51.0
7.0
Celtic
Sea
6.5
Bristol Channel
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 SuoNada.”
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 buoyancy 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
Exchange
Characteristic ecohydrodynamic
features
Tolo Harbour
Persistent thermal stratification and
slowly flushed.
Short seasonal stratification and
rapidly flushed.
Tidally stirred, with infrequent
thermo-haline stratification and
moderately flushed.
High frequency of HABs
Seasonal thermo-haline stratification
and moderately flushed.
Benign summer blooms and
occasional HABs
Thermal stratification and numerous
frontal boundaries
Region of freshwater influence,
tidally stirred and intermittent
thermohaline stratification
High frequency of HABs
Victoria Harbour
Carlingford Lough
Loch Striven
Harmful and benign blooms
Low frequency of HABs
Low frequency of benign summer
blooms
Regional Seas
Seto Inland Sea
Eastern Irish Sea
Infrequent occurrence of HABs
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
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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 turbidity (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).
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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.
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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.
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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
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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
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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 Pseudonitzschia 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 Pseudonitzschia 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 Pseudonitzschia 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 Pseudonitzschia in thin layers, have the potential to directly affect the toxicity of shellfish; and there is
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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
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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
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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
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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.
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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.
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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:
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“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.
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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, nearshore, 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 thermohaline 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.
- 188 -
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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
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Annex II
Pictures of selected species of phytoplankton
Dinoflagellates
Alexandrium tamarense (SEM 40 ) (AFBI)
Alexandrium tamarense (AFBI)
(22 – 44 x 20 - 36 µm, l x w)
Dinophysis acuminata (AFBI)
(38 – 58 µm long)
40
Dinophysis acuta (L Naustvoll)
(54 – 85 µm long)
Scanning Electron Micrograph showing the thecal plates
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Prorocentrum lima (SAMS)
(32 – 50 x 20 – 28 µm, l x w)
Karenia mikimotoi (Cefas)
(24 – 40 x 17 – 32 µm, l x w)
Prorocentrum minimum (AFBI)
(14 – 22 x 10 – 15 µm, l x w)
P. minimum (SEM) (AFBI)
Noctiluca scintillans (A Alazri)
(diameter 200 – 2,000 µm)
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Diatoms
Pseudo-nitzschia spp. (AFBI)
(33 – 160 µm long)
Chaetoceros brevis (E Capuzzo)
(7 – 40 µm diameter)
Leptocylindrus danicus (E Capuzzo)
(5 – 16 µm wide)
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Microflagellates
Chattonella sp. (L Naustvoll)
(30 – 50 µm long)
Chrysochromulina polylepis (L Naustvoll)
(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.)
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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 3-11, 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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