Climate Change and Marine and Freshwater Toxins [2nd, completely revised and extended Edition] 9783110625738, 9783110622928

Citation preview

Botana, Louzao, Vilariño (Eds.) Climate Change and Marine and Freshwater Toxins

Also of Interest Environmental Toxicology Botana (Ed.),  ISBN ----, e-ISBN ----

Nanomaterials for Water Remediation Mishra, Hussain, Mishra (Eds.),  ISBN ----, e-ISBN ----

Aquatic Chemistry, for Water and Wastewater Treatment Applications Lahav, Birnhack,  ISBN ----, e-ISBN ----

Drinking Water Treatment, An Introduction Worch,  ISBN ----, e-ISBN ----

Food Contamination by Packaging, Migration of Chemicals from Food Contact Materials de Quirós, Cardama, Sendón, Ibarra,  ISBN ----, e-ISBN ----

Climate Change and Marine and Freshwater Toxins Edited by Luis M. Botana, M. Carmen Louzao and Natalia Vilariño 2nd Edition

Editors Prof. Luis M. Botana Universidad de Santiago de Compostela Facultad de Veterinaria Departamento de Farmacología 27002 Lugo, Spain [email protected] Prof. M. Carmen Louzao Universidad de Santiago de Compostela Facultad de Veterinaria Departamento de Farmacología 27002 Lugo, Spain [email protected] Prof. Natalia Vilariño Universidad de Santiago de Compostela Facultad de Veterinaria Departamento de Farmacología 27002 Lugo, Spain [email protected]

ISBN 978-3-11-062292-8 e-ISBN (PDF) 978-3-11-062573-8 e-ISBN (EPUB) 978-3-11-062302-4 Library of Congress Control Number: 2020950014 Bibliographic information published by the Deutsche Nationalbibliothek The Deutsche Nationalbibliothek lists this publication in the Deutsche Nationalbibliografie; detailed bibliographic data are available on the Internet at http://dnb.dnb.de. © 2021 Walter de Gruyter GmbH, Berlin/Boston Cover image: abadonian/iStock/Getty Images Plus Typesetting: Integra Software Services Pvt. Ltd. Printing and binding: CPI books GmbH, Leck www.degruyter.com

Preface Climate change and its far-reaching impacts compel us to question the value we give to nature and the human relationship with it. Water touches many subject areas that are important in our daily lives. At some point, it influences what we eat and drink, and how we access it. Therefore, we decided to edit a book on climate change and water toxins that includes climate trends and effects, physicochemical measurements, water quality parameters, marine and freshwater toxins, toxin detection, phytoplankton and zooplankton, invertebrates and fish. There is no historical record to compare the amounts of toxins that are existent now and a century or more ago. Toxins are identifiable as a result of modern science, and thereby their presence, structure or levels in food have only been known for a short time. The use of mass spectrometers is rather recent, and the existence of certified standards only goes back a few years. Therefore, it is very complex to establish a solid link, using the scientific method, between climate change and toxins. But it is clear that something is happening – not only because modern technology allows us to track the changes easily but also because the trend is that more and different toxins are appearing in new locations and products. Although climate change is frequently related to extreme weather episodes and rising sea levels in the media, a lesser known fact is that new toxins will appear in areas and products where they presently do not occur. Despite the fact that scientific evidence may not always be available to prove or disprove perceived potential harms of climate change and their links with toxins, this book offers quantitative compelling evidence of the many complex interactions that must be considered from primary toxin producers up the food chain to humans. In the case of marine toxins, although ballast water, international trade and so on may be a source of new intoxications and blooms, it is very clear that some regions are hot spots for many compounds. Likewise, eutrophication of lakes is a source of cyanobacterial blooms. The United States had never had a diarrheic episode until Texas witnessed one a few years ago. Europe had never had a tetrodotoxin intoxication from shellfish until a few years ago; ciguatoxin intoxications are becoming frequent after ingestion of fish from the Southern European Atlantic Ocean; and aerosols with ostreocin from Gambierdiscus are now a problem in Mediterranean beaches every year. A similar problem is being observed in freshwater, as the expansion of cyanobacteria and their toxins has become a worldwide problem; this adds to the deleterious effect of human pollution in drinking water. Something is happening that was not previously reported and may be explained by increased water temperatures in both lakes and seas. This book intends to cover the main aspects of the possible relation between climate change and freshwater and marine toxins: prediction models and management of harmful algal blooms; influence on food security and food production; legislation; drinking water and cyanobacteria blooms; and sex change in toxin vectors. https://doi.org/10.1515/9783110625738-202

VI

Preface

This last topic, sex change, serves as an introduction to a new area of research – the role of climate change in basic physiological processes. Very little information is currently available on this subject. This book has brought together a group of international experts. Contributing authors expand the framework of possibilities for appropriate assessment of climate change impacts on marine and freshwater toxins, which in turn directly impacts the natural environment, human health and sustainability. The book is an excellent introduction to this complex topic or a useful supplement to courses in the field of ecotoxicology. In short, it is a must-read book for all who are interested in toxins and how climatic conditions can modify them – from the general public or students to toxicologists, food technologists, pharmacologists, analytical chemists, ecologists, biologists, veterinarians and physicians. Last, and by no means least, we thank all the authors. They were not only very generous with their time, and were also bold enough to commit to write a chapter on an especially difficult topic and use their prestigious names in their chapters. For this, we are greatly thankful to all of them. We hope the book helps in understanding the potential risks caused by climate in a particularly sensitive area: food and drinking water.

Contents Preface

V

Josefino C. Comiso 1 Variability and trends of global sea ice cover and sea level: effects on physicochemical parameters 1 Begoña Espiña, Marta Prado, Stephanie Vial, Verónica C. Martins, Soraia P.S. Fernandes, Marilia B. dos Santos, Laura M. Salonen and Raquel B. Queirós 2 New techniques in environment monitoring 35 Mikko Nikinmaa and Katja Anttila 3 Responses of marine animals to ocean acidification

107

Shauna Murray, Uwe John, Henna Savela and Anke Kremp 4 Alexandrium spp.: genetic and ecological factors influencing saxitoxin production and proliferation 133 Laura T. Kelly, Jonathan Puddick, Hugo Borges, Daniel R. Dietrich, David P. Hamilton and Susanna A. Wood 5 Potential effects of climate change on cyanobacterial toxin production 167 Gustaaf M. Hallegraeff 6 Harmful marine algal blooms and climate change: progress on a formidable predictive challenge 195 Elke S. Reichwaldt, Som Cit Sinang and Anas Ghadouani 7 Global warming, climate patterns and toxic cyanobacteria

209

Panagiota Katikou 8 Human impact in Mediterranean coastal ecosystems and climate change: emerging toxins 253 Hazel Farrell, Gurjeet S. Kohli and Shauna A. Murray 9 Gambierdiscus, the cause of ciguatera fish poisoning: an increased human health threat influenced by climate change 303

VIII

Contents

Mireille Chinain, Clémence M.i. Gatti, Hélène Martin-Yken, Mélanie Roué and H. Taiana Darius 10 Ciguatera poisoning: an increasing burden for Pacific island communities in light of climate change? 369 Dani J. Barrington, Xi Xiao, Liah X. Coggins and Anas Ghadouani 11 Control and management of harmful algal blooms 429 Patricia M. Glibert, Arthur H. W. Beusen, Alexander F. Bouwman, JoAnn M. Burkholder, Kevin J. Flynn, Cynthia A. Heil, Ming Li, Chih-Hsien (Michelle) Lin, Christopher J. Madden, Aditee Mitra, William Nardin, Greg Silsbe, Yang Song and Fan Zhang 12 Multifaceted climatic change and nutrient effects on harmful algae require multifaceted models 473 Joaquín Espinosa, Sara Silva-Salvado and Óscar García-Martín 13 Global climate change profile and its possible effects on the reproductive cycle, sex expression and sex change of shellfish as marine toxin vectors 519 M. Carmen Louzao, Natalia Vilariño and Luis M. Botana 14 Effects on world food production and security 579 Natalia Vilariño, M. Carmen Louzao, María Fraga and Luis M. Botana 15 From science to policy: dynamic adaptation of legal regulations on aquatic biotoxins 607 Index

655

Josefino C. Comiso

1 Variability and trends of global sea ice cover and sea level: effects on physicochemical parameters 1.1 Introduction The rapid decline in the Arctic summer ice cover minimum and the acceleration of sea-level rise, as reported in recent years [1–3], have gained a great deal of attention and are regarded among the most visible signals of anthropogenic global warming. Both phenomena have been linked to climate change either directly or indirectly and are expected to cause profound changes in the physicochemical characteristics of the polar and extrapolar regions. Historically, the high latitude regions have received little interest mainly because of general inaccessibility, harsh weather conditions and the paucity of data. The advent of satellite remote sensing has completely changed this, and our ability to monitor polar regions and especially sea ice cover, other components of the cryosphere and sea-level rise has been vastly improved. In particular, polar orbiting satellite data have yielded four decades of consistent and continuous global data sets at a temporal resolution of at least twice daily. But more importantly, the data have yielded strong evidence that dramatic changes related to climate are occurring in the polar regions. The yearly Arctic summer ice cover minimum has been studied and used to indicate that the perennial ice cover has been rapidly declining – even with only approximately 22 years (i.e., 1978 to 2000) of satellite data [4]. By perennial ice, we mean the thick ice type that normally survives the spring and summer melt. It is the mainstay of the Arctic sea ice cover and is known to have been in existence for at least 1,450 years [5]. Following a dramatic decline in the Arctic summer sea ice cover in 2007 [1], the extent of the perennial ice was a record low in 2012 [6] and was found to be less than half of its extents observed in the 1980s. This has led to a realization that the Arctic region has been changing fast; this has ignited many international and national research projects in the region. While a sea ice cover decline has been expected from modeling studies, the rate of decline as projected by the models is significantly less than that actually observed from satellite data [7]. This means that there are still gaps in our knowledge of the physics of the Arctic climate system. The observed decline is consistent with the amplification of global warming in the Arctic by more than three times as has been reported [8]. Such amplification is in part the result of a phenomenon called “ice-albedo feedback” that is associated with the decline of the summer ice cover [9]. Such warming has affected the rest of the cryosphere in the region, including Josefino C. Comiso, Earth Sciences Division, NASA Goddard Space Flight Center, Maryland, USA https://doi.org/10.1515/9783110625738-001

2

Josefino C. Comiso

the glaciers in North America and the vast ice sheet of Greenland [10]; the latter has a sea-level equivalence of greater than 7 m. The observed physical changes in the Southern Hemisphere is different from those in the Arctic and appear counterintuitive but not totally unexpected, considering that the impact of global warming around the globe is not uniform [11]. Sea ice cover is observed to be expanding and some cooling is observed in large areas of the Antarctic region [12, 13]. Quantitative estimates of the rate of loss of mass observed due to melting of the massive ice sheet in Antarctica, which has a sea-level equivalence of more than 60 m, are also of interest but results from different investigators have been inconsistent [10]. This chapter presents an overview of the current state of the global sea ice cover, changes in surface temperature, related rise in sea level and associated changes in the physicochemical characteristics of the marine environment and global climate. Sea ice is already a part of the ocean and causes a negligible increase in seawater level when it melts. However, the loss of sea ice in recent years, especially in the spring and summer, has contributed to the aforementioned amplified warming in the Arctic that in turn has caused more land ice to melt and thus a higher rate of sea-level rise. It will likely cause a change in the primary productivity of the region as well. The direct impacts of changes in the global sea ice cover on the physicochemical characteristics of the affected regions are expected to be profound, but specific details are basically unknown and still the subject of many modeling and observational studies. Most studies focus on the impact of climate change in general on marine ecosystems [14], but the impact on polar marine ecosystems has become the subject of strong interest because of the rapidly retreating Arctic ice cover [15]. A significant increase in sea level is expected to cause serious negative changes on the global marine ecosystem, but most studies concentrate mainly on impacts that primarily occur in coastal regions [16] where the effects are most visible. Because of rapid changes in the polar environment in recent years, this chapter has been updated from a previous version to include data up to 2019. The extended data record provides the means to better interpret the observed phenomena and enhances our current understanding of the physics of the system. The use of updated data also allows for an assessment of previous forecasts’ accuracy that relied on numerical models or statistical studies.

1.2 Variability and trends of global sea ice The sea ice cover has been considered a key component of the Earth’s climate system. Because of its high albedo and good thermal insulating property, it is very effective in limiting heat and salinity fluxes between the ocean and the atmosphere. Sea ice also redistributes surface salinity and alters surface density; this causes vertical circulation that enables upper layers of the oceans to be replenished with nutrients, oxygen and other chemicals. The ice-covered regions make up three of only

1 Variability and trends of global sea ice cover and sea level

3

four regions where deep ocean convection has been observed worldwide. They have also been the primary source of bottom water that is an essential component of global ocean thermohaline circulation. Latent heat and sensible heat polynyas are also formed in ice-covered regions; this causes an alteration in the physical and chemical properties of the ocean [17]. Sea ice melt in the spring and summer also causes the formation of a stable melt surface layer that is exposed to abundant sunlight and serves as a platform for efficient photosynthesis. This phenomenon makes the region a site for phytoplankton blooms and high primary productivity during the spring and summer [18]. Sea ice also covers a large fraction (about 6%) of the global oceans and can affect shipping, fisheries and mineral exploration. Among the most important tools that have been used to study the large-scale characteristics, variability and trends of global sea ice cover are satellite passive microwave sensors. The first imaging system was the Nimbus-5/ESMR that was launched in December 1972 and was the first to reveal the true extent and variability of global sea ice cover at a temporal resolution of about 3 days. Because of the large contrast in the brightness temperature of sea ice-covered and ice-free ocean, the sensor was able to provide the extent and general characteristics of the sea ice cover [19]. However, with only one channel available at 19 GHz and horizontal polarization, the data could only provide rough estimates of the ice concentration within the pack because of varying surface emissivity associated with different ice types and snow cover conditions. Accurate, consistent and continuous monitoring of sea ice cover started with the Nimbus-7 scanning multichannel microwave radiometer which was launched in October 1978 and was succeeded by a series of the defense meteorological satellite program (DMSP)/special scanning microwave imager that started in July 1987 and continued up to the present. The two sensors are both dual polarized, multifrequency and conically scanning systems for consistent coverage of the surface at a resolution of about 25 km. An even more capable system called advanced microwave scanning radiometer (AMSR-E) was launched on board the EOS/Aqua satellite in May 2002; it provides more accurate brightness temperature data at a significantly higher resolution [20]. AMSR-E provided high quality sea ice data until it suffered instrumental problems and was turned off in 2011. It was succeeded by AMSR-2, which was launched on board global climate observation mission (GCOM-W) by Japan aerospace exploration agency (JAXA) in 2012. Analyses of time series of sea ice data from various passive microwave sensors have been reported in several publications [12, 13, 21–24]. Different techniques were used to retrieve sea ice concentrations and therefore ice extent and ice area. However, the results on seasonal and interannual variability of sea ice have been generally consistent [25, 26]. The trends reported for different time periods are generally different – albeit slightly – because of the relatively short record length and the large interannual variability of ice cover. Color-coded maps of ice concentration averages during maximum and minimum extents for each year from November 1978 to December 2019 are presented in Figure 1.1.

4

Josefino C. Comiso

(a) NH minimum average with 2019 minimum contour

(b) NH maximum average with 2019 maximum contour

(c) SH minimum average with 2019 minimum contour

(d) SH maximum average with 2019 maximum contour

0%

25%

50%

75%

100%

Ice concentration Figure 1.1: Color-coded ice concentration average maps during maximum and minimum extents in the (a, b) Northern Hemisphere and (c, d) Southern Hemisphere. The red contours represent averages minimum or maximum data in 2019.

1 Variability and trends of global sea ice cover and sea level

5

Because of large interannual variability in the location of the ice edges, the images show smearing at the ice margins. In the Northern Hemisphere, the sea ice cover in the Arctic Basin is confined by surrounding land areas, but the sea ice reaches the peripheral seas and can go as far south as 44° N. Note how one year can be so different from the average. The contour for maximum extent in 2019 is much more expansive than average in the Sea of Okhotsk (top of Figure 1.1(b)) while much less expansive in the Bering Sea (middle left). In the Southern Hemisphere, the sea ice cover surrounds the Antarctic continent and is generally symmetric in winter and on the average covers the Southern Ocean up to latitudes near 55° S. The sea ice in the Arctic is exposed to colder temperatures and is generally thicker than in the Antarctic. However, the ice in the Antarctic is more expansive, partly because no land boundary at the ice edges limits its growth. The trends in ice cover in the two hemispheres are different: it is declining significantly in the Northern Hemisphere while it is expanding, albeit modestly, in the Southern Hemisphere. To some extent, this phenomenon is caused by differences in the geographical environment of the ice cover and the climate system of the two regions. It is, however, interesting to note that when the monthly values from the two hemispheres are added together for each month the sea ice cover shows a relatively uniform distribution with a slightly negative trend of about 2% per decade.

1.2.1 Arctic region The large-scale variability of the sea ice cover has been quantified using the terms ice extent and ice area. By definition, ice extent represents the integral sum of the area of all data elements in the map with ice concentration greater than 15%. Ice area represents the sum of the products of the area of each data element and the ice concentration. In particular, ice extent represents the total area of the ice-covered region and provides the means to study how the fraction of the ocean with ice cover on it is changing. Ice area, on the other hand, provides the actual area covered by sea ice and the means to estimate the total ice volume assuming that the average thickness is known. The estimated ice volume can in turn be used for mass balance studies. A plot of monthly sea ice extents in the Northern Hemisphere from November 1978 to November 2019 is presented in Figure 1.2(a). The plot, which is an update of those reported previously [13, 27, 28], shows large seasonality of the sea ice cover with the winter extent as high as 16 × 106 km2 and the summer extent as low as 3.5 × 106 km2. The plot also shows a slight increase in the amplitude of the yearly seasonality from the 1978 to 2006 period to the more recent 2007 to 2019 period. It is apparent that the Arctic sea ice cover has become more seasonal in the last 13 years because the sea ice cover at the end of the summer is declining more rapidly than sea ice during the winter. To assess interannual variability and trends more quantitatively, monthly anomalies in the ice extents are presented in Figure 1.2(b). The anomalies were estimated

6

Josefino C. Comiso

Northern Hemisphere 20 (a) Ice extent

Extent (106 km2)

15

10

5

0 3 (b) Ice extent anomaly 2

Extent (106 km2 )

1 0 –1 –2 –3 –4

Trend: –57834.2 ± 1539.6 km2/year (–4.5 ± 0.1 %/dec )

1980

1990

2000

2010

2020

Year Figure 1.2: Plots of (a) monthly averages and (b) monthly anomalies of sea ice extent in the Northern Hemisphere for the period from November 1978 to October 2019.

by subtracting the climatological monthly averages (using averages of data from 1978 to 2019) from each monthly average. The plot shows high values in 1980 and 1996 and unusually low values in 2007, 2012 and 2019. The period from 1997 to 2006 also shows a moderate and steady decline but no unusual yearly changes. After 2006, the interannual variability became very strong in part because of inconsistent and relatively low values in the summer ice extent. In 2012, the monthly anomaly went down to a record low value of almost 3 × 106 km2. The results of a linear regression analysis also indicate an average decline of 4.5% per decade from 1978 to 2019. To gain better insight into the changes in sea ice cover, the maximum and minimum extents and areas of annual sea ice cover are presented in Figure 1.3(a) and (b),

1 Variability and trends of global sea ice cover and sea level

7

Figure 1.3: Plots of yearly maximum (blue), minimum (red), mean (green) and seasonal ice (gold) of (a) sea ice extent and (b) sea ice area from 1979 to 2019 in the Northern Hemisphere.

respectively. The 5-day running average of daily extents was used to estimate the maximum and minimum for each year. The running average is used to minimize the effect of daily variability in the extent due to temporal changes in wind direction and other factors. The maximum extent is shown to be declining at the rate of 2.7% per decade while the minimum extent is declining at the rate of 11.2% per decade. The ice area is declining at a slightly different rate with the maximum and minimum trends being 2.6% and 11.7% per decade, respectively. The actual numerical changes in area and extent and corresponding statistical errors are listed in the figures. The results show that although both cases show negative changes, there is a large difference in the change of ice extent during the winter/growth period and the change in ice extent during the summer/melt period. The change in winter is usually influenced

8

Josefino C. Comiso

mainly by surface temperature and wind circulation, while the change in summer is primarily controlled by the thickness of the ice floes and their ability to survive the summer melt. A fraction of the thicker ice floes are advected out of the region mainly through Fram Strait, but the interannual changes during the satellite era are relatively minor according to recent studies [29, 30]. The change in minimum ice cover represents a change in perennial ice cover, which is ice that survives the summer melt. The perennial ice is known to be the mainstay of the Arctic sea ice cover and assumed to be present every summer since it has been as been observed in situ for at least 1,450 years [5]. However, a continuation of a decline of 11% to 12% per decade implies that summer ice could disappear during the twenty-first century. It may actually disappear sooner than expected, since actual data show more rapid decline than what modeling studies predict [7]. Studies of changes in the thickness of the ice have been undertaken using a combination of submarine, mooring and satellite data; there is a consensus that the average thickness is declining as well [31]. A quantitative assessment of the interannual variability of Arctic sea ice cover is presented in Figure 1.3. The yearly averages (in green) do not show the fluctuations in the monthly plots, but the yearly trends in the ice extent and ice area are −4.43 ± 0.23 and −4.56 ± 0.27% per decade, respectively, which are consistent with results using monthly anomalies (see Figure 1.2(b)). The values derived from monthly anomalies are usually used because of higher statistical accuracy. The differences between the maximum extents (in winter) and the minimum extent (in the previous summer) provides the mean to assess how the extent of seasonal ice has been changing; these are shown (in yellow) in Figure 1.3. The interannual changes are relatively small except in 2007 and 2012. The trends for ice extent and ice area for seasonal ice in the Arctic are 5.53 ± 1.01% and 5.20 ± 0.94% per decade, respectively. These values indicate that the pan-Arctic sea ice cover is becoming more and more seasonal. The trends in ice cover are not uniform spatially and can be very different for different regions and seasons as depicted in the color-coded trend maps shown in Figure 1.4. In winter and spring, the significant negative trends occur in the peripheral seas and especially in the Sea of Okhotsk, Baffin Bay, Barents Sea and Greenland Sea. It is interesting to note that the positive trends in winter and spring in the Bering Sea reported previously are now mainly negative suggesting significant warming in this region in recent years. The trends in ice cover in the Central Arctic are almost all zero because the ice cover is fully consolidated most of the time during this period. During the summer and autumn, the trends are strongly negative in the areas where the ice has retreated the most during the 1978 to 2019 period.

1 Variability and trends of global sea ice cover and sea level

Northern Hemisphere seasonal ice concentration trends 1979–2019 (a) Winter (DJF)

(b) Spring (MAM)

90°E

90°W

60°N 50°N

(c) Summer (JJA)

(d) Autumn (SON)

–2.4 –2.0 –1.6 –1.2 –0.8 –0.4 0.0 0.4 0.8 1.2 1.6 2.0 2.4 % Ice concentration/year Figure 1.4: Color-coded trend maps of sea ice concentration for (a) winter, (b) spring, (c) summer and (d) autumn using satellite data from 1979 to 2019. Trends are indicated in yearly change in area (km) and in percentage change.

9

10

Josefino C. Comiso

1.2.2 Antarctic region The monthly variability in the extent of Antarctic sea ice cover as observed using passive microwave data is depicted in Figure 1.5(a). It is apparent that ice cover in the Southern Ocean is more extensive in the winter but less extensive in the summer than those in the Arctic region [32]. The ice cover is thus primarily seasonal, and perennial ice is almost all of the second-year ice type because it tends to be advected out of the perennial ice regions every winter. In the first two decades, the ice seasons were generally very similar and typically would range in extent from 3.5 to 19 × 106 km2. In 2014, however, the range shifted up to about 4 to 20 × 106 km2 but shortly after that it went down to 2.5 to 18.5 × 106 km2.

Figure 1.5: Plots of (a) monthly sea ice extent and (b) monthly ice anomaly in the Southern Hemisphere for the period from November 1978 to October 2019.

1 Variability and trends of global sea ice cover and sea level

11

The monthly and interannual changes are more evident in the monthly anomaly plot shown in Figure 1.5(b). A large fluctuation in the monthly anomalies is apparent and indicative of large interannual changes for each month. The plot also shows the positive trend as described earlier. The result of regression analysis indicates a positive trend as described earlier but the trend has gone down to 0.4 ± 0.2% per decade. In light of what has been observed in the Arctic sea ice cover, such a trend is unexpected and has been the subject of much research activity [33–37]. Some look at longer-term time series and indicate that the ice cover was actually more extensive in the 1940s and 1950s according to ship observations [38, 39]. The changes have been primarily regional with the most positive trends occurring in the Ross Sea; it is strongly negative in the Bellingshausen/Amundsen Seas. Plots of yearly maximum and minimum extents and areas are presented in Figure 1.6. The day when the yearly maximum occurred is again determined using 5-day running averages to minimize short-term effects. The maximum was relatively uniform in the first 20 years, but has been on the rise in recent years with the maximum exceeding 20 × 106 km2 for the first time in 2014 during the satellite era. It is interesting to note that the following years show a sharp decline in the ice extent and ice area reaching a record minimum in 2017 and a slight recovery after that. This unexpected turn of events is intriguing and adds to the mysteries in the observed characteristics of the Antarctic sea ice cover [40]. The extent of the ice cover is affected by several factors, the contribution of which has not been accurately quantified. In particular, the rate of ice growth is enhanced by colder temperatures at the ice margins and by strong winds off coastal polynyas where ice is produced almost continuously [41, 42]. A freshening of the water due to melting ice shelves or iceberg calving also increases the rate of seawater freezing [37]. The ozone hole has been postulated as the cause for a deepening of the lows in the West Antarctic; this has been confirmed by numerical models [36]. This in turn causes stronger winds off the Ross Ice Shelf and higher ice production in the Ross Sea region as has been reported in a number of recent studies [35, 43]. The dramatic decline after the record high in 2014 has been associated to sustained changes in the ocean as caused by positive Southern Annular Mode and negative Interdecadal Pacific Oscillation [44]. Overall, the physics of the changes that includes climate variability and feedbacks are not well understood especially since climate models are not able to capture the observed changes [45]. Yearly minimum and maximum extents and ice areas are presented in Figure 1.6; the trends have become very moderate following the relative declines in the years after 2014. The trend for maximum ice extent is 0.42 ± 0.31% per decade while that of minimum extent is even negative at −0.68 ± 1.70% per decade. The corresponding trends for ice area are slightly different at 0.71 ± 0.34% and 1.41 ± 2.00% per decade, respectively but have consistent ice extent values within

12

Josefino C. Comiso

Figure 1.6: Plots of yearly maximum (blue), minimum (red), mean (green) and seasonal ice (gold) of (a) sea ice extent and (b) sea ice area from 1979 to 2019 in the Southern Hemisphere.

statistical errors. The larger discrepancy in the summer is likely associated with interannual variations in average ice concentration. The trends for yearly ice extent and area are 0.14 ± 0.50% and 0.68 ± 0.55% per decade, respectively while the corresponding trends for seasonal ice cover are 0.45 ± 0.45% and 0.40 ± 0.47% per decade. Although it was previously reported that the trend in Antarctic sea ice is significantly positive [46], it is apparent that the trend is no longer noteworthy and appears negligible. The trends of average ice concentration in the Southern Ocean for the different seasons are presented in Figure 1.7. It is apparent that in summer and autumn the trends are clearly divided into two areas: negative trends in Bellingshausen/ Amundsen Seas region and positive trends in other regions. During winter and spring, the patterns are quite different in that there is an alternating positive and negative trend around the continent. The trends are indicative of the mode 2 patterns of sea-level pressure that have been identified and studied previously [47, 48]. A pattern of a propagating wave called Antarctic circumpolar wave was identified and observed in the 1980s and 1990s [47], but was not so evident in the 2000s [28]. Instead, the

1 Variability and trends of global sea ice cover and sea level

13

Southern Hemisphere seasonal ice concentration trends 1979–2019 (a) Summer (DJF)

90oW

50o S

(b) Autumn (MAM)

90o E

60o S

(c) Winter (JJA)

(d) Spring (SON)

–2.4 –2.0 –1.6 –1.2 –0.8 –0.4 0.0 0.4 0.8 1.2 1.6 2.0 2.4 % Ice concentration/year

Figure 1.7: Color-coded trend maps of sea ice concentration for (a) summer, (b) fall, (c) winter and spring for the 1979 to 2019 period.

wave became more stationary with more persistent positive trends in the Ross Sea region. Again, this is consistent with results of modeling studies, as previously mentioned, indicating that the ozone hole had led to the deepening of the lows in the lower troposphere in West Antarctica leading to strong winds off the Ross Ice Shelf [35]. Strong winds caused the formation of larger coastal polynyas in the region and the production of more ice.

14

Josefino C. Comiso

1.3 Variability and trends in sea level One of the most serious impacts of global warming is the rise in global mean sea level (GMSL). Sea-level rise is a big concern because a large fraction of the inhabitants of our planet lives in coastal areas. It has been estimated that an increase in sea level by a few meters would cause the displacement of several hundred million people, immeasurably large economic losses and a mass destruction of the environment. The impacts on ecology and biodiversity are also expected to be profound [15]. In GMSL we refer to the height of the sea with respect to a benchmark (e.g., fixed reference such as a land feature) and averaged over long enough period to minimize if not eliminate the effects of big waves and tides. It also takes into account isostatic rebound as may be caused by the melt of ice sheets or glaciers over land and crustal movements. For a long time, the sea level has been monitored using tide gauges that were installed in cities and towns around the world. The gauges provide accurate and continuous readings of the sea level in areas where they are properly maintained. However, the measurements represent regional changes, and they do not necessarily represent changes in GMSL. This problem was minimized if not eliminated with the advent of satellite radar and laser altimeters that provide ocean topography measurements at high precision globally. The rise in sea level is caused mainly by two factors: (a) a warming of the ocean and (b) the introduction of liquid or solid water from land into the ocean. The melt of sea ice provides a negligible contribution to sea-level rise because sea ice is already part of the ocean.

1.3.1 Contributions from warming oceans Liquid water is known to expand as its temperature increases on account of the enhanced kinetic energy of the water molecules; they thus require more volume for the same number of molecules. The ocean is a vast storage of heat and energy, and it is estimated that 90% of the additional heat absorbed by the Earth in the last 50 years due to global warming has been stored in the ocean [49]. Getting estimates of the change in volume associated with global warming is not so straightforward because of many complex ocean processes associated with atmospheric forcing, ocean dynamics and different physical characteristics of the ocean in different regions. The expansion of the ocean depends on the quantity of heat absorbed and on water temperature, pressure and to a smaller extent, salinity. Greater expansion is expected to occur in warmer and saltier water under greater pressure. It has been estimated that for a simplified environment, seawater with a depth of 1 km expands by about 1 or 2 cm for every 0.1 °C increase in temperature [50]. If this estimate is correct, an increase of 4 °C for a doubling of the atmospheric CO2 could cause a significant sea-level rise of 40 to 80 cm. However, it turned out that there are many

1 Variability and trends of global sea ice cover and sea level

15

10 AVISO/CNES: trend = 3.31 mm year–1 CSIRO: trend = 3.36 mm year–1 University of Colorado: trend = 3.40 mm year–1 SL_cci/ESA: trend = 3.28 mm year–1 NASA/GSFC: trend = 3.44 mm year–1 NOAA: trend = 3.37 mm year–1 Averaged GMSL: trend = 3.35 mm year–1

Global mean sea level (cm)

8

6

4

2

0

–2 1992

1996

2000

2004

2008

2012

2016

250

200

Church & White (GRL, 2005)

Sea-level anomaly (mm)

University of Colorado (2014 Release 1) 150

100

3.2 mm year–1 (1993– 2013) 0.8 mm year–1 (1870–1924)

1.9 mm year–1 (1925–1992)

50

0 1880

1900

1920

1940

1960

1980

2000

Figure 1.8: (a) Altimetry data from six groups (CU, NOAA, GSFC, ESA, AVISO and CSIRO) for the period 1993 to 2017 with mean of the six shown and (b) yearly average GMSL reconstructed from tide gauges (1900–2013) by Church and White and the University of Colorado.

complications that need to be accounted for in making such an estimate for ocean expansion in the global oceans [51]. Data on sea-level rise as caused by thermal expansion has been relatively sparse until the 1980s when dedicated hydrographic measurements became available. Examples of repeated basin scale hydrographic measurements were those

16

Josefino C. Comiso

made for the North Atlantic [51, 52] and for the Southwest Pacific [53]. The results of these measurements as summarized [54] indicate that thermal expansion caused a sea-level rise of about 1 mm per year. A major cause of uncertainties to estimates of the rate of sea-level rise has been attributed to the occurrences of mesoscale eddies and large interannual variability in surface topography. Ocean thermal expansion has been cited as a major contributor to sea-level rise in the twentieth century; its effect is expected to continue in the twenty-first century [49, 55]. About half of the sea-level rise over the last few decades has been attributed to the warming of the ocean [56]. Combined with contributions from other sources, data for sea-level rise using modern techniques are presented in Figure 1.8(a), while the GMSL as reconstructed from tide gauges by different investigators are presented in Figure 1.8(b).The data shows some yearly variability, but it is apparent that the rate of increase has been considerably higher in the last decade. The series of altimeter data starting with GEOSAT in 1985 and followed by the more advanced systems (the most prominent of which are the TOPEX and Jason series) are presented in Figure 1.9. It is evident that there is a lot of noise in the data, especially with GEOSAT. There were also overlaps in measurements by the different sensors from 1992 to 2004; it is apparent that there are some inconsistencies. Such inconsistencies were resolved through the use of precise and well-documented reference frames (e.g., land). The long-term gauge data provided a rate of sealevel rise of 1.32 mm per year while the altimeter data provided 2.7 mm per year for the period 1985 to 2004. Again, this is an indication of an accelerated sea-level rise assuming that the two data sets provide consistent measurements of sea level. Data

100

TTM

Estimated global (81 °S to 81 °N) sea-level rise = 2.7±0.4 mm year–1

GFO

Mean sea-level variation (mm)

80 60

GEOSAT GM

40

TOPEX

GEOSAT ENVISAT Jason

ERM

20

ERS–2

0 –20 1985

ERS–1

1990

1995

2000

2005

Figure 1.9: Global mean sea level as measured by different altimetry missions from 1985 to 2004 [57]. A precise and well-documented reference frame that is monitored for several years is essential to ensure the consistency of the measurements. (With permission from Wiley-Blackwell, LTD.).

1 Variability and trends of global sea ice cover and sea level

17

from TOPEX and Jason, which are very similar systems, shows sea-level rise of 3.2 ± 0.4 mm per year for the period 1993 to 2009 [50].

1.3.2 Contributions from glaciers, ice sheets and others As indicated earlier, the melt of sea ice does not contribute to sea-level rise because it is already part of the ocean. However, through ice-albedo feedback and other effects from the retreat of sea ice, especially in spring and summer, a general warming of the region occurs; this makes other components of the cryosphere more vulnerable. Sea level is affected by the transfer of mass from land to the ocean. The contributions from snow and permafrost are through river runoff; there has not been any indication of significant increases in these contributions in recent years. The most important components of the cryosphere that could significantly, if not drastically, affect sea level are the glaciers and huge ice sheets in Greenland and Antarctica. The effect of climate change on glaciers is long term and it is hard to make attribution of interannual changes in the volume because the changes may be the result of climate forcing from the previous century. But during the twentieth century, the contribution of glacier melt to sea-level rise has been estimated to be considerable and actually exceeded the contribution from ice sheets [55]. This is despite the fact that the combined volume of all glaciers represents only about 1% of global ice volume. The big potential contributors to sealevel rise are the ice sheet of Greenland which has a sea-level equivalence of more than 7 m, and the Antarctic ice sheet which has a sea-level equivalence of more than 60 m. The impact of the loss of mass in glaciers is more than that associated with sealevel increase. Glaciers help regulate the seasonal water cycle and provide fresh water to neighboring regions during the dry season. The retreat of glaciers also causes a destabilization of mountain slopes and sometimes leads to the formation of meltwater lakes that are unstable and can cause flooding. Attribution of the loss of glaciers to anthropogenic causes has been studied and found to be detectable only in recent years (1991 to 2010) at 69 ± 24% of total loss when the anthropogenic signal becomes significant [58]. The ability to assess the location of significant mass losses in Antarctica and Greenland has been made possible by new technologies. In particular, many techniques have been adopted, including the use of radar altimetry, LIDAR altimetry and SAR interferometry; this results in estimates that are generally plausible but sometimes inconsistent [10]. Although data resolution is not as good as with other sensors, satellite data from the Gravity Recovery and Climate Experiment (GRACE) have allowed investigators to pin down the exact locations where the loss of mass is significant as illustrated in Figure 1.10(a) and (b). Thus, in Antarctica, mass loss is most prominent in West Antarctica in the general location of the Pine Island glaciers where massive icebergs have been calving here in recent years. West

18

Josefino C. Comiso

Antarctica has been an object of intense research because the ice sheet is sitting on a bedrock that is primarily below sea level, and the average temperature during the summer is just approximately −6 °C below the freezing temperature. In Greenland, there is significant mass loss in an area of the southern and western region that occupies more than half of the total area. Quantitative assessments of the mass loss per year of the glaciers and those from Greenland and Antarctica are presented in Figure 1.10(c). It is apparent that at least up to the present, the contribution from glaciers is more than double those from either Greenland or Antarctica.

2,000

0

–10 –8

–6

–4 –2

0

2

4 Ice gain

Ice loss (a)

3,000 2,000 1,000

2010

2008

2006

2004

2002

2000

1998

1996

1994

0 1992

Cumulative ice mass loss (Gt) (b)

4,000

2012

16 14 12 10 8 6 4 2 0 –2

5,000

SLE (mm)

Change 2003–2012 (cm of water / year–1)

Figure 1.10: Distribution of mass loss expressed in cm of water per year for the period 2003 to 2012 as determined by the Gravity Recovery and Climate Experiment (GRACE) time-variable gravity for (a) Antarctica and (b) Greenland. The graph at the bottom provides the total mass loss of ice from glaciers and the Greenland and Antarctic ice sheets in Gt and sea-level equivalent (mm) [52]. (With permission from IPCC and Cambridge University Press.).

1 Variability and trends of global sea ice cover and sea level

19

The concern has been the stability of the ice sheets; it is currently not well established. In West Antarctica, there is concern that the intense calving in the Amundsen Sea region will continue if not accelerate and cause a considerably larger contribution to sea-level rise. In Greenland, the area of the surface that experiences melt during the spring and summer has been increasing. The associated meltwater has been postulated to percolate to the bottom where it serves as a lubricant to the ice sheet and causes considerable changes in the dynamics of the system [59]. However, other scenarios are possible. For example, the meltwater could fill up the pores in the ice sheet and freeze; after that, it would be unlikely that liquid water would percolate all the way to the bottom in subsequent melt periods. Figure 1.11 shows yearly maps of Greenland that depicts areas that experienced melt during the years 2000 to 2016 as derived from MODIS surface temperature data [60]. It is apparent that the area of melt was most expansive in 2012 with almost the entire ice sheet having experienced some melt period. In 2002, the area of melt was also extensive, but not quite as bad as in 2012. After 2012, however, the melt areas appeared relatively normal.

Figure 1.11: Maps of maximum annual surface melt on the Greenland ice sheet from 2000 to 2016 as derived from MODIS monthly IST product (from [60]).

1.4 Effects on physicochemical parameters The impacts of the observed trends in sea ice cover and sea-level rise on the physicochemical characteristics in the regions are expected to be profound. The rapid retreat of sea ice cover in the summer will likely lead to a basically ice-free Arctic Ocean with a totally different environment and ecosystem. We have already observed that warming in the Arctic is amplified; this makes the other components of the cryosphere vulnerable and will likely cause a rise in sea level. A most important impact would be the change in physical characteristics including the vertical structure and circulation of the Arctic Ocean. An important layer in the Arctic Ocean is

20

Josefino C. Comiso

the halocline, which has kept deep-ocean warm water from upwelling to the surface and melting the ice. Without the halocline, the physical characteristics of the Arctic Ocean will change, and habitats of the ocean will have to adjust accordingly. The impacts of sea-level rise have direct and indirect components. Examples of the direct component are the displacement of hundreds of millions of inhabitants on the planet and the loss of important components of the climate system such as wetlands, mangroves and natural resources in the region. Among the indirect components, a large fraction of species and organisms will likely disappear, and the biodiversity of the region will change considerably. Also, sea-level rise will lead to the salinization of underground water near coastal areas; this is a critical source of freshwater and it is used in agricultural production. This leads to serious problems for hundreds of millions of people who depend on this source for their freshwater supply.

1.4.1 Large-scale changes in surface temperature Since 1900, the global surface temperature of the Earth as derived from meteorological stations and other data (i.e., GISS data) has been increasing at the rate of about 0.08 °C per decade [8]. For the period from 1981 to 2012, the same data shows a rate of increase of 0.17 °C per decade for the entire globe. This indicates an acceleration of warming (approximately twofold) since 1981 despite a reported hiatus since 1998 that has been observed and hypothesized to be due to natural climate variability. The same data also shows a rate of increase of about 0.60 °C per decade for the Arctic region (>64° N) while the satellite AVHRR data yields a rate of warming estimated to be 0.69 °C per decade for the same region [8]. The results show that the rate of warming in the Arctic is more than three times the rate of warming globally. This is often referred to as the amplification of global warming in the Arctic [61]. The AVHRR trend value likely provides the more accurate assessment because the data is more comprehensive and covers the entire Arctic region. Changes in water temperature have many implications for aquatic organisms at all trophic levels. Most organisms are sensitive to changes in water temperature, and most of them will not survive in extremely high temperatures. Correlation analysis of the plankton concentration with surface temperature data has indicated that depending on the region, bloom patterns are enhanced in regions like the Bering Sea and Okhotsk Sea as the sea ice recedes; however, the plankton concentration starts to decline as the temperature reaches a certain threshold value. Increases in water temperatures to more than 29 °C have led to bleaching coral reefs in many parts of the world. It is also well known that the warming of Eastern Pacific Ocean during El Niño–Southern Oscillation normally leads to the demise of millions of fish in the region. Tuna are usually found in waters that have temperatures of about 28 °C. Heat waves have occurred occasionally, and in Europe, a heat wave

21

1 Variability and trends of global sea ice cover and sea level

killed several thousands of people in August 2003 when 40 °C was reached and sustained for about a week. Also, changes in temperatures that are a few degrees above normal would usually shorten mating seasons and keep organisms from being able to reproduce. In the Arctic region, the color-coded trend maps of surface temperatures for the different seasons using satellite infrared data from 1981 to 2019 are presented in Figure 1.12. Overall, the trends are dominantly positive, but it is surprising that there are regions in Alaska and parts of Canada and Siberia where the trends are negative in the summer. Such trends are actually less negative than previously reported and consistent with a general warming in these regions in more recent

(a) Winter trend 1981–2019 180E

(c) Summer trend 1982–2019

(b) Spring trend 1982–2019

60N

K/dec

90E

(d) Autumn trend 1981–2019

>2.05 2.0 1.9 1.8 1.7 1.6 1.5 1.4 1.3 1.2 1.1 1.0 0.9 0.8 0.7 0.6 0.5 0.4 0.3 0.2 0.1 0.0 –0.1 –0.2 –0.3 –0.4 –0.5 –0.6 –0.7 –0.8 –0.9 –1.0 –1.1 –1.2 –1.3 –1.4 –1.5 –1.6 –1.7 –1.8 –1.9 –2.0 2.05 1.9 2.0 1.8 1.7 1.6 1.5 1.4 1.3 1.2 1.1 1.0 0.9 0.8 0.7 0.6 0.5 0.4 0.3 0.2 0.1 0.0 –0.1 –0.2 –0.3 –0.4 –0.5 –0.6 –0.7 –0.8 –0.9 –1.0 –1.1 –1.2 –1.3 –1.4 –1.5 –1.6 –1.7 –1.8 –1.9 –2.0 40. 30. 20.

90 E

10. 5. 3. 2. 1. Jul.

Sep.

Nov. .5 .3 .2 .1 .05 .03 .02 .01

Figure 1.14: Monthly climatology of chlorophyll pigment concentration in the Northern Hemisphere every other month from January to November. The climatology is from data starting in 1998 to 2019 using SeaWiFS and MODIS data.

surface temperature could also be a factor, especially in the Baffin Bay/Labrador Sea and North Atlantic regions. In the Southern Hemisphere, the plankton distribution of climatological data from 1997 to 2019 for every other month, as depicted in Figure 1.16, is more symmetric and appears to be enhanced near the polar fronts and follows a circular pattern north of the Antarctic Circumpolar Current [28, 57]. The bloom patterns are again most pronounced in spring and summer and especially in November and January. The most intense blooms occur off the shores of Argentina and are most likely associated with the abundance of nutrients and iron than the other regions. In winter, the regions immediately north of the sea ice cover have relatively low plankton concentrations in part because of minimal of solar exposure and perhaps the lack of nutrients and iron. But ocean acidification, especially where it persists, is also a possibility.

1 Variability and trends of global sea ice cover and sea level

80

25

(b) Amerasian Arctic

(a) Eurasian Arctic

Trend = 1.19 ± 2.69 (g C m–2 year–1 dec–1)

60 40 Trend = 12.74 ± 4.16 (g C m–2 year–1 dec–1) 20 90

(d) Bering Sea

(c) Sea of Okhotsk

Primary productivity (g C m–2 year–1 )

80

Trend = 2.45 ± 2.43 (g C m–2 year–1 dec–1)

70 60 Trend = 2.08 ± 2.83 (g C m–2 year–1 dec–1) 50 100 (e) Barents Sea

(f) Greenland Sea Trend = 6.00 ± 2.07 (g C m–2 year–1 dec–1)

90 80 70 –2 –1 –1 60 Trend = 12.85 ± 3.24 (g C m year dec ) 50 70 (g) Hudson Bay

60

(h) Baffin Bay/Labrador Sea

Trend = 5.46 ± 2.51 (g C m–2 year–1 dec–1)

50 Trend = 5.67 ± 2.12 (g C m–2 year–1 dec–1)

40 30 70 (i) North Atlantic

(j) Average of nine regions

65 Trend = 5.89 ± 1.89 (g C m–2 year–1 dec–1) 60 55 50

Trend = 6.04 ± 1.59 (g C m–2 year–1 dec–1)

45 2005

2010

2015

2005

2010

2015

Figure 1.15: Primary productivity (2003–2018, March–September only) in nine different regions of the Northern Hemisphere (for a definition of the regions see Comiso, 2015), as well as the average of these nine regions, derived using chlorophyll-a concentrations from MODIS-Aqua data, the NOAA 1/4° daily Optimum Interpolation Sea Surface Temperature dataset (or daily OISST) that uses satellite sea surface temperatures from AVHRR and additional parameters. Values are calculated based on the techniques described by Behrensfield and Falkowski (1997) and represent net primary productivity (NPP).

The influence of sea ice on the primary production in the Southern Ocean has been discussed in detail elsewhere [18, 64]. Plots of primary productivity in the Southern Ocean are presented in Figure 1.17. The data shows a negative trend for all sectors combined and also for each individual sector. This appears to be in

26

Josefino C. Comiso

Jan.

Mar. 45 E

mg/m3 >30. 20. 10. 30 S May

135 E 5. Jul.

3. 2. 1. .5 .3 .2 .1

Sep.

Nov. .05 .03 .02 .01

Figure 1.16: Color-coded maps of monthly climatology of chlorophyll pigment concentration in the Southern for every other month from January to November. The climatology data is from 1998 to 2019 using SeaWiFS and MODIS data.

contradiction to the positive trends in productivity in the region as reported previously for the period 1997 to 2008 [64] but this is mainly due to persistently low values since 2011. The reversal of the trend after the addition of a few years of data reveals a weakness of doing trend analysis on data with relatively short record length. The reason for the consistently low values from 2011 to 2019 is not known and may be caused by the negative trends in the ocean’s surface temperature or the

1 Variability and trends of global sea ice cover and sea level

30

(a) Weddell Sea

(b) Weddell Sea

Trend = –4.35 ± 1.32g C m-2 year-1 dec-1

Trend = –5.70 ± 1.40g C m-2 year-1 dec-1

27

25

20

Primary Production g C m-2 year-1

15 45 (c) W. Pacific Ocean

(d) Ross Sea Trend = –7.92 ± 2.17g C m-2 year-1 dec-1

40 35 30 25 20 Trend = –7.38 ± 2.35g C m-2 year-1 dec-1

60 (e) Bellingshausen, Amundsen Seas (f) Average of areas Trend = –7.24 ± 1.99g C m-2 year-1 dec-1

50 40 30 20

Trend = –10.86 ± 2.91g C m-2 year-1 dec-1

2000

2005

2010

2015

2000

2005

2010

2015

2020

Figure 1.17: Plots of yearly primary productivity from 1998 to 2019 in different regions of the Southern Hemisphere as derived using Chlorophyll a concentrations from SeaWiFS and MODIS data, AVHRR surface temperature and other parameters.

positive trends in the sea ice cover especially in the summer, but more research is needed to establish this. The enhanced values from 2001 to 2010 also need some explanation, but again further studies are required. The ocean color data have been invaluable in many other marine science studies. For example, the data can be used to identify regions where harmful algal blooms (sometimes called red tides) occur. In this case, the signal is caused by a bloom of toxic red dinoflagellates. These are normally depicted as high values in plankton concentration that are occasionally observed in coastal waters. The data

28

Josefino C. Comiso

could therefore provide advanced and critical information about onset and magnitude of these blooms. On the other hand, since plankton is at the bottom of the food web the data has been used by the fishing industry to locate schools of fishes. Ocean color data are also invaluable in the assessment of the effects of ocean acidification as discussed in the next section. There are already some observations of areas that are highly depleted in plankton concentration and can be candidate areas for studying the effects of ocean acidification.

1.4.3 Changes in other physicochemical parameters The aforementioned changes in the global sea ice cover and the rise in sea level will undoubtedly cause pronounced shifts and reorganizations in global and regional ecosystems and biogeochemical cycles [65, 66]. Finding direct linkages of these changes to actual changes in physicochemical parameters (other than surface temperature and plankton concentration) is not trivial because of the lack of reliable numerical models that can be used for sensitivity studies and in situ data to validate the results of such studies.

1.4.3.1 pH and ocean acidification A large fraction of atmospheric carbon dioxide is taken up by the ocean. After being absorbed, the carbon dioxide dissolves and forms carbonic acid, which breaks down quickly into bicarbonate (HCO3− ) and a hydrogen ion (H+). The acidity of liquid water is usually quantified by its pH, which is a measure of its acid/base activity and defined as the negative common logarithm of the activity/concentration of hydrogen ions: pH = − log½H +  The pH of pure water at room temperature is 7 and is regarded as the pH for natural waters that are not acidic. Freshwater with pH less than 7 is considered acidic while those greater than 7 represent base saturation or alkalinity. For ocean seawater, the pH level is slightly higher and is currently at 8.2. The introduction of more hydrogen ions would reduce the pH and make it more acidic. Since carbon dioxide is taken up more rapidly in colder water, the regions that are most vulnerable are the polar regions. Thus, in the Arctic basin where sea ice is retreating, the presence of more cold open water areas that are being exposed due to ice cover decline means enhanced acidification in the region. The images in Figures 1.14 and 1.16 show low plankton concentrations in parts of the ocean. Such low values may be in part a manifestation of acidification in the region. The Arctic Ocean is actually regarded as the first ocean where a rapid spread in acidification has been observed

1 Variability and trends of global sea ice cover and sea level

29

[67]. As observed in this ocean, the rate is twice as fast as in the Pacific and Atlantic Oceans. Such acidification has serious implications for marine life including clams, mussels and sea snails that have difficulty in maintaining their shells in acidic water.

1.4.3.2 Dissolved oxygen The availability of dissolved oxygen (DO) in the ocean is required by all forms of aquatic life. The concentration of DO is influenced primarily by biological activities and through photosynthesis in aquatic plants. Thus, DO is relatively high during the daytime and is reduced during nighttime. Oxygen also tends to be more soluble in colder than warmer water and may be more readily available in the polar seas. The typical DO of unpolluted freshwater is about 10 mg L–1. It has been observed that fish kills occur when the DO levels goes down to less than 2 mg L–1. Again, a key source of DO is ocean plankton. Thus, a depletion of plankton as may be caused by the loss of sea ice, ocean acidification or other factors would cause a reduction of DO.

1.4.3.3 Electrical conductivity and turbidity The electrical conductivity of water is mainly influenced by the amount of salt in the water. The more saline the water, the higher the electrical conductivity. Globally, the spatial variations of the ocean’s salinity are primarily controlled by precipitation, evaporation and sea ice cover. Precipitation causes the introduction of freshwater, which causes a reduction of salinity; evaporation does the opposite. In the polar oceans, the impact of sea ice on salinity depends on the season. During the growth period, the ocean in the underside of sea ice is usually cold and saline because of sea ice production. Sea ice can absorb only less than 30% of seawater salt during formation; this causes the water on the underside of the ice to be relatively saline. This is especially the case where the rate of ice production is highest in autumn and winter as in coastal polynyas and ice edges. During the spring and summer, the opposite happens with the introduction of melt water that is less saline than the regular salt water. The positive trend in ice extent in the Antarctic means that the rate of ice production is getting higher and therefore more high-density saline water is being formed. On the other hand, the negative trend in the Arctic during the summer means a freshening of the surface water with the introduction of more meltwater. Turbidity or total suspended solids affect the transparency and light scattering of water and is usually composed of fine clay, silt particles, plankton, organic or inorganic compounds and microorganisms. The size of suspended particles ranges from 10 nm to 0.1 mm but are normally defined as those that do not pass through a 45 μm filter. In the Arctic, turbidity is usually affected mainly by river runoff, windinduced upwelling and changes in plankton concentration; in the Antarctic, winds

30

Josefino C. Comiso

and plankton concentration are the key factors. Changes in the acidity of the ocean could affect turbidity, since it changes the solubility of suspended matter.

1.4.3.4 Nutrients There is large spatial variability in the distribution of nutrients in the oceans and in the polar regions; they may not be as readily available in some as in other regions. The most vulnerable areas are those that are farthest away from land. The loss of sea ice in the Arctic basin exposes more open water with surface layers consisting of meltwater that can be the platform for phytoplankton blooms. However, this would not materialize if nutrients were not available to support the blooms. Also, as the ice retreats to the deep ocean region, it is likely that these regions are depleted of nutrients. The reason for this is that when sea ice forms, the water on the underside becomes dense due to salinization and tends to go down to the deeper portion of the water. During the process, nutrients are entrained with the dense water into the deeper parts of the ocean where it may not be possible for them to resurface.

1.5 Discussion and conclusions The effect of global warming on sea ice cover and the GMSL is studied and the possible impacts to the physicochemical characteristics in the region is evaluated. Global warming is not uniformly distributed around the globe. The warming has been observed to be amplified by about three times and observed to be 0.69 °C per decade in the Arctic region (>64° N) compared to 0.17 °C per decade globally; however, no amplification is observed in the Antarctic region. The impact of ice-albedo feedback is apparent in the Arctic where dramatic reductions in summer ice is observed and the extent of the perennial sea ice cover has been reduced to less than half its value over a span of three decades. The impact of the decline of sea ice in the Arctic is a general warming in the region that has caused the volume of glaciers to decline and the Greenland ice sheet to lose mass, which together with the warming ocean has caused an increase in the rate of sea-level rise to about 3.4 mm per year from 1.32 as measured by the long-term gauge data. The trend in extent of the Antarctic sea ice cover continues to be positive, but has been reduced from 2% per decade for data up to 2014 to 0.4% per decade for data up to 2019 because of dramatic declines in extent in recent years. Positive trends are most significant in the Bellingshausen/Amundsen Seas adjacent to the West Antarctic region where mass loss is significant. The impacts of the observed decline of sea ice cover in the Arctic can be profound and would affect the physical oceanography, the primary productivity and the circulation patterns of the ocean and the atmosphere. The rate

1 Variability and trends of global sea ice cover and sea level

31

of sea-level rise has been modest [68] but accelerated to 3.4 mm per year which is still relatively low. However, since the ice sheets appear to be relatively unstable, large changes in the GMSL could occur in the future. The impacts of changes in sea ice and sea level on the physicochemical characteristics of the affected regions including the ecology, the environment and aquatic organisms at all trophic levels are also very negative and could get worse by the end of the twenty-first century.

References [1] [2] [3]

[4] [5] [6] [7] [8] [9] [10]

[11] [12] [13] [14] [15] [16] [17] [18]

Comiso JC, Parkinson CL, Gersten R, Stock L. Accelerated decline in the Arctic sea ice cover. Geophys Res Lett 2008;35:L01703. Doi: 10.1029/2007GL031972. Church JA, White NJ. A twentieth century acceleration in global sea-level rise. Geophys Res Lett 2006;33:L01602. Doi: 10.1029/2005GL024826. Rignot E, Velicogna I, van den Broeke MR, Monaghan A, Lenaerts JTM. Acceleration of the contribution of the Greenland and Antarctic Ice Sheets to sea level rise. Geophys Res Lett 2011;38:L05503. Doi: 10.1029/2011GL046583. Comiso JC. A rapidly declining Arctic perennial ice cover Geophys Res Letts 2002;29 (20):1956. Doi: 10.1029/2002GL015650. Kinnard C, Zdanowicz CM, Fisher DA, Isaksson E, De Vernal A, Thompson LG Reconstructed changes in Arctic sea ice over the past 1,450 years. Nature 2011;479:509–512. Parkinson CL, Comiso JC. On the 2012 record low Arctic sea ice cover: Combined impact of preconditioning and an August storm. Geophy Res Lett 2013;40:1–6. Doi: 10.1002/grl.50349. Stroeve J, Holland MM, Meier W, Scambos T, Serreze M. Arctic sea ice decline: Faster than forecast. Geophys Res Lett 2007;34:L09501. Doi: 10.1029.2007GL029703. Comiso JC, Hall DK Climate Trends in the Arctic, WIREs (Wiley Interdisciplinary Reviews) Climate Change: Advanced Review 2014; Doi: 10.1002/wcc.277. Lindsay R, Zhang J The thinning of Arctic sea ice, 1988–2003: Have we passed a tipping point? J. Climate 2005;18:4879–4894. Shepherd A, Ivins ER, Geruo A, Barletta VR, Bentley MJ, Bettadpur S, et al. A reconciled estimate of ice-sheet mass balance. Science 2012;338:1183–1189. Doi: 10.1126/ science.1228102. Hansen J, Ruedy R, Sato M, Lo K. Global surface temperature change. Rev Geophys 2010;48. Doi: 10.1029/2010RG000345. Zwally HJ, Comiso JC, Parkinson CL, Cavalieri DJ, Gloersen P. Variability of the Antarctic sea ice cover. J Geophys Res 2002;107(C5):1029–1047. Comiso JC, Nishio F. Trends in the sea ice cover using enhanced and compatible AMSR-E, SSM/I and SMMR data. J Geophys Res 2008;11:C02S07. Doi: 10.1029/2007JC004257. Doney SC, Ruckelshaus M, Duffy JE, Barry JP, Chan F, et al. Climate change impacts on marine ecosystems. Annu Rev Mar Sci 2012;4:11–37. Schofield O, Ducklow HW, Martison DG, Meredith MP, Moline MA, Fraser WR How do polar marine ecosystems respond to rapid climate change?. Science 2010;328:1520–1523. Nicholls RJ, Cazenave A Sea-level rise and its impact on coastal zones. Science 2010;328: 1517–1520. Gordon AL, Comiso JC Polynyas in the Southern Ocean. Sci Am 1988;256:90–97. Smith WO Jr, Comiso JC. Southern ocean primary productivity: Variability and a view to the future, in Smithsonian at the Poles: Contributions to the International Polar Year Science.

32

[19] [20] [21]

[22] [23] [24] [25]

[26]

[27] [28] [29] [30] [31] [32] [33]

[34]

[35] [36]

[37] [38]

Josefino C. Comiso

Krupnik I, Lang MA, Miller SE [eds.]. Washington, DC; Smithsonian Institution Scholarly Press: 2009. 309–318. Zwally HJ, Comiso JC, Parkinson CL, Campbell WJ, Carsey FD, Gloersen P Antarctic Sea Ice 1973–1976 from Satellite Passive Microwave Observations. NASA Spec Publ 1983;459:206. Comiso JC, Cavalieri DJ, Markus T. Sea ice concentration, ice temperature, and snow depth, using AMSR-E data. IEEE TGRS 2003;41(2):243–252. Gloersen P, Campbell WJ, Cavalieri D, Comiso JC, Parkinson CL, Zwally HJ. Arctic and Antarctic Sea Ice, 1978–1987: Satellite Passive Microwave Observations and Analysis. NASA Spec Publ 511:1992. Bjorgo E, Johannessen OM, Miles MW. Analysis of merged SMMR-SSMI time series of Arctic and Antarctic sea ice parameters 1978–1995. Geophys Res Lett 1997;24(4):413–416. Parkinson CL, Cavalieri DJ, Gloersen P, Zwally HJ, Comiso JC. Arctic sea ice extents, areas, and trends, 1978–1996. J Geophys Res 1999;104(C9):20837–20856. Cavalieri DJ, Parkinson CL Arctic sea ice variability and trends, 1979–2010. The Cryosphere 2012;6:957–979. Comiso JC, Parkinson CL. Arctic sea ice parameters from AMSR-E using two techniques, and comparisons with sea ice from SSM/I. J Geophys Res 2008;113:C02S05. Doi: 10.1029/ 2007JC004255. Parkinson CL, Comiso JC. Antarctic sea ice from AMSR-E from two algorithms and comparisons with sea ice from SSM/I. J Geophys Res 2008;113:C02S06. Doi: 10.1029/ 2007JC004253. Parkinson CL, Cavalieri DJ. Arctic sea ice variability and trends, 1970–2006. J Geophys Res 2008;113:C07003. Doi: 10.1029/2007JC004558. Comiso JC. Polar Oceans from Space. New York; Springer Publishing: 2010. Doi: 10.1007/ 978-0-387-68300-3. Kwok R, Rothrock DA. Variability of Fram Strait ice flux and North Atlantic Oscillation. J Geophy Res 1999;104(C3):5177–5189. Vinje T. Fram strait ice fluxes and atmospheric circulation: 1950–2000. J Climate 2001: 3508–3517. Kwok R, Untersteiner N The thinning of Arctic sea ice. Physics Today 2011;64:36–41. Zwally HJ, Comiso JC, Parkinson CL, Cavalieri DJ, Gloersen P. Variability of the Antarctic sea ice cover. J Geophys Res 2002;107(C5):1029–1047. Kwok R, Comiso JC, Martin S, Drucker RS. Ross Sea Polynyas: Response of ice concentration retrievals to large areas of thin ice. J Geophys Res 2007;112:C12012. Doi: 10.1029/ 2006JC003967. Martin S, Drucker RS, Kwok R The areas and ice production of the western and central Ross Sea polynyas, 1992–2002, and their relation to the B-15 and C-19 iceberg events of 2000 and 2002. J Marine Systems 2007;68:201–214. Comiso JC, Kwok R, Martin S, Gordon A. Variability and trends in sea ice and ice production in the Ross Sea. J Geophys Res 2011;116:C04021. Doi: 1029/2010JC006391. Turner J, Comiso JC, Marshall GJ, Lachlan-Cope TA, Bracegirdle T, Maksym T, Meredith M, Wang Z. Non-annular atmospheric circulation change induced by stratospheric ozone depletion and its role in the recent increase of Antarctic sea ice extent. Geophy Res Lett 2009;36:L08502. Doi: 10.1029/2009GL037524. Zhang J. Increasing Antarctic sea ice under warming atmospheric and oceanic conditions. J Climate 2007:2515–2529. de la Mare WK. Abrupt mid-twentieth century decline in Antarctic sea-ice extent from whaling records. Nature 1997;389:57–60. Doi: 10.1038/37956.

1 Variability and trends of global sea ice cover and sea level

33

[39] Ackley SF, Wadhams P, Comiso JC, Worby A. Decadal decrease of Antarctic sea ice extent from whaling records revisited on the basis of historical and modern sea ice records. Polar Res 2003;22(1):10–25. [40] Turner J, Comiso JC, et.al. Comment: Solve Antarctica’s sea-ice puzzle. Nature 2017;547 (7663), Doi: 10.1038/54727a. [41] Zwally HJ, Comiso JC, Gordon AL. Antarctic offshore leads and polynyas and oceanographic effects. In: Jacobs. S [ed.] Oceanology of the Antarctic Continental Shelf. Antarctic Research: 1985;Vol. 43:203–226. [42] Comiso JC, Gordon AL Interannual variabilities of summer ice minimum, coastal polynyas, and bottom water formation in the Weddell Sea, in Antarctic sea ice physical properties and processes. AGU Antarct Res Ser 1998;74:293–315. [43] Holland PR, Kwok R Wind-driven trends in Antarctic sea ice drift. Nat Geosci 2012;5:872–5. [44] Meehl GA, Arbetlster JM, Chung TY, Holland MM, DuVivier A, Thompson LA, Yang D, Bitz CM. Sustained ocean changes contributed to sudden Antarctic sea ice retreat in late 2016. Nat Commun 2019;10:1–9. Doi: doi.org/10.1038/s414-018-07865-9. [45] Maksym T Antarctic sea ice change: Contrasts, commonalities and causes. Annu Rev Mar Sci 2019;11:187–213. [46] Comiso JC, Gersten R, Stock L, Turner J, Perez G, Cho K. Positive trends in the Antarctic sea ice cover and associated changes in surface temperature. J Climate 30:2251–2267. Doi: 10.1175/JCLI-D-0408.I. [47] White WB, Peterson RG. An Antarctic circumpolar wave in surface pressure, wind, temperature and sea ice extent. Nature 1996: 380-699-702. Doi: 10.1038/380699a0. [48] Yuan X, Li C. Climate modes in southern high latitudes and their impacts on Antarctic sea ice. J Geophys Res 2008;113:C06S91. Doi: 10.1029/2006JC004067. [49] Bindhoff N, Willebrand J, Casenabe A, Gregory JM. Observations: ocean climate change and sea level. In: Solomon S, Qin D, Manning M, Marquis M, Averyt K, Tignor MMB, et al. [ed.] Climate Change 2007: The Physical Sciences Basis, Contributions of Working Group 1 to the Fourth Assessment Report of the Intergovernmental Panel on Climate Change. Cambridge, UK; Cambridge University Press: 385–432. [50] Church JA, Roemmich D, Domingues CM, Willis JK, White NJ, Gilson JE, Stammer D, et al. Ocean Temperature and Salinity Contributions to Global and Regional Sea-Level Change. In: Church JA, Woodworth PL, AARup T, Wilson WS [ed.] Understanding sea-level rise and variability. Ox- ford, UK; Blackwell Publishing: 2010. 143–176. [51] Roemmich D, Wunch C Apparent changes in the climatic state of the deep North Atlantic Ocean. Nature 1984;307:447–450. [52] Levitus S Interpentadal variability of steric sea level and geopotential thickness of the North Atlantic Ocean, 1970–74 versus 1955–59. J Geophys Res 1992;95:5233–5238. [53] Bindhoff N, Church J Warming of the water column in the southwest Pacific Ocean. Nature 1992;357:59–62. [54] Church JA, Gregory JM, Huybrechts P, Kuhn M, Lambeck K, et al. Changes in Sea Level. In: Houghton JT, Ding Y, Griggs DJ, Noguer M, van der Linden PJ, Dai X [ed.] Climate Change 2001: The Physical Sciences Basis, Contributions of Working Group 1 to the Third Assessment Report of the Intergovernmental Panel on Climate Change. Cambridge, UK; Cambridge University Press: 2001. 639–94. [55] Meehl GA, Stocker TF, Collins WD, Friedlingstein P, AT G, Gregory JM, et al. Global climate projections. In: Solomon S, Qin D, Manning M, Marquis M, Averyt K, Tignor MMB, et al. [ed.] Climate Change 2007: The physical science basis. Contribution of Working Group 1 to the Fourth Assessment Report of the Intergovernmental Panel on Climate Change. Cambridge, UK; Cambridge University Press: 2007. 747–845.

34

Josefino C. Comiso

[56] IPCC2013. Climate change 2013: The Physical Basis. New York; Cambridge University Press: 2013. 1535. [57] Moore JK Abbott MR and Richman JG. Location and dynamics of the Antarctic Polar Front from satellite sea surface temperature data. J Geophy Res 1999;104:3059–3073. [58] Marzelon B, Cogley JG, Richter K, Parkes D. Attribution of global glacier mass loss to anthropogenic and natural causes. Science 2014. Doi: 10.1120. [59] Zwally HJ, Abdalati W, Herring T, Larson K. Saba J and Steffen K. Surface melt-induced acceleration of Greenland ice-sheet flow. Science 2002;297(5579):218–222. [60] Hall DK, Cullather RL, DiGirolamo NE, Comiso JC, Medley BC, Nowicki SM. A multilayer surface temperature, surface albedo, and water vapor product of Greenland from MODIS. Remote Sens 2018;10:555. Doi: doi.10.3390/rs10040555. [61] Holland MM, Bitz CM. Polar amplification of climate change in coupled models. Climate Dynamics 2003;21:221–232. Doi: 10.1007/s00382-003-0332-6. [62] Mildrexler D, Oyle D, Comiso JC. Surface temperature inter-relationships, in Taking the temperature of the Earth. ed by [Hulley G, Ghent D] Springer. Doi: org/10.1016/B978-0-12-. [63] Wassman P, Ratkova T, Andreassen I, Vernet M, Pedersen G, Rey F Spring bloom development in the marginal ice zone and central Barents Sea. Marine Ecology 1999;20:321–346. [64] Smith W Jr, Comiso JC. The influence of sea ice on primary production in the Southern Ocean: A satellite perspective. J Geophys Res 2008;113:C05S93. Doi: 10.1029/2007JC004251. [65] Grebmeier JM, Overland JE, Moore SE, Farley EV, Camack EC, et al. A major shift in the Northern Bering Sea. Science 2006;311:1461–1464. [66] Overland JE, Stabeno PJ. Is the climate of the Bering Sea warming and affecting the ecosystem? EOS, Transaction. AGU 2004;85(33):309–316. [67] Qi D, Chen L, Gao Z, Zhong W, Feely RA, Anderson LG, Sun H, Chen J, Chen M, Zhan L, Zhang Y, Zhang Y, Cai WJ. Increase in acidifying water in the western. Arctic Ocean, Nature Climate Change 2017:195-xxx. Doi: 10.1038/nclimate3228. [68] Church JA, White NJ Sea level rise from the late 19th to early twentieth century. Surv Geophys 2011;32:585–602.

Begoña Espiña, Marta Prado, Stephanie Vial, Verónica C. Martins, Soraia P.S. Fernandes, Marilia B. dos Santos, Laura M. Salonen and Raquel B. Queirós

2 New techniques in environment monitoring 2.1 Introduction Microalgae in marine, and cyanobacteria in both marine and freshwaters are able to produce harmful effects, including a broad range of phenomena referred to as “harmful algal blooms” (HABs). In recent decades, scientists have observed an increase in the frequency, severity and geographic distribution of HABs worldwide. Recent research suggests that the impacts of climate change may promote the growth and dominance of HABs through a variety of mechanisms, including – warmer water temperatures, – changes in salinity, – increases in atmospheric carbon dioxide concentrations, – changes in rainfall patterns, – intensifying of coastal upwelling, – sea-level rise. An adequate management of HABs requires monitoring of microalgae. The global scale, the severity of some intoxications derived by them and their relation with environmental phenomena such as eutrophication highlight the need of adequate monitoring tools for microalgae and cyanobacteria in order to control and prevent HABs. Such monitoring tools should allow the fast, accurate and specific detection of HAB species in monitoring programs, which offers important advantages such as understanding the presence of HAB species and their distribution and dispersion mechanisms, which contributes as well to the prevention or mitigation of the harmful effects of HAB on human health, marine ecosystem and its related economic activities [1]. There are several physicochemical parameters that have demonstrated impact and relevance in the occurrence of HABs and that can be correlated with microalgae abundance in order to monitor and predict them. These parameters are usually measured in water quality control programs and include the use of

Begoña Espiña, Marta Prado, Marilia B. dos Santos, Laura M. Salonen, Raquel B. Queirós, International Iberian Nanotechnology Laboratory (INL), Braga, Portugal Soraia P.S. Fernandes, International Iberian Nanotechnology Laboratory (INL), Braga, Portugal; Department of Chemistry, QOPNA, University of Aveiro, Campus Universitário de Santiago, Aveiro, Portugal Stephanie Vial, 3B’s Research Group, I3Bs – Universidade do Minho, Barco GMR, Portugal Verónica C. Martins, INESC Microsistemas e Nanotecnologias Rua Alves Redol, Lisbon, Portugal https://doi.org/10.1515/9783110625738-002

36

Begoña Espiña et al.

methods (sensors, in most part of cases) to measure conductivity, pH, dissolved oxygen, salinity, temperature, depth, turbidity, ammonium and nitrate. In this chapter, we will not focus on those parameters but in environment monitoring tools whose main aim is to give information about the biological composition of waters and toxin profiles. The description of recent advances in new technologies for analytical systems and biosensing (in lab bench analysis and in situ), including nanotechnology-based and molecular biology-based methods, and their application to HAB monitoring and control will be our main goal. Environmental monitoring covers a large range of fields related to air, soil and water analysis, and it can take many forms and levels of technical sophistication. This chapter covers the technological progress in the biosensing area that has led to improved detection capabilities at the sensitivity, accuracy and multiplexability of biosensors applied to waterborne toxins, in particular, toxins produced by phytoplankton (e.g., dinoflagellates and algae). Particular attention is devoted to the nanomaterials investigated as enhancing strategies applied to biosensors for waterborne toxins. Additionally, novel molecular strategies based on genetic tools for DNA/RNA amplification and detection techniques are revised.

2.2 In situ harmful algal bloom monitoring The increase in the frequency of occurrence of HABs makes their prediction and early detection an even more important concern in environmental monitoring. However, early detection is difficult to attain and the time of response increases when monitoring relies only in in situ sample collection and further laboratory bench analysis. Taking this into account, new automated systems for HAB monitoring are necessary. It is important to clarify that currently monitoring of HABs is based on the detection and/or quantification of significant increases in the phytoplanktonic biomass that can include, or not, the occurrence of a toxinogenic HAB. In this sense, harmfulness of the exponential growth of phytoplankton and/or cyanobacteria is not always based on the production of toxins as many species do not or, at least were not reported to produce biotoxins, but their massive growth can imply other noxious effects on the ecosystem such as oxygen deployment, decrease in the light penetration depth and nutrient depletion. Nevertheless, the exponential growth of toxinogenic species gives a new relevance to the HAB that should be taken into account. Even so, very few automated monitoring systems can discriminate toxic HABs from the ones that are not.

2 New techniques in environment monitoring

37

Taking this into account, we will devote this section to new trends, advances and currently used automated monitoring systems for HABs but trying to focus, when possible, in technologies that could discriminate between toxinogenic and nontoxinogenic ones.

2.2.1 Optical remote sensing Remotely sensed data collection and interpretation is the front line in monitoring and forecasting algal blooms. The use of Earth Observation (EO) remote sensing data for the study of HABs has been the object of many advances and studies such as in the Algal Bloom Detection, Monitoring and Prediction project, a Concerted Action funded by the European Commission and carried out from April 1997 to March 1999 [2]. Several institutions in oceanic countries develop, implement and operate services using satellite EO data for monitoring HABs and other water quality parameters. This way, the Nansen Environmental and Remote Sensing Center in Bergen (Norway) as well as the Monitoring and Event Response for Harmful Algal Blooms Research Program from the National Oceanic and Atmospheric Administration (USA) use satellite imagery data collected from the European Space Agency’s Medium Resolution Imaging Spectrometer (MERIS) and other EO sensors to forecast the occurrence of HAB from numerical ocean models based on optical signals derived from diverse parameters such as chlorophyll a, yellow substance and sea surface temperature (SST). MERIS instrument has been particularly effective to estimate biomass mainly due to its 300 m resolution, 2-day repeat orbit and sufficient spectral bands. Spatiotemporal results from MERIS are still being analyzed, reporting bloom abundances or pointing out to the high relevance of nutrients such as phosphate [3]. Satellite remote sensing provides a potential technology for identifying cyanoHABs in multiple waterbodies and across geopolitical boundaries. MERIS was used to quantify cyanoHAB surface area extent, transferable to different spatial areas, in Florida, Ohio and California for the test period of 2008 to 2012 [4]. In mid-2015, Sentinel-3, part of a series of Sentinel satellites from Copernicus program was launched including the Ocean Land Color Instrument (OLCI). The OLCI operates across 21 wavelength bands from ultraviolet (UV) to near-infrared and uses optimized cameras directed to reduce the effects of sun glint. Recently, a mobile device application that uses data from OLCI in near real time called Cyanobacteria Assessment Network was demonstrated to perform initial water quality assessments and quick alerts on potential cyanobacterial bloom threats [5]. More recently, lower altitude hyperspectral imaging (HSI) sensors have been developed providing improved and detailed spectral signatures of the water surface. This diverse spectral information supports the ability of the system to sense a

38

Begoña Espiña et al.

wide range of surfaces, habitats and taxonomic groups, allowing to discriminate between different types of phytoplankton. The obtained spectral information may act as an early warning water HAB detection system; however, due to the high costs associated with flying HSI-equipped aircraft, this system is not viable as a first line of monitoring [6]. However, despite the usefulness of EO, there is no evident relationship between the concentration or the biological composition of an algal bloom, on the one hand, and the observed reflectance spectrum from the water in which the bloom occurs on the other hand. The observed reflectance spectrum depends, for one, on the pigments present in the bloom. There are three main groups of photosynthetic pigments: chlorophylls, carotenoids and phycobiliproteins (phycobilins). Common to all photosynthetic organisms that produce oxygen, chlorophyll a is often used to calculate overall phytoplankton abundance, although it is highly variable depending on the physiological conditions, degree of light adaptation and state of degradation [7–9]. There are other oceanic physicochemical parameters that are used to monitor and forecast microalgae growth by optical remote sensing that are even more imprecise in terms of determining the biological composition or harmfulness of the episode, but can help in early warning and tracking of HABs, above all, when combined with other technologies. This is the case of the SST tracking by infrared emission: in oligotrophic and mixed waters, sun energy can penetrate as much as 30 m in depth, implying a high-temperature mitigation rate. However, when there are high concentrations of phytoplankton in surface waters relatively scarce solar energy can penetrate the water column, meaning that most part of it is absorbed to heat in the upper few meters. Additionally, SST allows the inference of the availability of some necessary nutrients for phytoplankton. Some of them, such as nitrite, are only supplied in normal conditions from the waters below the thermocline, mainly in upwelling areas. So, SST would act as a tracer of nitrite concentration [2]. It is also very important to keep in mind that availability or scarcity of certain nutrients may favor/disfavor the growth of certain phytoplankton species (toxinogenic or not). Taking this into account, a deep knowledge about the favorable ecophysiological conditions for the harmful species growing is essential. As reviewed, optical remote sensing systems constitute highly valuable tools to help and support to forecast and monitor the occurrence of HABs; nevertheless, currently these tools are not still absolute solutions, as they present some limitations that can only be overcome by in situ sampling and analysis.

2 New techniques in environment monitoring

39

2.2.2 Automated monitoring Early warning systems are needed in order to prevent and minimize the consequences of occurrence of HABs. Regular monitoring usually implies collection of samples at specific space and time points and posterior lab bench analysis; for example, optical microscopy for phytoplankton species identification and analytical chemistry procedures for toxin profiles in the case of contaminated seafood or freshwater. Those are time-consuming activities and mean high costs in terms of necessary instrumental and specialized staff. Due to that the frequency of sampling and processing and, consequently, data delivery time lapse are under the desirable objectives. Ideally, the early warning systems should be continuously working devices distributed in proper sampling points and be able to collect and analyze representative samples of plankton and deliver specific data about their biological composition and toxinogenicity on a regular basis. In fact, there are already manufactured available devices with approaches not so far from this challenging goal. McLane Research Laboratories, Inc. (Massachusetts, USA) produces automated technologies for in situ water sampling coupled to analysis. One of them, the Environmental Sample Processor (ESP; see Figure 2.1), developed by the Monterey Bay Aquarium Research Institute, provides in situ collection and analysis of water samples from the subsurface ocean (“Environmental Sample Processor|Mclane Research Labs,” n.d.) (http://www.mclanelabs.com/master_page/product-type/ samplers/environmental-sample-processor). The instrument is an electromechanical fluidics system designed to collect discrete water samples, concentrate microorganisms or particles and automate application of molecular probes in order to identify microorganisms and their gene products. Data generated are then available for remote retrieval and analysis in near real time. The system is a modular design consisting of a core sample processor (the ESP), analytical modules and sampling modules. The core ESP provides the primary interface between the environment and a set of DNA and antibody-based sample processing technologies that are applied onboard the instrument in real time. In addition, the ESP can be used to archive samples for a variety of analyses after the instrument is returned to a laboratory. The system provides expandability to allow installation and control of secondary analytical modules for parallel processing of collected samples. This system has been used to track HABs by analyzing water samples for the presence of toxic diatom species of the genus Pseudo-nitzschia, the dinoflagellate Alexandrium catenella and the raphidophyte Heterosigma akashiwo. Additionally, an enzyme-linked immunosorbent assay (ELISA) test for domoic acid (DA) detection and quantification was included. This represents the first example of a complete automatic environmental sampler that can monitor the HAB producer together with its toxic product, Pseudo-nitzschia and DA [10].

40

Begoña Espiña et al.

Figure 2.1: Environmental Sample Processor (ESP). Printed with permission from McLane Research Laboratories.

Sample processing modules are able to perform quantitative polymerase chain reaction (qPCR) to amplify DNA and/or detection by sandwich hybridization assay (SHA) of the microorganisms present in the sample as well. This device has a version that can operate at 4,000 m of depth, the deep-sea ESP. This is probably not so relevant for monitoring of HABs because most part of the toxin-producing phytoplankton is in the surface waters. However, it demonstrated its usefulness to survey aerobic methanotrophs [11]. ESPs, even though being promissory, present several drawbacks and issues that should be solved before being implemented in a systematic way as monitoring tools; ESP devices are huge and very heavy implying that their deployment and mooring are really complicated and expensive procedures. Additionally, those systems are quite expensive (more than $200 K) and present serious limitations concerning the number of samples that can be analyzed per deployment; typically, up to 4 HABs (consisting of several species and toxins). Additionally, the range of toxins and their producers that can be analyzed should be increased. ESP can only be used to detect DA and saxitoxin (STX) and their producers so far. Currently, ESP is in the prototyping stage for its 3 G ESP. This version has been engineered to fit into the payload of a long-range autonomous underwater vehicle and deployment may be up to 3 weeks and cover about 1,800 km. Moreover, analytical capabilities include onboard digital PCR and metabolite/toxin detection via reusable miniature

2 New techniques in environment monitoring

41

SPR sensor chips. Several field trials of the 3 G ESP have been completed successfully in Monterey Bay, California, including an end-to-end test of the SPR-based measurement of DA during one flight in early 2015 [12]. Alternately, the Imaging FlowCytobot (IFCB) is an in situ automated submersible imaging flow cytometer that generates images of particles in-flow taken from the aquatic environment up to 40 m of depth. This system would allow to image all the planktonic cells present in seawater. The IFCB uses a combination of flow cytometry and video technology to capture high-resolution (HR) images of suspended particles. Laser-induced fluorescence and light scattering from individual particles are measured and used to trigger targeted image acquisition; the optical and image data are then transmitted to shore in real time. Collected images during continuous monitoring can be processed externally with automated image classification software. Images can be classified to the genus or even species level with demonstrated accuracy comparable to that of human experts. IFCB generates HR (1 µm) images of suspended particles in the size range 30 psu [96, 208]. In the brackish Baltic Sea, the only low salinity Alexandrium, A. ostenfeldii, grows best at 10 psu [38, 91]. Salinity can be a strong driver of local adaptation and even result in genetic differentiation. Contrasting reaction norms in oceanic and freshwater-influenced A. ostenfeldii populations [91] are reflected by phylogenetic differentiation patterns [64]. To some extent, patterns in reaction norms to salinity seem to be species related. Generally, A. minutum and A. ostenfeldii (syn. with A. peruvianum) have wider salinity reaction norms than the A. tamarense complex sensu [48, 209] or A. tamiyavanichi, which are less tolerant to low salinities [210]. It is not clear whether this reflects or defines differences in the habitat ranges between the species. Presently, A. ostenfeldii blooms are expanding in low salinity habitats worldwide. The new bloom populations represent a low salinity adapted phylogenetic group [211, 212]. Nevertheless, differences in the shapes of their reaction norms reflect specific conditions of the respective water bodies [213]. Many studies mention optimum salinities of 25 to 30 for different Alexandrium species, which is the typical range in most periodically freshwater influenced coastal environments where the genus forms blooms [96, 214]. Of these, many species and populations can tolerate salinities of 10 psu and still grow well at 15, implying that freshening of coastal waters as a predicted consequence of climate change [174] will not be physiologically limiting for most Alexandrium spp. [169, 199]. Though growth rates as such may not be affected by lower salinities, associated morphological and physiological traits such as cell size and pigment content might be influenced

150

Shauna Murray, Uwe John, Henna Savela and Anke Kremp

significantly [96]. Climate-related salinity changes might affect Alexandrium blooms in brackish low salinity habitats. Predicted salinity reduction (from 6 to 4 psu) might drive the species to the salinity tolerance limits [38, 176] and this may counteract the stimulating effects observed at increased temperature and CO2. Conversely, salinity effects on the growth of Alexandrium are dependent on other environmental factors. At suboptimal temperatures, the optimum salinity interval becomes narrower in the western Mediterranean A. minutum [96]. Though this concerns temperatures that are lower than at present, such interactive effects may generally be responsible for reduced tolerances to decreasing salinities and have been reported whenever tested experimentally or in modeling studies [176]. Salinity effects on toxin content and composition are similarly complex and contradictory as temperature effects. A number of studies report a link between low salinities and high toxin content for different Alexandrium species [178, 199, 207] suggesting that suboptimal salinity conditions favor STX production. However, such a relationship could not be substantiated as growth rates (reflecting favorable conditions) are typically not correlated with toxin content [98, 167, 199], suggesting that salinity affects toxicity independently from growth rates. Reports of high cellular toxin quota at high and favorable salinities [91, 199], or the opposite [167], support the concept of specific responses to salinity. Also effects on toxin composition are inconsistent, ranging from no compositional changes detected in Canadian A. catenella [98] to significant changes in individual analog proportions [91]. These authors also reported complete absence of PSTs from high salinity adapted A. ostenfeldii populations, but concluded that this was not a direct salinity effect but genetically predetermined due to the lack of an essential motif of the sxt gene [91]. Very complex changes in analog composition were observed in A. minutum when exposed to different salinities at different temperatures [167], once more emphasizing that toxin patterns are the result of interactive effects of different environmental parameters. With the available information on the relationship between salinity and PSTs, conclusive predictions on consequences of salinity changes for toxicity and possible impacts of Alexandrium spp. blooms cannot be made. Freshening of surface waters could affect Alexandrium blooms and toxic outbreaks by modifying the patterns of water column stratification. Dinoflagellates typically benefit from stratified conditions since these lead to the exclusion of nonmotile competitors and allow motile dinoflagellates to actively aggregate. Stratification due to river discharge may be associated with increased supply of humic substances, promoting growth of dinoflagellates [179]. Alexandrium blooms are commonly associated with stratified water [208, 215]. In temperate coastal waters, stratification often occurs at times of increased freshwater runoff due to heavy rainfall or ice melt in the spring. River plumes in estuaries have been demonstrated to play an important role in bloom development and shellfish toxicity of A. catenella in high latitude waters [196, 216]. The relationship of salinity stratification and bloom formation was found to be a

4 Alexandrium spp.: genetic and ecological factors

151

result of both the accumulation due to physical processes and active proliferation in the low salinity layer [215]. Interestingly, resting cyst formation of some Alexandrium species seems to be correlated with low salinities [216], which might be a strategy to maintain the seed beds in the low salinity river plume areas and assure persistent reseeding despite offshore advection of surface layer cells into higher salinity waters. In the Gulf of Maine such a mechanism apparently contributes to the formation of widely dispersed A. catenella blooms [126]. Such an indirect relationship of low salinity and favorable habitat conditions was shown to drive the ongoing expansion of A. pseudogonyaulax in coastal N European waters [217] and predicted for other Alexandrium species in a modeling study specifically considering the interaction of climate factors and habitat conditions [218].

4.6 Adaptation to changing climate conditions Despite the extensive efforts to characterize the effects of climate-related environmental variables on different Alexandrium species and regional Alexandrium blooms, it remains a challenge to predict the response of these harmful dinoflagellates to climate change and assess potential consequences related to their toxicity. “Case”specific, often contrasting, results and reports on factor interactions at different levels have generated a complex picture that cannot easily be explained. This situation is emphasized by the different assessment outcomes regarding climate change impacts on Alexandrium blooms and PST events [219, 220]. These inconsistencies might also point to crucial gaps in our understanding or insufficiency of approaches. Most of the climate change-related environmental effects on proliferation and toxicity of Alexandrium determined so far reflect immediate short-term responses in simplified systems. The respective laboratory studies were typically performed using single isolates. The few studies that have used more than one strain per species [43, 102, 103, 199] determined in some cases considerable differences in response patterns among the tested isolates. This is in some ways unsurprising and emphasizes the need to consider intraspecific variability in studies aiming to understand the effects of climate change on Alexandrium species. In fact, variability in ecologically relevant traits and the genetic basis of such phenotypic variability has long been recognized [221] in phytoplankton and has recently become the subject of intense research in the genus Alexandrium as well. Growing evidence suggests that considerable genetic diversity exists within Alexandrium populations [95, 142, 189]. When coupled to diversity of adaptively significant phenotypic traits, such standing genetic diversity provides the basis for selection [222]. Large variability in lytic capacity and cellular toxin content was found in a bloom population of A. catenella (former A. tamarense) from the North Sea [95]; this indicates that genetic diversity

152

Shauna Murray, Uwe John, Henna Savela and Anke Kremp

reflects diversity of the phenotype. Considerable intraspecific diversity in STX production has previously been shown for species of Alexandrium [8]. Studies on A. ostenfeldii revealed a surprising intraspecific diversity of cyst populations in response to climate factors, suggesting that the seed pool of Baltic A. ostenfeldii populations contains genotypes that will be favored by future temperature, pCO2 and salinity conditions (Figure 4.2 [140]). Some but not all of the favored genotypes had high cellular toxin contents or high proportions of STX. These results indicate that selection from standing genetic variation is an important mechanism of adaptation to changing conditions.

20 °C/385 μatm

24 °C/750 μatm

1.0

Growth rate (μd–1)

0.8 *

*

0.6 0.4 0.2 0.0

(a)

(c)

1.2

PST STX

1.0 0.8 0.6

(b)

*

*

AOF0924

0.2

AOF0920

0.4 *

AOF0935

AOF0933

AOF0930

AOF0929

AOF0917

AOF0909

AOF0935

AOF0933

AOF0930

AOF0929

AOF0924

AOF0920

AOF0917

0.0 AOF0909

PST/STX content (pg cell–1 )

1.4

(d)

Figure 4.2: Growth rates and corresponding cellular saxitoxin content (total PST and STX) of eight clonal Alexandrium ostenfeldii isolates grown at present-day bloom conditions (a, b) and greenhouse conditions (c, d). Asterisks denote significant changes. Data adapted from Kremp et al. [43].

4 Alexandrium spp.: genetic and ecological factors

153

Long-term evolutionary response to climate change scenarios, including the incorporation of new mutations into the gene pool, is also an important mechanism of adaptation. Only two studies have so far addressed long-term adaptation in Alexandrium. In a study of A. minutum, evidence was found of change after 2 years of exposure to greenhouse conditions when performing reciprocal growth experiments with acclimated and nonacclimated subclones [223]. Interestingly, the evolution of toxicity followed a pattern of neutral evolution, or random mutation, as indicated by large variations of toxin quota among ten long-term adapted subclones [223]. The only other long-term study on A. pacificum emphasized the importance of community effects in long-term adaptations [200]. These authors did not detect a direct effect on growth after 12 months of exposure to high CO2, despite increased competitive success of adapted A. pacificum in reciprocal community experiments [200]. Studies of the regulation and expression of STX synthesis genes, for multiple clones of more of the STX-producing species of Alexandrium, will provide important insights into the processes behind the observed effects of environmental variables on toxin patterns. Finally, numerical models accommodating the complexity of interactive factors and their effects on different organizational levels, including life cycle stages, will enable us to better predict the consequences of climate change for Alexandrium blooms and toxicity in different habitats around the world.

References [1] [2] [3] [4] [5]

[6] [7] [8]

[9]

Llewellyn L, Negri A, Robertson A. Paralytic shellfish toxins in tropical oceans. Toxin Rev 2006;25:159–196. Azanza RV, Taylor FJ. Are Pyrodinium blooms in the Southeast Asian region recurring and spreading? A view at the end of the millennium. Ambio 2001;30:356–364. Anderson DM. Bloom dynamics of toxic Alexandrium species in the Northeastern US. Limnology and Oceanogr 1997;42:1009–1022. Anderson DM, Glibert PM, Burkholder JM. Harmful algal blooms and eutrophication: nutrient sources, composition, and consequences. Estuaries and Coasts 2002;25:704–726. Campbell A, Hudson D, McLeod C, Nicholls C, Pointon A Tactical Research Fund: Review of the 2012 paralytic shellfish toxin event in Tasmania associated with the dinoflagellate alga, Alexandrium tamarense, FRDC Project 2012/060 Appendix to the final report. SafeFish; Ade- laide: 2013. Okolodkov YB. The global distributional patterns of toxic, bloom dinoflagellates recorded from the Eurasian Arctic. Harmful Algae 2005;4:351–369. Lim PT, Usup G, Leaw CP, Ogata T. First report of Alexandrium taylori and Alexandrium peruvianum (Dinophyceae) in Malaysia waters. Harmful Algae 2005;4:391–400. Anderson D, Kulis D, Doucette G, Gallagher J, Balech E. Biogeography of toxic dinoflagellates in the genus Alexandrium from the northeastern United States and Canada. Mar Biol 1994;120(3):467–478. Touzet N, Franco JM, Raine R. Morphogenetic diversity and biotoxin composition of Alexandrium (Dinophyceae) in Irish coastal waters. Harmful Algae 2008;7:782–797.

154

[10]

[11] [12]

[13]

[14]

[15] [16] [17] [18]

[19]

[20] [21] [22]

[23] [24]

[25]

[26]

[27]

Shauna Murray, Uwe John, Henna Savela and Anke Kremp

Farrell H, Brett S, Ajani P, Murray S. Distribution of the genus Alexandrium (Halim) and paralytic shellfish toxins along the coastline of New South Wales, Australia. Mar Poll Bull 2013;72:133–145. Usup G, Leaw CP, Asmat A, Lim PT. Alexandrium (Dinophyceae) species in Malaysian waters. Harmful Algae 2002;1:265–275. Penna A, Fraga S, Masó M, Giacobbe MG, Bravo I, Garcés E, Vila M, Bertozzini E, Andreoni F, Luglié A, Vernesi C. Phylogenetic relationships among the Mediterranean Alexandrium (Dinophyceae) species based on sequences of 5.8S gene and Internal Transcript Spacers of the rRNA operon. European J Phycol 2008;43:163–178. MacKenzie L, de Salas M, Adamson J, Beuzenberg V. The dinoflagellate genus Alexandrium (Halim) in New Zealand coastal waters: comparative morphology, toxicity and molecular genetics. Harmful Algae 2004;3(1):71–92. Ruiz Sebastián C, Etheridge SM, Cook PA, O’Ryan C, Pitcher GC. Phylogenetic analysis of toxic Alexandrium (Dinophyceae) isolates from South Africa: implications for the global phylogeography of the Alexandrium tamarense species complex. Phycologia 2005;44:49–60. Smayda TJ. Harmful algal blooms: Their ecophysiology and general relevance to phytoplankton blooms in the sea. Limnol Oceanogr 1997;42:1137–1153. Smayda TJ. Adaptive ecology, growth strategies and the global bloom expansion of dinoflagellates. J Oceanogr 2002;58:281–294. Cembella A. Chemical ecology of eukaryotic microalgae in marine ecosystems. Phycologia 2003;42:420–447. Tillmann U, Alpermann T, John U, Cembella A. Allelochemical interactions and short-term effects of the dinoflagellate Alexandrium on selected photoautotrophic and heterotrophic protists. Harmful Algae 2008;7:52–64. Tillmann U, Alpermann T, da Purificação RC, Krock B, Cembella A. Intra-population clonal variability in allelochemical potency of the toxigenic dinoflagellate Alexandrium tamarense. Harmful Algae 2009;8:759–769. Legrand C, Rengefors K, Fistarol GO, Granéli E. Allelopathy in phytoplankton – Biochemical, ecological and evolutionary aspects. Phycologia 2003;42:406–419. Selander E, Thor P, Toth G, Pavia H. Copepods induce paralytic shellfish toxin production in marine dinoflagellates. Proc R Soc London 2006;B 273:1673–1680. Anderson DM,. Physiology and bloom dynamics of toxic Alexandrium species, with emphasis on life cycle transitions. In: Anderson DM, Cembella AD, Hallegraeff GM [Eds.] Physiological ecology of harmful algal blooms. NATO ASI Series. Vol. G 41; Heidelberg; Springer-Verlag Berlin: 1998. 29–48. Legrand C, Carlsson P. Uptake of high molecular weight dextran by the dinoflagellate Alexandrium. Aquat Microb Ecol 1998;16:81–86. Tillmann U, John U. Toxic effects of Alexandrium spp. on heterotrophic dinoflagellates: an allelochemical defence mechanism independent of PSP toxins. Mar Ecol Prog Ser 2002;230: 47–58. Anderson DM, Alpermann TJ, Cembella AD, Collos Y, Masseret E, Montresor M. The globally distributed genus Alexandrium: Multifaceted roles in marine ecosystems and impacts on human health. Harmful Algae 2012;14:10–35. Gettings RM, Townsend DW, Thomas MA, Karp-Boss L. Dynamics of late spring and summer phytoplankton communities on Georges Bank, with emphasis on diatoms, Alexandrium spp, and other dinoflagellates. Deep Sea Res II 2014;TS;103:120–138. Giacobbe MG, Oliva FD, Maimone G. Environmental factors and seasonal occurrence of the dinoflagellate Alexandrium minutum, a PSP potential producer, in a Mediterranean lagoon. Estuar Coast Shelf Sci 1996;42(5):539–549.

4 Alexandrium spp.: genetic and ecological factors

155

[28] McGillicuddy DJ Jr, Brosnahan ML, Couture DA, He R, Keafer BA, Manning JP, Pilskaln C, Townsend DW, Anderson DM. A red tide of Alexandrium fundyense in the Gulf of Maine. Deep Sea Res II 2014;103:174–184. [29] Anderson DM, Courture DA, Kleindinst JL, Kaefer BA, McGillicuddy DJ, Martin J, Richlen M, Hickey JM, Solow AR. Understanding interannual, decadal level variability in paralytic shellfish poisoning toxicity in the Gulf of Maine: The HAB Index. Deep Sea Res II 2014;TS;103: 264–276. [30] Lewitus AJ, Horner RA, Caron DA, Garcia-Mendoza E, Hickey BM, Hunter M, Huppert DD, Kudela RM, Langlois GW, Largier JL. Harmful algal blooms along the North American west coast region: History, trends, causes, and impacts. Harmful Algae 2012;19:133–159. [31] He R, McGillicuddy DJ, Keafer BA, Anderson DM. Historic 2005 toxic bloom of Alexandrium fundyense in the western Gulf of Maine: 2. Coupled Biophysical Numerical Modeling. Journal of Geophysical Research-Oceans 2008;113:C07040. [32] McCoy GR, McNamee S, Campbell K, Elliot CT, Fleming GTA, Raine R. Monitoring a toxic bloom of Alexandrium minutum using novel microarray and multiplex surface plasmon resonance biosensor technology. Harmful Algae 2014;32:40–48. [33] Moore SK, Mantua NJ, Hickey BM, Trainer VL. Recent trends in paralytic shellfish toxins in Puget Sound, relationships to climate, and capacity for prediction of toxic events. Harmful Algae 2006;8:463–477. [34] Pitcher GC, Cembella AD, Joyce LB, Larsen J, Probyn TA, Sebastian CR. The dinoflagellate Alexandrium minutum in Cape Town harbour (South Africa): bloom characteristics, phylogenetic analysis and toxin composition. Harmful Algae 2007;6(6):823–836. [35] Burson A, Matthijs HCP, de Bruijne W, Talens R, Hoogenboom R, Gerssen A, Visser PM, Stomp M, Steur K, van Scheppingen Y, Huisman J. Termination of a toxic Alexandrium bloom with hydrogen peroxide. Harmful Algae 2014;31:125–135. [36] Konovalova GV. The morphology of Alexandrium ostenfeldii (Dinophyta) from littoral waters of eastern Kamchatka. Botanichyeskii Zhurnal (Leningrad) 1991;76:79–94. [37] Penna A, Garcés E, Vila M, Giacobbe MG, Fraga S, Lugliè A, Bravo I, Bertozzini E, Vernesi C. Alexandrium catenella (Dinophyceae), a toxic ribotype expanding in the NW Mediterranean Sea. Mar Biol 2005;148:13–23. [38] Kremp A, Lindholm T, Dreßler N, Erler K, Gerdts G, Eirtovaara S, Leskinen E. Bloom forming Alexandrium ostenfeldii (Dinophyceae) in shallow waters of the Åland Archipelago, Northern Baltic Sea. Harmful Algae 2009;8:318–328. [39] Hallegraeff GM. Transport of toxic dinoflagellates via ship’s ballast water: bioeconomic risk assessment and efficacy of possible ballast water management strategies. Mar Ecol Progr Ser 1998;168:297–309. [40] Lilly EL, Kulis DM, Gentien P, Anderson DM. Paralytic shellfish poisoning toxins in France linked to a human-introduced strain of Alexandrium pacificum from the western Pacific: Evidence from DNA and toxin analysis. J Plankton Res 2002;24:443–452. [41] Hattenrath TK, Anderson DM, Gobler CJ. The influence of anthropogenic nitrogen loading and meteorological conditions on the dynamics and toxicity of Alexandrium funduyense blooms in a New York estuary. Harmful Algae 2010;9:402–412. [42] Garcés E, Camp J. Habitat changes in the Mediterranean Sea and the consequences for harmful algal blooms formation. In: Stambler N [ed.] Life in the Mediterranean Sea: A Look at Habitat Changes. Hauppauge, NY; Nova Science Publishers, Inc.: 2012. 519–541. [43] Kremp A, Godhe A, Egardt J, Dupont S, Suikkanen S, Casabianca S, Penna A. Intra-specific variability in the response of bloom forming marine microalgae to changing climatic conditions. Ecol Evol 2012;2:1195–1207.

156

Shauna Murray, Uwe John, Henna Savela and Anke Kremp

[44] Van de Waal DB, Smith VH, Declerck SAJ, Stam ECM, Elser JJ. Stoichiometric regulation of phytoplankton toxins. Ecol Lett 2014:1–7. Doi: 10.1111/ele.12280. [45] Halim Y. Alexandrium minutum, n gen. n sp.dinoflagelle provocant des eaux rouges. Vie Milieu 1960;11:102–105. [46] Balech E. The genus Alexandrium Halim (Dinoflagellata). Sherkin Island, Cork, Ireland; Sherkin Island Press: 1995. [47] Bolch CJS, de Salas MF. A review of the molecular evidence for ballast water introduction of the toxic dinoflagellates Gymnodinium catenatum and the Alexandrium. Harmful Algae 2007;6(4):465–485. [48] John U, Litaker RW, Montresor M, Murray S, Brosnahan ML, Anderson DM. Formal revision of the Alexandrium tamarense species complex (Dinophyceae) taxonomy: the introduction of five species with emphasis on molecular-based (rDNA) classification. Protist 2014;165(6): 779–804. [49] Murray SA, Hoppenrath M, Orr RJS, Bolch C, John U, Diwan R, Yauwenas R, Harwood T, de Salas M, Neilan B, Hallegraeff G. Alexandrium diversaporum sp. nov, a new non-saxitoxin producing species: Phylogeny, morphology, and sxtA genes. Harmful Algae 2014;31:54–65. [50] Leaw CP, Lim PT, Ng BK, Cheah MY, Ahmad A, Usup G. Phylogenetic analysis of Alexandrium species and Pyrodinium bahamense (Dinophyceae) based on theca morphology and nuclear ribosomal gene sequence. J Phycol 2005;44(5):550–565. [51] Kim KY, Yoshida M, Fukuyo Y, Kim CH. Morphological observation of Alexandrium tamarense (Lebour) Balech, A. catenella (Whedon et Kofoid) Balech and one related morphotype (Dinophyceae) in Korea. Algae 2002;17:11–19. [52] Hansen PJ, Cembella AD, Moestrup Ø. The marine dinoflagellate Alexandrium ostenfeldii: paralytic shellfish toxin concentration, composition, and toxicity to a tintinnid ciliate. J Phycol 1992;28:597–603. [53] Gayoso AM, Fulco VK. Occurrence patterns of Alexandrium tamarense (Lebour) Balech populations in the Golfo Nuevo (Patagonia, Argentina), with observations on ventral pore occurrence in natural and cultured cells. Harmful Algae 2006;5:233–241. [54] Orlova TY, Morozova TV, Gribble KE, Kulis DM, Anderson DM. Dinoflagellate cysts in recent marine sediments from the east coast of Russia. Botanica marina 2004;47:184–201. [55] Orr RJS, Stüken A, Rundberget T, Eikrem W, Jakobsen KS. Improved phylogenetic resolution of toxic and non-toxic Alexandrium strains using a concatenated rDNA approach. Harmful Algae 2011;10(6):676–688. [56] Murray SA, Wiese M, Brett S, Orr R, Neilan BA, Hallegraeff G. A reinvestigation of saxitoxin production and sxtA in the ‘non-toxic’ Alexandrium tamarense Group V clade. Harmful Algae 2012;18:96–104. [57] Balech E. The genus Alexandrium or Gonyaulax of the tamarensis group. In: Anderson DM, White AW, Baden DG [eds] Toxic Dinoflagellates, Proceedings of the Third International Conference on Toxic Dinoflagellates. New York; Elsevier: 1985. 33–38. [58] Hallegraeff G, Bolch C, Blackburn S, Oshima Y. Species of the toxigenic dinoflagellate genus Alexandrium in southeastern Australian waters. Bot Mar 1991;34(6):575–588. [59] Nguyen-Ngoc L. An autecological study of the potentially toxic dinoflagellate Alexandrium affine isolated from Vietnamese waters. Harmful Algae 2004;3(2):117–129. [60] MacKenzie L, Todd K. Alexandrium camurascutulum sp. nov. (Dinophyceae): a new dinoflagellate species from New Zealand. Harmful Algae 2002;1(3):295–300. [61] Montresor M, John U, Beran A, Medlin LK. Alexandrium tamutum sp. nov. (Dinophyceae): A new nontoxic species in the genus Alexandrium. J Phycol 2004;40(2):398–411.

4 Alexandrium spp.: genetic and ecological factors

157

[62] John U, Fensome RA, Medlin LK. The Application of a molecular clock based on molecular sequences and the fossil record to explain biogeographic distributions within the Alexandrium tamarense ’species complex’ (Dinophyceae). Mol Biol Evol 2003;20:1015–1027. [63] Lilly EL, Halanych KM, Anderson DM,. Species boundaries and global biogeography of the Alexandrium tamarense complex (Dinophyceae). J Phycol 2007;43(6):1329–1338. [64] Kremp A, Tahvanainen P, Litaker W, Krock B, Suikkanen S, Leaw CP, Tomas C. Phylogenetic relationships, morphological variation and toxin patterns in the Alexandrium ostenfeldii (Dinophyceae) complex: implications for species boundaries and identities. J Phycol 2014;50: 81–100. [65] Costas E, Zardoya R, Bautista J, Garrido A, Rojo C, López-Rodas V. Morphospecies vs. genospecies in toxic marine dinoflagellates: an analysis of Gymnodinium catenatum/ Gyrodinium impudicum and Alexandrium minutum/A. lusitanicum using antibodies, lectins, and gene sequences1. J Phycol 1995;31(5):801–807. [66] Touzet N, Franco JM, Raine R. Characterization of nontoxic and toxin producing strains of Alexandrium minutum (Dinophyceae) in Irish coastal waters. Appl Environ Microbiol 2007;73 (10):3333–3342. [67] Destombe C, Cembella AD, Murphy CA, Ragan MA. Nucleotide sequence of the 18S ribosomal RNA genes from the marine dinoflagellate Alexandrium tamarense (Gonyaulacales, Dinophyta). Phycologia 1992;31:121–124. [68] Orlova TY, Selina MS, Lilly EL, Kulis DM, Anderson DM. Morphogenetic and toxin composition variability of Alexandrium tamarense (Dinophyceae) from the east coast of Russia. Phycologia 2007;46(5):534–548. [69] Taylor FJR. Toxic dinoflagellates: Taxonomic and biogeographic aspects with emphasis on Protogonyaulax. In: Ragelis EP [ed] Seafood Toxins. Washington, D.C.; American Chemical Society: 1984. 77–97. [70] Miranda LN, Zhuang YY, Zhang H, Lin S. Phylogenetic analysis guided by intragenomic SSU rDNA polymorphism refines classification of “Alexandrium tamarense” species complex. Harmful Algae 2012;16:35–48. [71] Scholin C, Herzog AM, Sogin M, Anderson DM. Identification of group- and strain-specific genetic markers for globally distributed Alexandrium (Dinophyceae). II. Sequence analysis of a fragment of the LSU rDNA gene. J Phycol 1994;30:999–1011. [72] Wang L, Zhuang YY, Zhang H, Lin X, Lin S. DNA barcoding species in Alexandrium tamarense complex using ITS and proposing designation of five species. Harmful Algae 2014;31: 100–113. [73] Stern RF, Horak A, Andrew RL, Coffroth M-A, Andersen RA, et al. Environmental Barcoding Reveals Massive Dinoflagellate Diversity in Marine Environments. PLoS ONE 2010;5(11): e13991. [74] Hoppenrath M, Leander BS. Dinoflagellate Phylogeny as Inferred from Heat Shock Protein 90 and Ribosomal Gene Sequences. PLoS ONE 2010;5(10):e13220. [75] Murray SA, et al. Gene duplication, loss and selection in the evolution of saxitoxin biosynthesis in alveolates. Mol Phylogenet Evol 2015;92:165–180. [76] Kim KY, Yoshida M, Kim CH. Molecular phylogeny of three hitherto unreported Alexandrium species: Alexandrium hiranoi, Alexandrium leei and Alexandrium satoanum (Gonyaulacales, Dinophyceae) inferred from the 18S and 26S rDNA sequence data. Phycologia 2005;44: 361–368. [77] Tang YZ, Kong L, Holmes MJ. Dinoflagellate Alexandrium leei (Dinophyceae) from Singapore coastal waters produces a water-soluble ichthyotoxin. Marine Biology 2007;150(4):541–549. [78] Masseret E, Grzebyk D, Nagai S, Genovesi B, Lasserre B, Laabir M, Collos Y, Vaquer A, Berrebi P. Unexpected genetic diversity among and within populations of the toxic

158

[79] [80] [81] [82] [83] [84]

[85]

[86] [87]

[88]

[89] [90]

[91] [92]

[93]

[94]

[95]

[96]

Shauna Murray, Uwe John, Henna Savela and Anke Kremp

dinoflagellate Alexandrium catenella as revealed by nuclear microsatellite markers. Appl Environ Microbiol 2009;75:2037–2045. Wiese M, D’Agostino PM, Mihali TK, Moffitt MC, Neilan BA. Neurotoxic alkaloids: Saxitoxin and its analogs. Mar Drugs 2010;8:2185–2211. Oshima Y. Postcolumn derivatization liquid chromatography method for PSP toxins. JAOAC 1995;78(2):528–532. Shimizu Y, Norte M, Hori A, Genenah A, Kobayashi M. Biosynthesis of saxitoxin analogues: the unexpected pathway. J Am Chem Soc 1984;106:6433–6434. Shimizu Y. Microalgal metabolites. Chem Rev 1993;93:1685–1698. Kellmann R, Neilan BA. Biochemical characterization of paralytic shellfish toxin biosynthesis in vitro. J Phycol 2007;43:497–508. Kellmann R, Mihali TK, Jeon YJ, Pickford R, Pomati F, Neilan BA. Biosynthetic intermediate analysis and functional homology reveal a saxitoxin gene cluster in cyanobacteria. Appl Environ Microbiol 2008;74:4044–4053. Band-Schmidt CJ, Lilly EL, Anderson DM. Identification of Alexandrium affine and A. margalefii (Dinophyceae) using DNA sequencing and LSU rDNA-based RFLP-PCR assays. Phycologia 2003;42(3):261–268. Stüken A, Orr RJS, Kellmann R, Murray SA, Neilan BA, Jakobsen KS. Discovery of nuclearencoded genes for the neurotoxin saxitoxin in dinoflagellates. PloS One 2011;6(5):e20096. Wang D-Z, Shu-Gang Z, Gu H-F, Chana LL, Hong H. Paralytic shellfish toxin profiles and toxin variability of the genus Alexandrium (Dinophyceae) isolated from the Southeast China Sea. Toxicon 2006;48:138–151. Ciminiello P, Fattorusso E, Forino M, Montresor M. Saxitoxin and neosaxitoxin as toxic principles of Alexandrium andersoni (Dinophyceae) from the Gulf of Naples, Italy. Toxicon 2000;38(12):1871–1877. Sampedro N, Franco JM, Zapata M, Riobó P, Garcés E, Penna A, Camp J. The toxicity and intraspecific variability of Alexandrium andersonii Balech. Harmful Algae 2013;25:26–38. Yang I, John U, Beszteri S, Glöckner G, Krock B, Goesmann A, Cembella A. Comparative gene expression in toxic versus non-toxic strains of the marine dinoflagellate Alexandrium minutum BMC Genomics 2010;11:248. Suikkanen S, Kremp A, Hautala H, Krock B. Effects of salinity on growth rate and PST/ spirolide production of the dinoflagellate Alexandrium ostenfeldii. Harmful Algae 2013;26:52–59. Menezes M, Varela D, de Oliveira Troença LA, da Silva Tamanaha M, Paredes J. Identification of the toxic algae Alexandrium tamiyavanichi (Dinophyceae) from Northeastern Brazil: a combined morphological and rDNA sequence (partial LSU and ITS) approach. J Phycol 2010;46:1239–1251. Cho Y, Hiramatsu K, Ogawa M, Omura T, Ishimaru T, Oshima Y. Non-toxic and toxic subclones obtained from a toxic clonal culture of Alexandrium tamarense (Dinophyceae): Toxicity and molecular biological feature. Harmful Algae 2008;7:740–751. Alpermann TJ, Beszteri B, John U, Tillmann U, Cembella AD. Implications of life-history transitions on the population genetic structure of the toxigenic marine dinoflagellate Alexandrium tamarense. Mol Ecol 2009;18(10):2122–2133. Alpermann TJ, Tillmann U, Beszteri B, Cembella AD, John U. Phenotypic variation and genotypic diversity in a planktonic population of the toxigenic marine dinoflagellate Alexandrium tamarense (Dinophyceae). J Phycol 2010;46:18–32. Laabir M, Collos Y, Masseret E, Grzebyk D, Abadie E, Savar V, Sibat M, Amzil Z. Influence of Environmental Factors on the Paralytic Shellfish Toxin Content and Profile of Alexandrium catenella (Dinophyceae) Isolated from the Mediterranean Sea. Mar Drugs 2013;11:1583–1601.

4 Alexandrium spp.: genetic and ecological factors

159

[97] Ignatiades L, Gotsis-Skreta O, Metaxatos A. Field and culture studies on the ecophysiology of the toxic dinoflagellate Alexandrium minutum (Halim) present in Greek coastal waters. Harmful Algae 2007;6:153–165. [98] Parkhill JP, Cembella AD. Effects of salinity, light and inorganic nitrogen on growth and toxigenicity of the marine dinoflagellate Alexandrium tamarense from northeastern Canada. J Plankton Res 1999;21:939–955. [99] Wohlrab S, Iversen M, John U, Molecular A. Co-evolutionary Context for Grazer Induced Toxin Production in Alexandrium tamarense. PLoS ONE 2010;5(11):e15039. [100] Yang I, Beszteri S, Tillmann U, Cembella A, John U. Growth- and nutrient-dependent gene expression in the toxigenic marine dinoflagellate Alexandrium minutum. Harmful Algae 2011;12:55–69. [101] Yang I, Selander E, Pavia H, John U. Grazer-induced toxin formation in dinoflagellates: a transcriptomic model study. Eur J Phycol 2011;46(1):66–73. [102] Van de Waal DB, Eberlein T, John U, Wohlrab S, Rost B. Impact of elevated pCO2 on paralytic shellfish poisoning toxin content and composition in Alexandrium tamarense. Toxicon 2014: 58–67. [103] Van de Waal DB, Tillmann U, Zhu M, Koch B, Rost B, John U. Nutrient pulse induces dynamic changes in cellular C:N:P, amino acids, and paralytic shellfish poisoning toxins in Alexandrium tamarense. Mar Ecol Prog Ser 2013;493:57–69. [104] Yang I, Beszteri S, Tillmann U, Cembella A, John U Physiological and Gene Expression Responses to Salinity Stress in Alexandrium. Proceedings of the 13th International Conference on Harmful Algae, 3–7 November 2008,Hong K, China. Ho KC, Zhou MJ, Qi YZ (eds.). Hong Kong; Environmental Publication House Hong Kong: 2010. p. 180. [105] Piel J. Metabolites from symbiotic bacteria. Nat Prod Rep 2004;21:519–538. [106] Sato S, Shimizu Y. Purification of a fluorescent product from the bacterium Moraxella: aneosaxitoxin imposter. Xunta de Galicia and Intergovernmental Oceanographic Commission of UNESCO 1998;2:465–467. [107] Prol MJ, Guisande C, Barreiro A, Mı´guez B, de la Iglesia P, et al. Evaluation of the production of paralytic shellfish poisoning toxins by extracellular bacteria isolated from the toxic dinoflagellate Alexandrium minutum. Can J Microbiol 2009;55:943–954. [108] Baker TR, Doucette GJ, Powell CL, Boyer GL, Plumley FG. GTX(4) imposters: characterization of fluorescent compounds synthesized by Pseudomonas stutzeri SF/PS and Pseudomonas/ Alteromonas PTB-1, symbionts of saxitoxinproducing Alexandrium spp. Toxicon 2003;41: 339–347. [109] Hold GL, Smith EA, Birkbeck TH, Gallacher S. Comparison of paralytic shellfish toxin (PST) production by the dinoflagellates Alexandrium lusitanicum NEPCC 253 and Alexandrium tamarense NEPCC 407 in the presence and absence of bacteria. FEMS Microbiol Ecol 2001;36:223–234. [110] Orr RJS, Stüken A, Murray S, Jakobsen KS. Evolutionary acquisition and loss of saxitoxin biosynthesis in dinoflagellates: lessons from the second “core” gene – sxtG. Appl Environ Microbiol 2013;79:2128–2136. [111] Hackett JD, Wisecaver JH, Brosnahan ML, Kulis DM, Anderson DM, Bhattacharya D, Plumley FG, Erdner DL. Evolution of saxitoxin synthesis in cyanobacteria and dinoflagellates. Mol Biol Evol 2013;30:70–78. [112] Murray SA, Wiese M, Stüken A, Brett S, Kellman R, Hallegraeff GM, Neilan BA. sxtA-based quantitative molecular assay to identify saxitoxin-producing harmful algal blooms in marine waters. Appl Environ Microbiol 2011;77:7050–7057. [113] Stüken A, et al. Paralytic shellfish toxin content is related to genomic sxtA4 copy number in Alexandrium minutum strains. Front Microbiol 2015;6(404).

160

Shauna Murray, Uwe John, Henna Savela and Anke Kremp

[114] Savela H, et al. Quantity of the dinoflagellate sxtA4 gene and cell density correlates with paralytic shellfish toxin production in Alexandrium ostenfeldii blooms. Harmful Algae 2016;52:1–10. [115] Murray SA, Ruvindy R, Kohli GS, Anderson DM, Brosnahan ML. Evaluation of sxtA and rDNA qPCR assays through monitoring of an inshore bloom of Alexandrium catenella Group 1. Sci Rep 2019;9:14532. [116] Murray SA, et al. sxtA-based quantitative molecular assay to identify saxitoxin-producing harmful algal blooms in marine waters. Appl Environ Microb 2011;77:7050–7057. [117] Gao Y, et al. High specificity of a quantitative PCR assay targeting a saxitoxin gene for monitoring toxic algae associated with paralytic shellfish toxins in the Yellow Sea. Appl Environ Microbiol 2015;81:6973–6981. [118] Stüken A, et al. Novel hydrolysis-probe based qPCR assay to detect saxitoxin transcripts of dinoflagellates in environmental samples. Harmful Algae 2013;28:108–117. [119] Penna A, et al. The sxt gene and paralytic shellfish poisoning toxins as markers for the monitoring of toxic Alexandrium species blooms. Environ Sci Technol 2015;49:14230–14238. [120] Farrell H, et al. Molecular detection of the sxtA gene from saxitoxin-producing Alexandrium minutum in commercial oysters. J Shellfish Res 2016;35:169–177. [121] Bravo I, Garces E, Diogene J, Fraga S, Sampedro N, Figueroa RI. Resting cysts of the toxigenic dinoflagellate genus Alexandrium in recent sediments from the Western Mediterranean coast, including first description of cysts of A. kutnerae and A. peruvianum. Eur J Phycol 2006;41:293–302. [122] Figueroa RI, Garces E, Bravo I. Comparative study of the life cycles of Alexandrium tamutum and Alexandrium minutum (Gonyaulacales, Dinophyceae) in culture. Harmful Algae 2007;43 (5):1039–1053. [123] Horner RA, Greengrove CL, Postel JR, Gawel JE, Davies-Vollum KS, Cox AM. Spatial distribution of benthic cysts of Alexandrium catenella in surface sediments of Puget Sound, Washington, USA. Harmful Algae 2011;11:96–105. [124] Angles S, Jordi A, Garcés E, Basterretxea G, Palanques A. Alexandrium minutum resting cyst distribution dynamics in a confined site. Deep-Sea Res Pt II 2010;57(3–4):210–221. [125] Hakanen P, Suikkanen S, Franzén J, Franzén H, Kankaanpää H, Kremp A. Bloom and toxin dynamics of Alexandrium ostenfeldii in a shallow embayment at the SW coast of Finland, northern Baltic Sea. Harmful Algae 2012;15:91–99. [126] Anderson DM, Stock CA, Keafer BA, Bronzino Nelson A, Thompson B, McGillicuddy DJ, Keller M, Matrai PA, Martin J. Alexandrium fundyense cyst dynamics in the Gulf of Maine. Deep-Sea Res Pt II 2005;2(19–21):2522–2542. [127] Mardones JI, Bolch C, Guzmán L, Paredes J, Varela D, Hallegraeff GM. Role of resting cysts in Chilean Alexandrium catenella dinoflagellate blooms revisited. Harmful algae 2016;55: 238–249. [128] Anderson DM, Rengefors K. Community assembly and seasonal succession of marine dinoflagellates in a temperate estuary – the importance of life cycle events and predation. Limnol Oceanogr 2006;51(2):860–873. [129] Nı´Rathaille A, Raine R. Seasonality in the excystment of Alexandrium minutum and Alexandrium tamarense in Irish coastal waters. Harmful Algae 2011;10:629–635. [130] Lau WLS, Law K, Liow GR, Hii KS, Usup G, Lim PT, Leaw CP. Life-history stages of natural bloom populations and the bloom dynamics of a tropical Asian ribotype of Alexandrium minutum. Harmful algae 2017;70:52–63. [131] Fischer AD, Brosnahan ML, Anderson DM. Quantitative Response of Alexandrium catenella Cyst Dormancy to Cold Exposure. Protist 2018;169(5):645–661.

4 Alexandrium spp.: genetic and ecological factors

161

[132] Natsuike M, Yokoyama K, Nishitani G, Yamada Y, Yoshinaga I, Ishikawa A. Germination fluctuation of toxic Alexandrium fundyense and A. pacificum cysts and the relationship with bloom occurrences in Kesennuma Bay, Japan. Harmful algae 2017;62:52–59. [133] Garcés E, Bravo I, Vila M, Figueroa RI, Maso´ M, Sampedro N. Relationship between vegetative cells and cyst production during Alexandrium minutum bloom in Arenys de Mar harbour (NW Mediterranean). J Plankton Res 2004;26(6):637–645. [134] Brosnahan ML, Ralston DK, Fischer AD, Solow AR, Anderson DM. Bloom termination of the toxic dinoflagellate Alexandrium catenella: Vertical migration behavior, sediment infiltration, and benthic cyst yield. Limnol Oceanogr 2017;62(6):2829–2849. [135] Anderson DM, Chrisholm SW, Watras CJ. Importance of life cycle events in the population dynamics of Gonyaulax tamarensis. Mar Biol 1983;76:179–189. [136] Bravo I, Fraga S, Figueroa RI, Pazos Y, Massanet A, Ramilo I. Bloom dynamics and life cycle strategies of two toxic dinoflagellates in a coastal upwelling system (NW Iberian Peninsula). Deep-Sea Res Pt II 2010;57(3–4):222–234. [137] Jerney J, Ahonen SA, Hakanen P, Suikkanen S, Kremp A. Generalist Life Cycle Aids Persistence of Alexandrium ostenfeldii (Dinophyceae) in Seasonal Coastal Habitats of the Baltic Sea1. J Phycol 2019. [138] Nagai S, Lian C, Yamaguchi S, Hamaguchi M, Matsuyama Y, Itakura S, Shimada H, Kaga S, Yamauchi H, Sonda Y, Nishikawa T, Kim CH, Hogetsu T. Microsatellite markers reveal population genetic structure of the toxic dinoflagellate Alexandrium tamarense (Dinophyceae) in Japanese coastal waters. J Phycol 2007;43:43–54. [139] Sundqvist L, Godhe A, Jonsson PR, Sefbom J. The anchoring effect-long-term dormancy and genetic population structure. ISME J 2018;12(12):2929. [140] Kremp A, Oja J, LeTortorec AH, Hakanen P, Tahvanainen P, Tuimala J, Suikkanen S. Diverse seed banks favour adaptation of microalgal populations to future climate conditions. Environmental microbiology 2016;18(2):679–691. [141] Erdner DL, Richlen M, McCauley LAR, Anderson DM. Intrapopulation diversity and dynamics of a widespread bloom of the toxic dinoflagellate Alexandrium fundyense. PLoS ONE 2011;6 (7):e22965. [142] Dia A, Guillou L, Mauger S, Bigeard E, Marie D, Valero M, Destombe C. Spatiotemporal changes in the genetic diversity of harmful algal blooms caused by the toxic dinoflagellate Alexandrium minutum. Mol Ecol 2014;23:549–560. [143] Spatharis S, Danielidis DS, Tsirtsis G. Recurrent Pseudo-nitzschia calliantha (Bacillariophyceae) and Alexandrium insuetum (Dinophyceae) winter blooms induced by agricultural runoff. Harmful Algae 2007;6:811–822. [144] Nagata T. Production mechanisms of dissolved organic matter. In: Kirchman DL [ed.] Microbial Ecology of the Oceans. New York; Wiley-Liss: 2000. 121–152. [145] Loureiro S, Garces E, Fernandez M, Vaque D, Camp J. Pseudo-nitzschia spp. (Bacillariophyceae) and dissolved organic matter (DOM) dynamics in the Ebro Delta (Alfacs Bay, NW Mediterranean Sea). Estuar Coast Shelf Sci 2009;83(4):539–549. [146] Stolte W, Panosso R, Gisselson LA, Granéli E. Utilization efficiency of nitrogen associated with riverine dissolved organic carbon (> 1 kDa) by two toxin-producing phytoplankton species. Aquat Microb Ecol 2002;29:97–105. [147] Fagerberg T, Carlsson P, Lundgren M. A large molecular size fraction of riverine high molecular weight dissolved organic matter (HMW DOM) stimulates growth of the harmful dinoflagellate Alexandrium minutum. Harmful Algae 2009;8:823–831. [148] Dyhrman ST, Anderson DM. Urease activity in cultures and field populations of the toxic dinoflagellate Alexandrium. Limnol Oceanogr 2003;48:647–655.

162

Shauna Murray, Uwe John, Henna Savela and Anke Kremp

[149] Ou LJ, Huang BQ, Lin LZ, Hong HS, Zhang F, Chen ZZ. Phosphorus stress of phytoplankton in the Taiwan Strait determined by bulk and single-cell alkaline phosphatase activity assays. Mar Ecol Prog Ser 2006;327:95–106. [150] Jacobson DM, Anderson DM. Widespread phagocytosis of ciliates and other protists by marine mixotrophic and heterotrophic thecate dinoflagellates. J Phycol 1996;32:279–285. [151] Tillmann U, Kremp A, Tahvanainen P, Krock B. Characterization of spirolide producing Alexandrium ostenfeldii (Dinophyceae) from the western Arctic. Harmful Algae 2014;39:259–270. [152] Tillmann U, Hansen PJ. Allelopathic effects of Alexandrium tamarense on other algae: evidence from mixed growth experiments. Aquat Microb Ecol 2009;57:101–112. [153] Weissbach A, Tillmann U, Legrand C. Allelopathic potential of the dinoflagellate Alexandrium tamarense on marine microbial communities. Harmful Algae 2010;10:9–18. [154] Fistarol GO, Legrand C, Rengefors K, Granéli E. Temporary cyst formation in phytoplankton: A response to allelopathic competitors?. Environ Microbiol 2004;6:791–798. [155] Hakanen P, Suikkanen S, Kremp A. Allelopathic activity of the toxic dinoflagellate Alexandrium ostenfeldii: intra-population variability and response of co-occurring dinoflagellates. Harmful Algae 2014;39:287–294. [156] John U, Tillmann U, Hülskötter J, Alpermann TJ, Wohlrab S, Van de Waal D Intraspecific facilitation by allelochemical mediated grazing protection within a toxigenic dinoflagellate population. Proceedings of the Royal Society B. In press; 2014. [157] Blossom HE, Markussen B, Daugbjerg N, Krock B, Norlin A, Hansen PJ. The Cost of Toxicity in Microalgae: Direct Evidence From the Dinoflagellate Alexandrium. Front Microbiol 2019:10. [158] Brandenburg KM, Wohlrab S, John U, Kremp A, Jerney J, Krock B, Van de Waal DB. Intraspecific trait variation and trade‐offs within and across populations of a toxic dinoflagellate. Ecol Lett 2018;21(10):1561–1571. [159] Tillmann U, John U, Cembell A. On the allelochemical potency of the marine dinoflagellate Alexandrium ostenfeldii against heterotrophic and autotrophic protists. J Plankton Res 2007;29:527–543. [160] Tillmann U, Krock B, Alpermann TJ, Cembella A. Bioactive compounds of marine dinoflagellate isolates from western Greenland and their phylogenetic association within the genus Alexandrium. Harmful algae 2016;51:67–80. [161] Sopanen S, Setälä O, Piiparinen J, Erler K, Kremp A. The toxic dinoflagellate Alexandrium ostenfeldii promotes incapacitation of the calanoid copepod Eurytemora affinis and Acartia bifilosa from the northern Baltic Sea. J Plankton Res 2011;33:1564–1573. [162] Smith VH. Eutrophication of freshwater and coastal marine ecosystems: a global problem. Environ Sci Pollut Res 2003;10:126–139. [163] Borkman, D. G., Smayda, T. J., Schwarz, E. N., Flewelling, L. J., & Tomas, C. R. (2014). Recurrent vernal presence of the toxic Alexandrium tamarense/Alexandrium fundyense (Dinoflagellata) species complex in Narragansett Bay, USA. Harmful algae, 32, 73–80. [164] Van der Lingen, C. D., Hutchings, L., Lamont, T., & Pitcher, G. C. (2016). Climate change, dinoflagellate blooms and sardine in the southern Benguela Current Large Marine Ecosystem. Environmental development, 17, 230–243. [165] Wells, M. L., Trainer, V. L., Smayda, T. J., Karlson, B. S., Trick, C. G., Kudela, R. M., … & Cochlan, W. P. (2015). Harmful algal blooms and climate change: Learning from the past and present to forecast the future. Harmful algae, 49, 68–93. [166] Han M, Lee H, Anderson DM, Kim B. Paralytic shellfish toxin production by the dinoflagellate Alexandrium pacificum (Chinhae Bay, Korea) in axenic, nutrient-limited chemostat cultures and nutrient-enriched batch cultures. Mar Pollut Bull 2016;104(1–2):34–43. [167] Laabir M, Jauzein C, Genovesi B, Masseret E, Grzebyk D, Cecchi P, Vaquer A, Perrin Y, Collos Y. Influence of temperature, salinity and irradiance on the growth and cell yield of the

4 Alexandrium spp.: genetic and ecological factors

[168] [169]

[170] [171] [172]

[173]

[174]

[175]

[176]

[177]

[178] [179] [180]

[181]

[182] [183] [184] [185]

163

harmful red tide dinoflagellate Alexandrium catenella colonising Mediterranean waters. J Plankton Res 2011;33:1550–1563. Reusch TBH, Boyd PW. Experimental evolution meets marine phytoplankton. Evolution 2013;67:1849–1859. Østergaard Jensen M, Moestrup Ø. Autoecology of the toxic dinoflagellate Alexandrium ostenfeldii: life history and growth at different temperatures and salinities. Eur J Phycol 1997;32:9–18. Gu H. Morphology, phylogenetic position, and ecophysiology of Alexandrium ostenfeldii (Dinophyceae) from the Bohai Sea, China. J Syst Evol 2011;49:606–616. Navarro JM, Munoz MG, Contreras AM. Temperature as a factor regulating growth and toxin content in the dinoflagellate Alexandrium catenella. Harmful Algae 2006;5:762–769. Nagai S, Matsuyam Y, Oh SJ, Itakura S. Effect of nutrients and temperature on encystment of the toxic dinoflagellate Alexandrium tamarense (Dinophyceae) isolated from Hiroshima Bay, Japan. Plankton Biol Ecol 2004;51(2):103–109. Shin HH, Baek SH, Han M-S, Oh SJ, Youn S-H, Kim YS, Kim D, Lim W-A. Resting cysts, and effects of temperature and salinity on the growth of vegetative cells of the potentially harmful species Alexandrium insuetum Balech (Dinophyceae). Harmful Algae 39:175–184. IPCC (Intergovernmental Panel on Climate Change). 2014. 5th Assessment Report. Geneva: IPCC Secretariat, World Meteorological Association; 2014. Available at: http://www.ipcc.ch/ report/ar5/ Tatters AO, Flewelling LJ, Fu F, Granholm AA, Hutchins DA. High CO2 promotes the production of paralytic shellfish poisoning toxins by Alexandrium catenella from Southern California waters. Harmful Algae 2013;30:37–43. Salgado P, Vazquez JA, Riobó P, Franco JM, Figueroa RI, Kremp A, Bravo I. A kinetic and factorial approach to study the effects of temperature and salinity on growth and toxin production by the dinoflagellate Alexandrium ostenfeldii from the Baltic Sea. PloS one 2015;10(12):e0143021. Eckford-Soper LK, Bresnan E, Lacaze JP, Green DH, Davidson K. The competitive dynamics of toxic Alexandrium fundyense and non-toxic Alexandrium tamarense: The role of temperature. Harmful algae 2016;53:135–144. Fu FX, Tatters AO, Hutchins DA. Global change and the future of harmful algal blooms in the ocean. Mar Ecol Prog Ser 2012;470:207–233. Hallegraeff GM. Ocean climate change, phytoplankton community responses, and harmful algal blooms: a formidable predictive challenge. J Phycol 2010;46:220–235. Gobler CJ, Doherty OM, Hattenrath-Lehmann TK, Griffith AW, Kang Y, Litaker RW. Ocean warming since 1982 has expanded the niche of toxic algal blooms in the North Atlantic and North Pacific oceans. Proc Natl Acad Sci U.S.A. 2017;114(19):4975–4980. Bill BD, Moore SK, Hay LR, Anderson DM, Trainer VL. Effects of temperature and salinity on the growth of Alexandrium (Dinophyceae) isolates from the Salish Sea. J Phycol 2016;52(2): 230–238. Moore SK, Johnstone JA, Banas NS, Salathe JE. Present-day and future climate pathways affecting Alexandrium blooms in Puget Sound, WA, USA. Harmful Algae 2015;48:1–11. Warns A, Hense I, Kremp A. Encystment of a cold-water dinoflagellate: From In Vitro to In Silico. J Mar Syst 2013;125:54–60. Anderson DM. Effects of temperature conditioning on development and germination of Gonyaulax tamarensis (Dinophyceae) hypnozygotes. J Phycol 1980;16(2):166–172. Cosgrove S, Nı´Rathaille A, Raine R. The influence of bloom intensity on the encystment rate and persistence of Alexandrium minutum in Cork Harbor, Ireland. Harmful Algae 2014;31:114–124.

164

Shauna Murray, Uwe John, Henna Savela and Anke Kremp

[186] Genovesi B, Laabir M, Masseret E, Collos Y, Vaquer A, Grzebyk D. Dormancy and germination features in resting cysts of Alexandrium tamarense species complex (Dinophyceae) can facilitate bloom formation in a shallow lagoon (Thau, southern France). J Plankton Res 2009;3:1209–1224. [187] Laanaia N, Vaquer A, Fiandrino A, Genovesi B, Pastoureaud A, Cecchi P, Collos Y. Wind and temperature controls on Alexandrium blooms (2000–2007) in Thau lagoon (Western Mediterranean). Harmful Algae 2013;28:31–36. [188] Figueroa RI, Vázquez JA, Massanet A, Murado MA, Bravo I. Interactive effects of salinity and temperature on planozygote and cyst formation on Alexandrium minutum (Dinophyceae). J Phycol 2011;47:13–24. [189] Tahvanainen P, Alpermann TJ, Figueroa RI, John U, Hakanen P, Nagai S, Blomster J, Kremp A. Patterns of post-glacial genetic differentiation in marginal populations of a marine microalga. PLoS ONE 2012;7(12):e53602. [190] Brandenburg KM, Velthuis M, Van de Waal DB. Meta‐analysis reveals enhanced growth of marine harmful algae from temperate regions with warming and elevated CO2 levels. Glob Chang Biol 2019. [191] McMinn A, Scott FJ. Dinoflagellates. In: Scott FJ, Marchant HJ [eds.]. Canberra & Hobart; Australian Biological Resources Study, Australian Antarctic Division: 2005. 202–250. [192] Baggesen C, Moestrup Ø, Daugbjerg N, Krock B, Cembella AD, Madsen S. Molecular phylogeny and toxin profiles of Alexandrium tamarense (Lebour) Balech (Dinophyceae) from the west coast of Greenland. Harmful Algae 2012;19:108–116. [193] Natsuike M, Nagai S, Matsuno K, Saito R, Tsukazaki C, Yamaguchi A, Imai I. Abundance and distribution of toxic Alexandrium tamarense resting cysts in the sediments of the Chukchi Sea and the eastern Bering Sea. Harmful Algae 2013;27:52–59. [194] Natsuike M, Saito R, Fujiwara A, Matsuno K, Yamaguchi A, Shiga N, . . . Imai I. Evidence of increased toxic Alexandrium tamarense dinoflagellate blooms in the eastern Bering Sea in the summers of 2004 and 2005. PloS one 2017;12(11):e0188565. [195] Vandersea MW, Kibler SR, Tester PA, Kristine Holderied DE, Hondolero KP, Baird S, Doroff A, Darcy Dugan R, Litaker W. Environmental factors influencing the distribution and abundance of Alexandrium catenella in Kachemak bay and lower cook inlet, Alaska. Harmful Algae 2018;77:81–92. [196] Tobin ED, Wallace CL, Crumpton C, Johnson G, Eckert GL. Environmental drivers of paralytic shellfish toxin producing Alexandrium catenella blooms in a fjord system of northern Southeast Alaska. Harmful algae 2019;88:101659. [197] Almandoz GO, Montoya NG, Hernando MP, Benavides HR, Carignan MP, Ferrario ME. Toxic strains of the Alexandrium ostenfeldii complex in southern South America (Beagle Channel, Argentina). Harmful Algae 2014;37:100–109. [198] Selina MS, Konovalova GV, Morozova TV, Orlova T. Genus Alexandrium Halim, 1960 (Dinophyta) from the Pacific Coast of Russia: Species Composition, Distribution, and Dynamics. Russ J Mar Biol 2006;32(6):321–332. [199] Etheridge SM, Roesler CS. Effects of temperature, irradiance, and salinity on photosynthesis, growth rates, total toxicity, and toxin composition for Alexandrium fundyense isolates. Deep Sea Res Part II 2005;52(19):2491–2500. [200] Tatters AO, Schnetzer A, Fu FX, Lie AAY, Caron DA, Hutchins DA. Short-versus long-term responses to changing CO2 in a coastal dinoflagellate bloom: implications for interspecific competitive interactions and community structure. Evolution 2013;67:1879–1891. [201] Eberlein T, Van de Waal DB, Rost B. Differential effects of ocean acidification on carbon acquisition in two bloom-forming dinoflagellate species. Physiol Plant 2014;151(4):468–479.

4 Alexandrium spp.: genetic and ecological factors

165

[202] Pang M, Xu J, Qu P, Mao X, Wu Z, Xin M, . . . Chen H. Effect of CO2 on growth and toxicity of Alexandrium tamarense from the East China Sea, a major producer of paralytic shellfish toxins. Harmful algae 2017;68:240–247. [203] Hattenrath-Lehmann TK, Smith JL, Wallace RB, Merlo LR, Koch F, Mittelsdorf H, Goleski JA, Anderson DM, Gobler CJ. The effects of elevated CO2 on the growth and toxicity of field populations and cultures of the saxitoxin-producing dinoflagellate, Alexandrium fundyense Limnol. Oceanogr 2015;60(1):198–214. [204] Mardones JI, Müller MN, Hallegraeff GM, Handling editor: Shubha Sathyendranath. Toxic dinoflagellate blooms of Alexandrium catenella in Chilean fjords: a resilient winner from climate change. ICES Journal of Marine Science 2016;74(4):988–995. [205] Hansen PJ. The effect of high pH on the growth and survival of marine phytoplankton: implications for species succession. Aquat Microb Ecol 2002;28:279–288. [206] Therriault JC, Painchaud J, Levasseur M. Factors controlling the occurrence of Protogonyaulax tamarensis and shellfish toxicity in the St. Lawrence Estuary: freshwater runoff and the stability of the water column. In: Anderson DM, White AW, Baden DG [editors] Toxic Dinoflagellates. New York; Elsevier Science: 1985. 141–146. [207] Grzebyk D, Béchemin C, Ward C, Vérité C, Codd G, Maestrini SY. Effects of salinity and two coastal waters on the growth and toxin content of the dinoflagellate Alexandrium minutum. J Plankton Res 2003;25:1185–1199. [208] Bravo I, Vila M, Maso M, Ramilo I, Figueroa RI. Alexandrium catenella and Alexandrium minutum blooms in the Mediterranean Sea: Toward the identification of ecological niches. Harmful Algae 2008;7:515–522. [209] Fraga S, Sampedro N, Larsen J, Moestrup Ø, Calado AJ. Arguments against the proposal 2302 by John & al. to reject the name Gonyaulax catenella Alexandrium catenella. Taxon 2015;64 (3):634–635. [210] Lewis AM, Coates LN, Turner AD, Percy L, Lewis J. A review of the global distribution of Alexandrium minutum (Dinophyceae) and comments on ecology and associated paralytic shellfish toxin profiles, with a focus on Northern Europe. J Phycol 2018;54(5):581–598. [211] Sildever S, Jerney J, Kremp A, Oikawa H, Sakamoto S, Yamaguchi M, . . . Shinada A. Genetic relatedness of a new Japanese isolates of Alexandrium ostenfeldii bloom population with global isolates. Harmful algae 2019;84:64–74. [212] Brandenburg KM, de Senerpont Domis LN, Wohlrab S, Krock B, John U, van Scheppingen Y, . . . Van de Waal DB. Combined physical, chemical and biological factors shape Alexandrium ostenfeldii blooms in the Netherlands. Harmful algae 2017;63:146–153. [213] Martens H, Van de Waal DB, Brandenburg KM, Krock B, Tillmann U. Salinity effects on growth and toxin production in an Alexandrium ostenfeldii (Dinophyceae) isolate from The Netherlands. J Plankton Res 2016;38(5):1302–1316. [214] Fabro E, Almandoz GO, Ferrario M, John U, Tillmann U, Toebe K, . . . Cembella A. Morphological, molecular, and toxin analysis of field populations of Alexandrium genus from the Argentine Sea. J Phycol 2017;53(6):1206–1222. [215] Fauchot J, Levasseur M, Roy S, Gagnon R, Weise AM. Environmental factors controlling Alexandrium tamarense (Dinophyceae) growth rate during a red tide event in the St. Lawrence Estuary (Canada). J Phycol 2005;41:263–272. [216] Weise AM, Levasseur M, Saucier FJ, Senneville S, Bonneau E, Roy S, Sauvé G, Michaud S, Fauchot J. The link between precipitation, river runoff, and blooms of the toxic dinoflagellate Alexandrium tamarense in the St. Lawrence. Can J Fish Aquat Sci 2002;59:464–473. [217] Kremp A, Hansen PJ, Tillmann U, Savela H, Suikkanen S, Voß D, . . . Krock B. Distributions of three Alexandrium species and their toxins across a salinity gradient suggest an increasing

166

[218]

[219]

[220]

[221]

[222] [223]

Shauna Murray, Uwe John, Henna Savela and Anke Kremp

impact of GDA producing A. pseudogonyaulax in shallow brackish waters of Northern Europe. Harmful algae 2019;87:101622. Townhill BL, Tinker J, Jones M, Pitois S, Creach V, Simpson SD, . . . Pinnegar JK. Harmful algal blooms and climate change: exploring future distribution changes. ICES Journal of Marine Science 2018;75(6):1882–1893. Moore SK, Mantua NJ, Kellogg JP, Newton JA. Local and large-scale climate forcing of Puget Sound oceanographic properties on seasonal to interdecadal timescales. Limnol Oceanogr 2008;53:1746–1758. Bresnan E, Davidson K, Edwards M, Fernand L, Gowen R, Hall A, Kennington K, McKinney A, Milligan S, Raine R, Silke J. Impacts of climate change on harmful algal blooms (HABs). MCCIP Science Review 2013:236–243. Brand LE. Genetic variability and spatial patterns of genetic differentiation in the reproductive rates of the marine coccolitophores Emiliana huxleyi and Gephyrocapsa oceanic. Limnology and Oceanogr 1982;27:236–245. Barrett RDH, Schluter D. Adaptation from standing genetic variation. Trends Ecol Evol 2007;23:38–44. Flores-Moya A, Rouco M, García-Sánchez MJ, García-Balboa C, González R, Costas E, LópezRodas V. Effects of adaptation, chance, and history on the evolution of the toxic dinoflagellate Alexandrium minutum under selection of increased temperature and acidification. Ecol Evol 2012;2:1251–1259.

Laura T. Kelly, Jonathan Puddick, Hugo Borges, Daniel R. Dietrich, David P. Hamilton and Susanna A. Wood

5 Potential effects of climate change on cyanobacterial toxin production 5.1 Introduction Cyanobacteria are a group of ancient oxygenic photosynthetic prokaryotic organisms originating between 3 and 4 billion years ago. They have been reported in a wide range of environments, including oceans, lakes and rivers, as well as extreme habitats such as geothermal springs, desert soils and the polar regions [1]. Cyanobacteria can exist as solitary, free-living cells or as colonies/filaments numbering a few cells to several thousand, enclosed in mucilage. While most cyanobacterial colonies/filaments and single cyanobacteria cells are microscopic, large populations become visible as mats, crusts and blooms. Populations may be planktonic (Figure 5.1(a) and (b)), suspended in the water column, or benthic, growing on bottom substrates and sometimes forming extensive mats (Figure 5.1(c) and (d)). Extensive blooms or benthic mats of cyanobacteria can be monocultures or consist of several species. These proliferations have been associated with increases in nutrient concentrations (i.e., eutrophication), which has been linked to human activities such as deforestation, agriculture and urbanization. Blooms may be catalyzed by other factors that have been linked to climate change, including increased water temperature, variations in precipitation, extended droughts and increased carbon dioxide levels (see Chapter 7). There are approximately 2,000 cyanobacteria species described worldwide and more than 50 are known to have strains which produce natural compounds which are toxic (cyanotoxins). Cyanotoxins are a threat to humans when ingested Acknowledgments: The authors thank the New Zealand Ministry of Business, Innovation and Employment (Trace metal limitation of phytoplankton growth in New Zealand lakes, C01X1711) and the Marsden Fund of the Royal Society of New Zealand (Blooming Buddies, CAW1601) for funding. SAW was supported by the National Institute of Water and Atmospheric Research Ltd. under the causes and effects of water quality degradation: eutrophication risk assessment program. DPH and SAW also acknowledge the Australian Research Council, project DP190101848: Next-generation models to predict cyanobacteria harmful algal blooms. Laura T. Kelly, Jonathan Puddick, Hugo Borges, Susanna A. Wood, Cawthron Institute, Nelson, New Zealand Daniel R. Dietrich, Human and Environmental Toxicology, University of Konstanz, Konstanz, Germany David P. Hamilton, Australian Rivers Institute, Griffith University, Brisbane, Australia https://doi.org/10.1515/9783110625738-005

168

Laura T. Kelly et al.

(a)

(b)

(c)

(d)

Figure 5.1: Images of (a and b) planktonic cyanobacterial shoreline scums (Microcystis sp.) and (c and d) benthic Microcoleus mats.

(via water supplies or accidental swallowing) or from contact (dermal or inhalation). Cyanotoxins exhibit a wide range of toxicity mechanisms including; hepatotoxicity, nephrotoxicity, neurotoxicity and dermatotoxicity [2]. Cyanotoxins can be divided into three broad groups based on chemical structures: cyclic peptides (microcystins and nodularins), alkaloids (cylindrospermopsins, saxitoxins and anatoxins) and lipopolysaccharides (LPS; Table 5.1). Table 5.1: Cyanotoxins produced by freshwater cyanobacteria (adapted from Table 3.1 of Reference [2]). Cyanotoxin

Primary target in mammals

Structural characteristics

Microcystins

Liver Protein phosphatase inhibitors

Cyclic heptapeptide

Nodularins

Liver Protein phosphatase inhibitors

Cyclic pentapeptide

Cylindrospermopsins

Liver Protein synthesis inhibitors

Tricyclic alkaloid with attached uracil

Saxitoxins

Nerve axon Blockage of voltage-gated sodium channels

Tricyclic alkaloid

Anatoxin-a

Nerve synapse Agonists of nicotinic acetylcholine receptors

Bicyclic secondary amine

169

5 Potential effects of climate change on cyanobacterial toxin production

Table 5.1 (continued ) Cyanotoxin

Primary target in mammals

Structural characteristics

Anatoxin-a(S)

Nerve synapse Acetylcholinesterase inhibitors

Organophosphate alkaloid

Lipopolysaccharides

Dermis Irritant of all exposed tissue

Lipopolysaccharide

5.1.1 Microcystins and nodularins Globally, microcystins are the most frequently found cyanotoxin [2]. Microcystins are cyclic peptides (Figure 5.2(f)) and to date, more than 250 variants have been isolated and characterized [3]. Microcystins and nodularins, both contain the unique beta-amino acid Adda, but while microcystins are heptapeptides containing five Dand two L-amino acids, nodularins contain only five amino acids (Figure 5.2(e)). While many cyanobacteria genera produce microcystins [2], the majority of nodularin reports have been from Nodularia spumigena, primarily a brackish-water species [2]. Exceptions are nodularin observed in Iningainema pulvinus from an Australian wetland [4] and a Nostoc species that grows symbiotically with cycads [5].

O O

O H N

O–

P

O

HO H 2N

O

N

O

O N+

N

NH

N H

N

(a)

OH

O

S

O

(b)

N

HN

NH

OH

OH

O

N H NH O

OH

NH NH

O

O

HN

O H N

NH CH2

O

O HN

H N O HO

OH

HN

(e)

O N

O

O O

O

OH

O

HN O

N N H

NH

(d) HO

O

H N N H

N

HN

(c)

O

NH2 O H

O

HN NH2

(f)

HN

NH2

Figure 5.2: Chemical structures of several common freshwater cyanotoxins: (a) anatoxin-a, (b) anatoxin-a(S). (c) cylindrospermopsin, (d) saxitoxin, (e) nodularin-R and (f) microcystin-LR.

170

Laura T. Kelly et al.

Each microcystin variant differs with respect to the methyl groups and two Lamino acids within the cyclic peptide ring. This results in pronounced differences in the toxicity of variants. Microcystins and nodularins are hepatotoxins that inhibit protein phosphatases 1 and 2A in affected organisms [6, 7]. Numerous incidents of animal and human poisonings have been attributed to microcystins and nodularins [8]. One of the most severe cases occurred in Brazil in 1996, when water supply to a hospital was contaminated with microcystins, with ensuing microcystin concentrations in the dialysis water of approximately 20 µg L–1. Fifty-six fatalities consequently occurred at a dialysis treatment clinic in the hospital [9].

5.1.2 Cylindrospermopsins Cylindrospermopsin is a tricyclic alkaloid (Figure 5.2(c)) that can also exist as a 7epicylindrospermopsin [10] and deoxy-cylindrospermopsin [11] analog. Cylindrospermopsins are potent protein synthesis inhibitors and cause extensive damage to the liver and kidney [12, 13]. Falconer and Humpage have also suggested that cylindrospermopsin may act directly as a tumor initiator [14]. Cylindrospermopsins have been reported from a range of cyanobacteria and in a variety of geographic locations including Europe, Asia and Australasia [15]. Cylindrospermopsin was implicated in one of the most significant cases of human poisoning from exposure to a cyanobacterial toxin when 148 people were hospitalized in 1979 with symptoms of gastro-enteritis after a local water supply on Palm Island (Australia) was dosed with copper sulfate to control a dense Raphidiopsis (previously Cylindrospermopsis) raciborskii bloom [16–18].

5.1.3 Saxitoxins Saxitoxins are alkaloids (Figure 5.2(d)) and are fast-acting neurotoxins that inhibit nerve conduction by blocking sodium channels [19]. Saxitoxins are commonly produced by marine dinoflagellates under the name of paralytic shellfish toxins. While saxitoxins from freshwater cyanobacteria have not been implicated in any cases of human intoxication [2], they have killed animals [20] and were identified in an extensive bloom of Dolichospermum (formally known as Anabaena) circinalis on the Murray Darling River (Australia) which resulted in the death of over 1,600 sheep and cattle [21].

5.1.4 Anatoxins These alkaloid toxins (Figure 5.2(a)) are powerful depolarizing neuromuscular blocking agents acting through the nicotinic acetylcholine receptor [22]. They are rapidly absorbed when ingested orally. Animals exposed to these alkaloid toxins

5 Potential effects of climate change on cyanobacterial toxin production

171

may undergo convulsions, coma, rigors, cyanosis, limb twitching, hypersalivation and death. Anatoxins have also been linked to animal and wildfowl poisonings, however there have been no verified reports of human poisonings/fatalities from anatoxin-a [8]. Anatoxins are produced by both planktonic [2] and benthic cyanobacteria species [23].

5.1.5 Anatoxin-a(S) Anatoxin-a(S) (Figure 5.2(b)), which is structurally different from anatoxin-a, is around ten-fold more potent than anatoxin-a, based on mouse bioassay tests [2]. Anatoxin-a(S) is a cholinesterase inhibitor that induces hypersalivation, diarrhea, shaking and nasal mucus discharge in mammals [24, 25]. It is thought to be produced only by Dolichospermum lemmermannii [26] and Dolichospermum flos-aquae [25]. Because it is rarely detected, very little research has been undertaken on this toxin and it is not discussed further in this chapter.

5.1.6 Lipopolysaccharides These toxins are integral components of the cell wall of all gram-negative bacteria, including cyanobacteria. The LPS are found in the outer cell membrane and form complexes with proteins and phospholipids [2]. LPS can elicit irritant and allergenic responses in humans and animals tissue [27, 28]. Although comparatively poorly studied, LPS have been implicated in human health problems associated with exposure to cyanobacteria [8]. Because of the limited information on these toxins, they are not discussed further in this chapter. Recent studies suggest that cyanobacteria will thrive under conditions of global climate change (see Chapter 7 and [29–35]). Cyanobacteria are considered the most important and dominant phototrophs in polar freshwater and terrestrial ecosystems, and recent research has shown that toxin production is more widespread than initially thought [36–39]. Climate change therefore has the potential to have a pronounced and dramatic effect on cyanobacterial composition and toxin production. In this chapter, we investigate how climate change may impact biomass. We also consider how climate change could impact cyanotoxin production through three key mechanisms: change in the concentration of toxin-producing species (including the expansion of species into new environments, e.g., from tropical to temperate regions), shifts in the relative abundance of toxic and nontoxic genotypes of bloom-forming species and variations in the relative amount of toxin produced at a strain level. We review data from both planktonic and benthic species. We also consider the geographic spread, including cyanobacteria in temperate, tropical and polar regions.

172

Laura T. Kelly et al.

5.2 Effects of climate change on common toxin-producing species In this section, we review four of the most common freshwater pelagic bloomforming genera: Microcystis, Raphidiopsis, Dolichospermum (formally Anabaena) (Figure 5.3(a), (b) and (c)) and Planktothrix. Benthic cyanobacteria have also become increasingly problematic in some countries [40]. We therefore use Microcoleus and the closely related taxon Phormidium (Figure 5.3(d)) as a case study to explore how climate change might affect benthic blooms in river environments.

(a)

(c)

(b)

(d)

Figure 5.3: Common bloom-forming cyanobacteria species: (a) Dolichospermum sp., (b) Raphidiopsis raciborskii, (c) Microcystis sp. and (d) Microcoleus sp.

5.2.1 Microcystis Microcystis is a ubiquitous genus that has been identified on all continents except Antarctica [41]. Microcystis spp. are well known as the dominant bloom-forming species and they are commonly associated with the production of microcystins. Microcystis cells lack individual sheaths, but under natural conditions are usually organized in colonies that can consist of many thousands of cells surrounded by a common mucilage [42]. Microcystis can regulate its position in the water column through gas vesicle production [43, 44] and negatively buoyant carbohydrate stores [45, 46], but can also remain buoyant in stratified waters. Both toxic and nontoxic genotypes exist and blooms are usually comprised of both, and are only distinguishable via molecular detection of genes in the microcystin synthetase gene operon [47].

5 Potential effects of climate change on cyanobacterial toxin production

173

Increased water temperature and density stratification induced by climate change [48] are likely to favor proliferation of Microcystis sp. Optimal temperatures for growth and photosynthesis are greater than 25 °C [49–52]. Intense stratification generally favors buoyancy-regulating cyanobacteria, and may also result in more extensive Microcystis blooms [31]. Davis et al. provided environmental and experimental data to suggest that predicted climatic changes in concert with eutrophication may promote the growth of toxic, rather than nontoxic Microcystis [53]. They conducted surveys at six study sites in temperate northeast United States and found that Microcystis became the dominant phytoplankton species at all sites as water temperature approached its annual maximum. They also showed that elevated temperatures resulted in significantly increased growth rates of toxic Microcystis strains in 83% of experiments, whereas growth rates of nontoxic strains only increased in 33%. Further evidence of shifts to toxic strains was provided by Dziallas and Grossart who incubated toxic and nontoxic strains of M. aeruginosa both with and without (i.e., under axenic conditions) bacteria in the laboratory at three temperatures (20, 26 and 32 °C) [54]. They observed a shift toward toxic strains at the higher temperature. In contrast, Martins et al. found selection of nontoxic Microcystis over toxic genotypes under favorable environmental growth conditions (including increased temperature) in a temperate reservoir in Northern Portugal [55]. Increasing atmospheric CO2 concentrations will likely favor the growth of buoyant species such as Microcystis. While blooms are rarely thought to be carbon limited due to an ability to use both CO2 and bicarbonate uptake pathways, using chemostat experiments, Verspagen et al. found higher growth rates and biomass of Microcystis as CO2 concentrations were increased, whereas under conditions of low CO2 and saturated bicarbonate, the strain was unable to sustain high biomass [56]. Recent research suggests that rising atmospheric CO2 levels may also affect the composition of toxic and nontoxic genotypes. Van de Waal et al. conducted chemostat experiments using mixed cultures of toxic and nontoxic M. aeruginosa [57]. The toxic strain reduced dissolved CO2 to lower concentrations than the nontoxic strain and became dominant at low CO2 levels. Conversely, the nontoxic strain could grow at lower light levels, and became dominant at high CO2 levels, but only under low-light conditions.

5.2.2 Raphidiopsis Raphidiopsis raciborskii is a solitary, filamentous diazotrophic freshwater cyanobacterium that was originally isolated from the lakes of Java (Indonesia) [58]. It poses a significant threat to ecological and human health because it can form dense blooms and produces cylindrospermopsins and saxitoxins [59]. It was initially believed to be confined to tropical environments, but over the past few decades, it appears to have expanded its range to many temperate regions across the globe [60–64].

174

Laura T. Kelly et al.

Several researchers have suggested that R. raciborskii will become increasingly dominant due to global warming [59, 60, 65]. Growth rates of R. raciborskii respond strongly to temperature between 20 °C and 35 °C, with maximum rates around 30 °C [66–68]. In tropical environments, R. raciborskii can bloom year-round, but it only dominates in temperate systems during periods of elevated temperature (i.e., summer months) [61, 62]. In temperate environments, it can form spore-like cells (akinetes) that allow it to overwinter and subsequently germinate under favorable temperatures (22 to 24 °C) [69]. Willis et al. found that water temperature had a greater influence on the growth of R. raciborskii than CO2 concentrations [70], however, increases in atmospheric CO2 concentrations are also predicted to have a positive effect on the growth of R. raciborskii [71]. Both toxic and nontoxic R. raciborskii strains exist, and among toxic genotypes, the amount of toxin produced per cell can vary markedly [66]. The characterization of the cylindrospermopsin biosynthesis pathway [72] and subsequent development of specific quantitative polymerase chain reaction assays [73] have enabled researchers to demonstrate that toxic and nontoxic genotypes cooccur in field populations [74]. To our knowledge, there are no published studies investigating how variables related to climate change will impact the relative abundance of toxinproducing versus nontoxin-producing genotypes, either in laboratory experiments or in the field. Although there is speculation that R. raciborskii blooms will expand and intensify with climate change, the paucity of knowledge on responses of specific toxic genotypes makes inferences regarding the potential toxicity of blooms uncertain. The niches of some R. raciborskii overlap with many Microcystis strains [75], further reinforcing the complexity of predicting cyanobacterial responses.

5.2.3 Dolichospermum Dolichospermum (formally known as Anabaena) is an ubiquitous filamentous diazotrophic genus found worldwide. Species within this genus have been reported to produce an array of cyanotoxins including; microcystins, anatoxin-a, anatoxin-a (S), cylindrospermopsin and saxitoxin [2]. The most commonly reported bloomforming species are D. circinalis, D. flos-aquae and D. lemmermannii. Dolichospermum most commonly blooms in waterbodies where there is strong temperature stratification. Both water temperature and stratification strength are predicted to increase with rising air temperature from climate change [65]. Dolichospermum may be selectively advantaged over other species in stratified systems: it can regulate buoyancy to allow for access to nutrients at depth or regulate light dose [76, 77]. It can also fix atmospheric nitrogen using specialized cells (heterocysts) under conditions of low availability of dissolved inorganic nitrogen [78, 79]. Using molecular techniques, researchers have shown that field populations of Dolichospermum comprise both toxic and nontoxic strains [80]. However, to our

5 Potential effects of climate change on cyanobacterial toxin production

175

knowledge there are no published studies which explore how the relative abundance of toxic and nontoxic genotypes will shift temporally.

5.2.4 Planktothrix Planktothrix is filamentous genus that is usually planktonic, forms solitary trichomes and has gas vacuoles. Two species, Planktothrix agardhii (green colored due to presence of phycocyanins) and P. rubescens (red colored due to phycoerythrins), are common in the northern hemisphere where they form blooms. Some strains produce microcystins [81]. While P. agardhii blooms in eutrophic lakes with low-light availability [82], P. rubescens usually blooms in deep oligotrophic lakes, in the low-light conditions of the metalimnion [83], often forming a “deep chlorophyll maximum.” Dokulil and Teubner analyzed a long-term dataset from Lake Mondsee (Austria) to explore the possible impacts of climate change on P. rubescens [84]. From their analysis, they suggest that P. rubescens will only benefit from climate warming early in the year, during late spring overturn and early summer. They also suggest that longer periods of summer stratification are unlikely to favor increasing biomass, although other studies provide contradictory results [85]. The potential impacts of climate change on P. agardhii are still unclear. Bonilla et al. assembled a large, global dataset for this species [82], and showed that although it was only observed in temperate and subtropical lakes, it could also bloom at a range of temperatures from 14 °C. Using an in-stream habitat assessment method, Heath et al. showed that decreases in summer minimum low flows had negligible change to the available Microcoleus habitat [101]. The authors concluded that the frequency of flushing flows (a flow that removes the Microcoleus from the river substrate), and not flow per se, was critical in determining the presence of Microcoleus. Climate change is predicted to cause shifts in precipitation through (a) strengthening of existing precipitation patterns; that is, wet regions become wetter and dry regions become drier, and (b) changing storm tracks; which should move away from the equator and toward the poles as atmospheric circulation changes [102]. In France and New Zealand, rivers in drier areas are likely to have prolonged low-flow periods which would be conducive to prolonged build-up of Microcoleus and Phormidium mats. Field and laboratory studies have demonstrated that Microcoleus mats are commonly a mixture of toxic and nontoxic genotypes. Among toxic genotypes, anatoxin concentrations can vary up to 100-fold [103, 104]. Heath et al. undertook a one-year study at eight sites on two rivers and observed that anatoxin-a and homoanatoxin-a occurrence was restricted to where water temperature exceeded 13.4 °C [100], and suggested that the toxin-producing strains may “out-compete” nontoxic Microcoleus strains under these conditions. More recent research has shown no correlation between water temperature and toxin production [105] or the relative abundance of toxic strains [106, 107]. Further culture- and laboratory-based studies are required to explore how temperature and other variables indirectly associated with climate change (e.g., river flow, rainfall) may impact toxin concentrations within the mats.

5 Potential effects of climate change on cyanobacterial toxin production

177

5.3 Effects of climate change on toxin regulation Environmental variables regulating cyanotoxin production have been a topic of intense interest over recent decades, however, the research has been strongly biased towards microcystin production. Most studies have used laboratory cultures to demonstrate relatively small (threefold to fourfold) changes in toxin production in response to changes in a number of different variables (e.g., light, nutrients or temperature). More recently, researchers have used field studies to examine the factors regulating toxin regulation [107, 108]. Here we review research on toxin regulation related to variables that are predicted to shift with climate change: temperature, light, dissolved carbon dioxide concentration and cell density (as a result of increased severity of cyanobacterial blooms).

5.3.1 Microcystins Although the biosynthetic pathway for microcystins is now known [47, 110], there is still uncertainty surrounding the transcriptional and post-translational regulation of the microcystin synthase. Microcystins were initially thought to be feeding deterrents for predators, and studies showing increased microcystin content in cyanobacteria exposed to predators supported this theory [111, 112]. However, the genes responsible for microcystin synthesis predate the eukaryotic lineage [113]. Many early studies focused on the effect of temperature on microcystin production and showed that toxin concentrations were maximal around 20 to 25 °C and decreased at higher and lower temperatures [52, 114–120]. More recent research incorporating gene expression has corroborated these results. For example, Dziallas et al. showed that microcystin transcripts were significantly lower at 32 °C than at 20 °C and 26 °C in axenic Microcystis cultures [54]. There are strong indications that light may play a regulatory role in microcystin production. Kaebernick et al. demonstrated that light quality specifically affects microcystin synthase expression, which is initiated at certain threshold intensities [121]. However, transcriptional changes in genes coding for microcystin production changes due to light did not correlate with observed variations in toxin production. Phelan and Downing showed a strong correlation between microcystin concentration and growth rate under high-light conditions in Microcystis aeruginosa PCC7806 [122]. However, no change was observed at optimal or low-light conditions, and media composition had no significant effect on the relationship between toxins and survival at high-light conditions. The authors therefore suggested a possible role for microcystin in protection against photooxidation [122]. Similarly, Geada et al. assessed microcystin production under different temperature, light, CO2 and pH conditions, finding that low light (10 μmol photons m−2 s−1) and elevated CO2 (5% v/v) resulted in the lowest microcystin concentrations [123]. Recently, researchers have provided evidence to support an intracellular function of microcystin in acclimation of Microcystis to high light and oxidative stress [124]. Sevilla

178

Laura T. Kelly et al.

et al. used quantitative reverse transcriptase polymerase chain reaction to monitor changes in the levels of transcripts encoding the microcystin-D-synthetase gene (mcyD) in M. aeruginosa PCC7806 subjected to oxidative agents and under different light intensities [125]. The study identified a link between microcystin synthesis and photosynthesis, and indicated that oxidative stress decreased microcystin synthesis. Using a DNA microarray based on the genome of M. aeruginosa PCC7806, Straub et al. studied the dynamics of gene expression during the light/dark cycle [126]. They showed that the biosynthesis of secondary metabolites, including microcystins, occurred mostly during the light period, suggesting that microcystin synthesis is connected with photosystem II. Light intensity has also been shown to alter ratios of microcystin variants. A strain of P. agardhii which produced [Asp3] congeners of both microcystin-LR and microcystin-RR exhibited increased levels of microcystin-LR at elevated irradiance levels [127]. The implication of this finding is that microcystin-LR is an order of magnitude more toxic than microcystin-RR, so the resulting phenotype from increased light intensity was more toxic. Ame and Wunderlin used batch culture and a natural population of M. aeruginosa. They found that, at a moderate temperature (20 °C), microcystin content would switch from being predominantly microcystin-LR to a less toxic phenotype that was predominantly microcystin-RR. At a higher temperature (28 °C), this did not occur and the levels of microcystin-LR and microcystinRR production remained constant [119]. A similar observation was reported by Song et al. using Microcystis viridis, with higher proportions of microcystin-RR at lower temperatures (15–20 °C) [128]. There is very limited data on the impact of pH on toxin production. Van der Westhuizen and Eloff suggested that Microcystis cells were more toxic at high and low pH values, corresponding to slower rates of growth [129]. Jahnichen et al. studied the impact of inorganic carbon availability on microcystin production, and suggested that microcystins may be involved in enhancing the efficiency of the adaptation of the photosynthetic apparatus to fluctuating inorganic carbon concentrations [130]. The potential effects of varying carbon dioxide levels on microcystin production from Microcystis are not well known. Van de Waal et al. showed that an excess supply of both nitrogen and carbon yielded high cellular nitrogen:carbon (N:C) ratios, accompanied by high cellular contents of total microcystin and the nitrogen-rich variant microcystin-RR [131]. They found comparable patterns in Microcystis-dominated lakes, where the relative microcystin-RR content increased with the seston N:C ratio. Using field-based studies, Wood et al. showed that Microcystis sp. can “switch” microcystin production on and off [108]. They also observed and experimentally induced c. 20-fold changes in microcystin quotas within a 5-h period and measured up to a 400-fold change in microcystin-E-synthetase gene (mcyE) expression [108, 109]. Both of these studies and a recent laboratory study [132] draw a correlation between microcystin production and cyanobacteria cell densities. These results

5 Potential effects of climate change on cyanobacterial toxin production

179

suggest that toxin production per cell may increase if blooms and scum formations intensify, as is predicted under future climate change scenarios [65].

5.3.2 Nodularins Little work has been performed on the environmental parameters that affect nodularin production. Lehtimaki et al. demonstrated that nodularin production was generally highest under conditions that promoted growth [133]. Intracellular nodularin concentrations increased at higher temperature (25–28 °C), higher irradiance (45–155 µmol photons m−2 s−1) and higher phosphate (PO4-P) concentrations (200–5,500 μg L−1). Teikari et al. showed that toxin-producing Nodularia can metabolize methylphosphonates, thus having a competitive advantage over other strains and cyanobacteria, especially under conditions of inorganic phosphorus limitation [134].

5.3.3 Cylindrospermopsins A study exploring the effect of temperature (20–35 °C) on growth and cylindrospermopsin content of R. raciborskii grown in batch culture showed a strong negative correlation between toxin content and temperature. When grown at 35 °C, none of the study strains produced any detectable cylindrospermopsin, but when transferred back to lower temperatures, cylindrospermopsin production was restored [66]. Preuβel et al. highlighted how toxin production can vary among strains of the same species [135]. They investigated the influence of light and temperature on cylindrospermopsin production of two Aphanizomenon flos-aquae strains using semicontinuous cultures. A light gradient from 10 to 60 μmol m−2 s−1, in combination with temperatures of 16, 20, and 25 °C, was assessed. Cylindrospermopsin concentrations showed a significant decrease with increasing temperature in one strain, while there was no clear relationship with temperature in the other. They also noted that cylindrospermopsin production at different light intensities varied at the three temperatures. In both strains, cylindrospermopsin concentrations increased significantly at 20 °C for light intensities from 10 to 60 μmol m−2 s−1, whereas concentrations decreased significantly at 25 °C in the same light gradient. They also suggested that conditions which constitute physiological stress (i.e., low light and temperature) triggered active release of the toxin into the medium. Dyble et al. investigated the role of light intensity on growth and cylindrospermopsin production in R. raciborskii cultures [136]. Maximum growth rates occurred at 75 μmol m−2 s−1, whereas maximum intracellular and extracellular toxin concentrations occurred in cultures grown under the highest light intensity (140 μmol m−2 s−1). The authors speculated that the highest intracellular toxin concentrations in the field

180

Laura T. Kelly et al.

are likely to occur in actively growing R. raciborskii populations exposed to light intensities of 75–150 μmol m−2 s−1 for more than two weeks. More recent studies provide evidence that cylindrospermopsin production in R. raciborskii is constitutive and the rate of production is the same as the cell division rate regardless of environmental conditions, although the amount produced can differ significantly among strains [137–139]. Strain-specific responses to nutrient concentrations are known to occur [70], however, to our knowledge there is no data on the relative abundance of toxic and nontoxic strains in relation to environmental conditions. Few studies have investigated changes in cylindrospermopsin production in benthic cyanobacteria. Bormans et al. [140] explored cylindrospermopsin production in Oscillatoria sp. PCC6506 and demonstrated that the total cylindrospermopsin content was highest during the exponential growth phase at intermediate light levels (10 μmol m−2 s−1) and during the stationary growth phase under lower and higher light levels.

5.3.4 Saxitoxins Only a small number of studies have explored changes in saxitoxin synthesis in relation to potential climate change. In a study on R. raciborskii (strain T3), saxitoxin synthesis was up-regulated at a light intensity of 100 µmol photons m−2 s−1 and not at 50 and 150 µmol photons m−2 s−1 [141]. This study also showed that saxitoxin production exhibited a circadian rhythm under the three light intensities tested. A study on a different R. raciborskii strain (C10) [142] found contradictory results. In this study, the extracellular and intracellular saxitoxin content of this strain was analyzed at two different temperatures; 19 and 25 °C. At 25 °C, an additional variant (dcSTX) was detected, but there was no significant change in toxin concentration. Comparing these data to other studies illustrates how responses to environmental parameters may differ among species. For example, Dias et al. studied Aphanizomenon sp. (LMECYA 31) and showed enhanced saxitoxin production at a high temperature (28 °C) compared to a lower temperature which was typical of in situ conditions (22 °C) [143].

5.3.5 Anatoxins Rapala et al. studied Dolichospermum and Aphanizomenon, and demonstrated that high temperature decreased the amount of the toxin produced regardless of growth rates [144]. Growth-limiting, low-light conditions and growth-inhibiting, high-light conditions decreased the amount of anatoxin in Dolichospermum cells. In contrast, the highest light flux in their study did not limit the growth or decrease the level of the toxin in the cells of Aphanizomenon. Rapala and Sivonen monitored anatoxin

5 Potential effects of climate change on cyanobacterial toxin production

181

production in Dolichospermum strains and observed that anatoxin-a production was slightly higher under suboptimal temperature and light levels [114]. Also, under low light, considerable amounts of extracellular anatoxin-a were detected. Araoz et al. studied Ocillatoria sp. (PCC6506), a benthic anatoxin-producing cyanobacteria, and noted shifts in the relative production of variants at different temperatures and when different carbon sources were utilized [145]. They observed that synthesis of anatoxin-a occurred at 22 °C when there was no carbon dioxide enrichment of the growth medium, while homanatoxin-a was produced when the medium was at 25 °C and supplemented with 10 mM sodium hydrogen carbonate to maintain carbon dioxide enrichment.

5.4 Climate change and its effect on cyanobacteria and toxin production in polar environments In Antarctica, cyanobacteria are considered the most important and dominant phototrophs in freshwater and terrestrial ecosystems [146]. Although less well studied, cyanobacteria are also abundant in the Arctic [147]. In both regions (where there is available water), they commonly form benthic or floating microbial mat communities which can be several centimeters thick (Figure 5.4). Nostocales and Oscillatoriales are usually dominant within these mats, and these taxa have adapted to tolerate harsh conditions, including high salinity, intense ultraviolet radiation and repeated cycles of desiccation/hydration [148, 149].

(a)

(b)

(c)

Figure 5.4: Antarctica cyanobacterial mats: (a) extensive mats in a meltwater pond in Miers Valley, (b) small pinnacles dominated by Leptolyngbya sp. and (c) a large colony of Nostoc sp.

182

Laura T. Kelly et al.

One of the fastest rates of climate warming has been documented in the Arctic and maritime Antarctic [150] with mean annual temperatures over the Antarctic Peninsula having increased by 0.5 °C per decade [151]. Freshwater habitats may be more severely impacted than terrestrial habitats as the relative increase in water temperature has been reported as two- to three-fold higher than the corresponding change of local air temperature [152]. A rise in temperature in ice-free areas is expected to increase the availability of liquid water (from glacial and permafrost melting) [153]. Such changes could dramatically increase water availability, possibly facilitating increased cyanobacterial growth. Wood et al. found high abundance of “dormant” cyanobacteria in Dry Valley soils [154]. As water availability increases, extensive macroscopic growth could occur within a relatively short time frame. Experiments using cloches (small covers that provided protection from wind and increased temperature and water availability) positioned on arid Dry Valley soils provided evidence to support the suggestion that increased water availability will favor cyanobacteria, and showed rapid growth of moss and cyanobacterial biomass (personal communication with D. D. Wynn-Williams as cited in [153]). Recent reports suggest that there is widespread microcystin production in Antarctic microbial mats including from the McMurdo Ice Shelf, Bratina Island [155, 156], McMurdo Dry Valleys [36] and the Antarctic Peninsula [39]. Although the microcystin producing species have not been definitively identified, Wood et al. used molecular techniques to provide compelling evidence that Nostoc sp. produced microcystins in at least some of the microcystin-positive samples [36]. Despite the high abundance of cyanobacteria in the Arctic [157], there are few reports of microcystins in cyanobacterial mats [37, 38, 158, 159], Arctic lakes [160] and one report in cyanobacteria from lichens in this region [161]. Kleinteich et al. demonstrated that microcystin production and species diversity increased in laboratory-cultured Arctic and Antarctic cyanobacterial mats when they were exposed to elevated temperatures (8–16 °C compared to 4 °C) [37]. It was suggested that this could be the result of increased toxin production per cell due to more favorable conditions. These conditions might include higher temperatures favoring the growth of toxin producing species and allowing them to become dominant during the experiment, or higher species diversity/abundance which could intensify interspecies competition at higher temperatures. In the case of the latter hypothesis, increased toxin concentrations could serve as a means for quorum-sensing, intraspecies communication or as allelopathic compounds. In addition to the potential ecosystem effects of increased microcystin production, Kleinteich et al. warned that changes in cyanobacterial diversity could lead to instability of the mat ecosystems, allowing the establishment of nonpolar mesophilic species [37]. Although less well studied, several other cyanotoxins have been detected in the polar regions. Kleinteich et al. confirmed the presence of cylindrospermopsin and deoxycylindrospermopsin in several cyanobacterial mats collected around Rothera (Antarctic Peninsula), although the concentrations measured were low compared

5 Potential effects of climate change on cyanobacterial toxin production

183

with benthic species of warmer climatic zones [39]. Saxitoxin was detected for the first time in a cyanobacterial mat collected from northern Baffin Island in the vicinity of Cape Hatt in the Arctic [38]. To date, the identity of the toxin producer(s) has not been determined. Anatoxins have been identified in biocrust samples in the Arctic on a single occasion [158]. Limited detections of this toxin may be the result of a lack of systematic surveys in polar regions, as several known producers are widespread in these environments (e.g., Microcoleus autumnalis). Although no data exist on how climate change will impact the abundance and diversity of toxins produced by cyanobacteria in polar regions rates of metabolism, nitrogen fixation and photosynthesis of Antarctic cyanobacteria are thought to be optimal around 15 °C [162]. Kleinteich et al. theorized that production of secondary metabolites may also be optimal at this temperature [37]. It is therefore plausible that climate change could yield increases in the concentration of these toxins, with as-yet unknown impacts on these fragile ecosystems.

5.5 Conclusions There is increasing evidence that regional and global climatic change will benefit many species of harmful bloom-forming cyanobacteria (see Chapter 7 or [29–33, 163, 164]). It is predicted that there will be increases in cyanobacteria growth rates, leading to greater longevity and severity of bloom events, and shifts in geographic distributions. Research indicates that many of the key bloom forming genera, that is, Microcystis, Raphidiopsis and Dolichospermum, will benefit under predicted climatic change scenarios such as: rising temperatures, enhanced water column vertical stratification and shifts in seasonal weather patterns (e.g., droughts, floods). It is extremely challenging to make generalizations about how climate change will impact cyanotoxin production. The studies to date have been undertaken using different conditions (e.g., growth media, batch vs. continuous culture), different methods for determining toxin concentrations (e.g., enzyme-linked immunosorbent assay, liquid chromatography-mass spectrometry) and have utilized different strategies for normalizing data to biomass (e.g., cell concentration, dry weight, optical density). Comparable studies have also shown contrasting responses between species (e.g., microcystin-producing Microcystis and Planktothrix), and even among strains of the same species (e.g., saxitoxin-producing R. raciborskii). We have only reviewed studies which have investigated variables closely associated with climate change. Many other variables may be indirectly related to climate change (i.e., nutrients, metal availability, abundance of parasitic pathogens, e.g., viruses and fungi [165]) and have also been shown to impact toxin production. Changes in these variables may have an equal or greater impact on toxin production, or act synergistically with the variables discussed. Studies that have undertaken

184

Laura T. Kelly et al.

multistressor experiments, for example investigating the effect on toxin production of temperature at a range of light intensities, have shown complex interactions and relationships, reinforcing how difficult it will be to predict the impacts of climate change in the natural environment. Finally, it should be noted that the majority of studies have been undertaken in the laboratory. While these have provided essential and foundational knowledge on toxin synthesis, and enabled detailed studies of single stressors, important natural variables and interactions that may regulate toxin production are removed. There is an urgent need for further field-based investigations that can better mimic the complex interactions and multiple stressors that might influence toxin production in cyanobacteria.

References [1]

[2] [3] [4]

[5]

[6]

[7]

[8]

[9]

[10]

[11]

Whitton BA, Potts M. Introduction to the Cyanobacteria, in The Ecology of Cyanobacteria, their Diversity in Time and Space. In: Whitton BA, Potts M [Editors.]. Dordrecht.; Kluwer Academic Publishers: 2000. 1–11. Chorus I, Bartram J, (eds). Toxic Cyanobacteria in Water: A Guide to their Public Health Consequences, Monitoring and Management. London; E & FN Spon: 1999. Meriluoto J, Spoof L, Codd GA. Handbook of Cyanobacterial Monitoring and Cyanotoxin Analysis. John Wiley & Sons: 2017. McGregor GB, Sendall BC. Iningainema pulvinus gen nov., sp nov. (Cyanobacteria, Scytonemataceae) a new nodularin producer from Edgbaston Reserve, north-eastern Australia. Harmful Algae 2017;62:10–19. Gehringer MM, Adler L, Roberts AA, Moffitt MC, Mihali TK, Mills TJT, Fieker C, Neilan BA. Nodularin, a cyanobacterial toxin, is synthesized in planta by symbiotic Nostoc sp. ISME J 2012;6(10):1834–1847. MacKintosh C, Beattie KA, Klumpp S, Cohen P, Codd GA. Cyanobacterial microcystin-LR is a potent and specific inhibitor of protein phosphatases 1 and 2A from both mammals and higher plants. FEBS Lett 1990;264(2):187–192. Honkanen RE, Dukelow M, Zwiller J, Moore RE, Khatra BS, Boynton AL. Cyanobacterial nodularin is a potent inhibitor of type 1 and type 2A protein phosphatases. Mol Pharmacol 1991;40(4):577–583. Ressom R, Soong FS, Fitzgerald DJ, Turczynowicz L, Saadi OE, Roder D, Maynard T, Falconer IR. Health Effects of Toxic Cyanobacteria (blue-green algae). Canberra; National Health and Medical Research Council: 1994. 108. Azevedo SMFO, Carmichael WW, Jochimsen EM, Rinehart KL, Lau S, Shaw GR, Eaglesham GK. Human intoxication by microcystins during renal dialysis treatment in Caruaru-Brazil. Toxicology 2002;181–182:441–446. Banker R, Teltsch B, Sukenik A, Carmeli S. 7-Epicylindrospermopsin, a toxic minor metabolite of the cyanobacterium Aphanizomenon ovalisporum from Lake Kinneret, Israel. J Nat Prod 2000;63(3):387–389. Norris RL, Eaglesham GK, Pierens G, Shaw GR, Smith MJ, Chiswell RK, Seawright AA, Moore MR. Deoxycylindrospermopsin, an analog of cylindrospermopsin from Cylindrospermopsis raciborskii. Environ Toxicol 1999;14(1):163–165.

5 Potential effects of climate change on cyanobacterial toxin production

[12]

[13]

[14]

[15] [16] [17]

[18]

[19] [20] [21] [22]

[23]

[24] [25] [26]

[27]

[28] [29] [30]

185

Terao K, Ohmori S, Igarashi K, Ohtani I, Watanabe MF, Harada KI, Ito E, Watanabe M. Electron microscopic studies on experimental poisoning in mice induced by cylindrospermopsin isolated from blue-green alga Umezakia natans. Toxicon 1994;32(7):833–843. Falconer IR, Hardy SJ, Humpage AR, Froscio SM, Tozer GJ, Hawkins PR. Hepatic and renal toxicity of the blue-green alga (cyanobacterium) Cylindrospermopsis raciborskii in male Swiss albino mice. Environ Toxicol 1999;14(1):143–150. Falconer IR, Humpage AR. Preliminary evidence for in vivo tumour initiation by oral administration of extracts of the blue-green alga Cylindrospermopsis raciborskii containing the toxin cylindrospermopsin. Environ Toxicol 2001;16(2):192–195. Moreira C, Azevedo J, Antunes A, Vasconcelos V. Cylindrospermopsin: Occurrence, methods of detection and toxicology. J Appl Microbiol 2013;114(3):605–620. Byth S. Palm Island mystery disease. Med J Aust 1980;2(1):40, 42. Bourke ATC, Hawes RB, Neilson A, Stallman ND. An outbreak of hepato-enteritis (the Palm Island mystery disease) possibly caused by algal intoxication. Toxicon 1983;21, Supplement 3 (0):45–48. Hawkins PR, Runnegar MT, Jackson AR, Falconer IR. Severe hepatotoxicity caused by the tropical cyanobacterium (blue-green alga) Cylindrospermopsis raciborskii (Woloszynska) Seenaya and Subba Raju isolated from a domestic water supply reservoir. Appl Environ Microbiol 1985;50(5):1292–1295. Adelman WJ, Fohlmeister JF, Sasner JJ, Ikawa M. Sodium channels blocked by aphantoxin obtained from the blue-green alga, Aphanizomenon flos-aquae. Toxicon 1982;20(2):513–516. Negri AP, Jones GJ, Hindmarsh M. Sheep mortality associated with paralytic shellfish poisons from the cyanobacterium Anabaena circinalis. Toxicon 1995;33(10):1321–1329. Bowling L, Baker P. Major cyanobacterial bloom in the Barwon-Darling River, Australia, in 1991, and underlying limnological conditions. Mar Freshwater Res 1996;47(4):643–657. Carmichael WW, An J. Using an enzyme linked immunosorbent assay (ELISA) and a protein phosphatase inhibition assay (PPIA) for the detection of microcystins and nodularins. Nat Toxins 1999;7(6):377–385. Gugger M, Lenoir S, Berger C, Ledreux A, Druart J-C, Humbert J-F, Guette C, Bernard C. First report in a river in France of the benthic cyanobacterium Phormidium favosum producing anatoxin-a associated with dog neurotoxicosis. Toxicon 2005;45(7):919–928. Carmichael WW. Cyanobacteria secondary metabolites – the cyanotoxins. J Appl Microbiol 1992;72(6):445–459. Mahmood NA, Carmichael WW. Anatoxin-a(s), an anticholinesterase from the cyanobacterium Anabaena flos-aquae NRC-525-17. Toxicon 1987;25(11):1221–1227. Henriksen P, Carmichael WW, An J, Moestrup Ø. Detection of an anatoxin-a(s)-like anticholinesterase in natural blooms and cultures of cyanobacteria/blue-green algae from danish lakes and in the stomach contents of poisoned birds. Toxicon 1997;35(6):901–913. Pilotto L, Hobson P, Burch MD, Ranmuthugala G, Attewell R, Weightman W. Acute skin irritant effects of cyanobacteria (blue-green algae) in healthy volunteers. Aust N Z J Public Health 2004;28(3):220–224. Torokne A, Palovics A, Bankine M. Allergenic (sensitization, skin and eye irritation) effects of freshwater cyanobacteria – Experimental evidence. Environ Toxicol 2001;16(6):512–516. O’Neil JM, Davis TW, Burford MA, Gobler CJ. The rise of harmful cyanobacteria blooms: The potential roles of eutrophication and climate change. Harmful Algae 2012;14:313–334. Kosten S, Huszar VLM, Bécares E, Costa LS, van Donk E, Hansson L-A, Jeppesen E, Kruk C, Lacerot G, Mazzeo N, De Meester L, Moss B, Lürling M, Nõges T, Romo S, Scheffer M. Warmer climates boost cyanobacterial dominance in shallow lakes. Glob Chang Biol 2012;18(1): 118–126.

186

[31]

[32] [33]

[34] [35]

[36]

[37]

[38]

[39]

[40]

[41]

[42] [43] [44] [45] [46] [47]

[48]

Laura T. Kelly et al.

Carey CC, Ibelings BW, Hoffmann EP, Hamilton DP, Brookes JD. Eco-physiological adaptations that favour freshwater cyanobacteria in a changing climate. Water Res 2012;46(5): 1394–1407. Hofer U. Climate change boosts cyanobacteria. Nat Rev Microbiol 2018;16(3):122–123. Yan X, Xu X, Wang M, Wang G, Wu S, Li Z, Sun H, Shi A, Yang Y. Climate warming and cyanobacteria blooms: Looks at their relationships from a new perspective. Water Res 2017;125:449–457. Paul VJ. Global Warming and Cyanobacterial Harmful Algal Blooms, in Cyanobacterial Harmful Algal Blooms: State of the Science and Research Needs. Springer: 2008. 239–257. Paerl H. Nutrient and Other Environmental Controls of Harmful Cyanobacterial Blooms Along the Freshwater–Marine Continuum, in Cyanobacterial Harmful Algal Blooms: State of the Science And Research Needs. Springer: 2008. 217–237. Wood SA, Mountfort D, Selwood AI, Holland PT, Puddick J, Cary SC. Widespread distribution and identification of eight novel microcystins in Antarctic cyanobacterial mats. Appl Environ Microbiol 2008;74(23):7243–7251. Kleinteich J, Wood SA, Kupper FC, Camacho A, Quesada A, Frickey T, Dietrich DR. Temperature-related changes in polar cyanobacterial mat diversity and toxin production. Nat Clim Chang 2012;2(5):356–360. Kleinteich J, Wood SA, Puddick J, Schleheck D, Kupper FC, Dietrich D. Potent toxins in Arctic environments – Presence of saxitoxins and an unusual microcystin variant in Arctic freshwater ecosystems. Chem Biol Interact 2013. doi.org/10.1016/j.cbi.2013.04.011. Kleinteich J, Hildebrand F, Wood SA, Ciŕs S, Agha R, Quesada A, Pearce DA, Convey P, K̈pper FC, Dietrich DR. Diversity of toxin and non-toxin containing cyanobacterial mats of meltwater ponds on the Antarctic Peninsula: A pyrosequencing approach. Antarct Sci 2014;26(05): 521–532. Quiblier C, Wood SA, Echenique-Subiabre I, Heath M, Villeneuve A, Humbert JF. A review of current knowledge on toxic benthic freshwater cyanobacteria – Ecology, toxin production and risk management. Water Res 2013;47(15):5464–5479. Harke MJ, Steffen MM, Gobler CJ, Otten TG, Wilhelm SW, Wood SA, Paerl HW. A review of the global ecology, genomics, and biogeography of the toxic cyanobacterium, Microcystis spp. Harmful Algae 2016;54:4–20. Komárek J, Anagnostidis K. Cyanoprokaryota 1 Teil: Chroococcales. Ettl H, et al. [ed]. Gustav Fisher Verlag Jena: 1999. 548. Walsby AE. Gas Vesicles. Annu Rev Plant Physiol 1975;26(1):427–439. Walsby AE. Gas vesicles. Microbiol Rev 1994;58(1):94–144. Visser PM, Ibelings BW, Mur LR. Autunmal sedimentation of Microcystis spp. as result of an increase in carbohydrate ballast at reduced temperature. J Plankton Res 1995;17(5):919–933. Visser PM, Passarge J, Mur LR. Modelling vertical migration of the cyanobacterium Microcystis. Hydrobiologia 1997;349(1–3):99–109. Tillett D, Dittmann E, Erhard M, von Döhren H, Börner T, Neilan BA. Structural organization of microcystin biosynthesis in Microcystis aeruginosa PCC7806: an integrated peptidepolyketide synthetase system. Chem Biol 2000;7(10):753–764. Trolle D, Hamilton DP, Hipsey MR, Bolding K, Bruggeman J, Mooij WM, Janse JH, Nielsen A, Jeppesen E, Elliott JA, Makler-Pick V, Petzoldt T, Rinke K, Flindt MR, Arhonditsis GB, Gal G, Bjerring R, Tominaga K, Hoen JT, Downing AS, Marques DM, Fragoso CR, Søndergaard M, Hanson PC. A community-based framework for aquatic ecosystem models. Hydrobiologia 2012;683(1):25–34.

5 Potential effects of climate change on cyanobacterial toxin production

187

[49] Robarts RD, Zohary T. Temperature effects on photosynthetic capacity, respiration, and growth rates of bloom-forming cyanobacteria. N. Z. J. Mar Freshwater Res 1987;21(3): 391–399. [50] Reynolds CS, Usher M, Saunders D, Dobson A, Peet R, Adam P, Birks H, Gustafsson L, McNeely J, Paine R. Ecology of Phytoplankton. Vol. 535. Cambridge University Press Cambridge: 2006. [51] Jöhnk KD, Huisman JEF, Sharples J, Sommeijer BEN, Visser PM, Stroom JM. Summer heatwaves promote blooms of harmful cyanobacteria. Glob Chang Biol 2008;14(3):495–512. [52] Bui T, Dao T-S, Vo T-G, Lürling M. Warming affects growth rates and microcystin production in tropical bloom-forming Microcystis strains. Toxins 2018;10(3):123. [53] Davis TW, Berry DL, Boyer GL, Gobler CJ. The effects of temperature and nutrients on the growth and dynamics of toxic and non-toxic strains of Microcystis during cyanobacteria blooms. Harmful Algae 2009;8(5):715–725. [54] Dziallas C, Grossart H-P. Increasing oxygen radicals and water temperature select for toxic Microcystis sp. PLoS ONE 2011;6(9):e25569. [55] Martins A, Moreira C, Vale M, Freitas M, Regueiras A, Antunes A, Vasconcelos V. Seasonal dynamics of Microcystis spp. and their toxigenicity as assessed by qPCR in a temperate reservoir. Mar Drugs 2011;9(10):1715–1730. [56] Verspagen JMH, Van de Waal DB, Finke JF, Visser PM, Van Donk E, Huisman J. Rising CO2 levels will intensify phytoplankton blooms in eutrophic and hypertrophic lakes. PLOS ONE 2014;9(8):e104325. [57] Van de Waal DB, Verspagen JMH, Finke JF, Vournazou V, Immers AK, Kardinaal WEA, Tonk L, Becker S, Van Donk E, Visser PM, Huisman J. Reversal in competitive dominance of a toxic versus non-toxic cyanobacterium in response to rising CO2. ISME J 2011;5(9):1438–1450. [58] Komárek JÍ, Kling H. Variation in six planktonic cyanophyte genera in Lake Victoria (East Africa). Arch Hydrobiol Suppl Algol Stud 1991;61:21–45. [59] Sinha R, Pearson LA, Davis TW, Burford MA, Orr PT, Neilan BA. Increased incidence of Cylindrospermopsis raciborskii in temperate zones – Is climate change responsible?. Water Res 2012;46(5):1408–1419. [60] Padisák J. Cylindrospermopsis raciborskii (Woloszynska) Seenayya et Subba Raju, an expanding, highly adaptive cyanobacterium: worldwide distribution and review of its ecology. Archiv fu¨r Hydrobiologie Supplementband: Monographische Beitra¨ge 1997;107(4): 563–593. [61] Hamilton PB, Ley LM, Dean S, Pick FR. The occurrence of the cyanobacterium Cylindrospermopsis raciborskii in Constance Lake: An exotic cyanoprokaryote new to Canada. Phycologia 2005;44(1):17–25. [62] Wood SA, Pochon X, Luttringer-Plu L, Vant BN, Hamilton DP. Recent invader or indicator of environmental change? A phylogenetic and ecological study of Cylindrospermopsis raciborskii in New Zealand. Harmful Algae 2014;39(0):64–74. [63] Rzymski P, Brygider A, Kokociński M. On the occurrence and toxicity of Cylindrospermopsis raciborskii in Poland. Limnological Review 2017;17(1):23–29. [64] Antunes JT, Leão PN, Vasconcelos VM. Cylindrospermopsis raciborskii: review of the distribution, phylogeography, and ecophysiology of a global invasive species. Front Microbiol 2015;6:473. [65] Paerl HW, Huisman J. Blooms like it hot. Science 2008;320(5872):57–58. [66] Saker ML, Griffiths DJ. The effect of temperature on growth and cylindrospermopsin content of seven isolates of Cylindrospermopsis raciborskii (Nostocales, Cyanophyceae) from water bodies in northern Australia. Phycologia 2000;39(4):349–354.

188

Laura T. Kelly et al.

[67] Briand J-F, Leboulanger C, Humbert J-F, Bernard C, Dufour P. Cylindrospermopsis raciborskii (cyanobacteria) invasion at mid-latitudes: selection, wide physiological tolerance, or global warming. J Phycol 2004;40(2):231–238. [68] Chonudomkul D, Yongmanitchai W, Theeragool G, Kawachi M, Kasai F, Kaya K, Watanabe MM. Morphology, genetic diversity, temperature tolerance and toxicity of Cylindrospermopsis raciborskii (Nostocales, Cyanobacteria) strains from Thailand and Japan. FEMS Microbiol Ecol 2004;48(3):345–355. [69] Padisák J, Reynolds CS. Selection of phytoplankton associations in Lake Balaton, Hungary, in response to eutrophication and restoration measures, with special reference to the cyanoprokaryotes. Hydrobiologia 1998;384(1–3):41–53. [70] Willis A, Chuang AW, Dyhrman S, Burford MA. Differential expression of phosphorus acquisition genes in response to phosphorus stress in two Raphidiopsis raciborskii strains. Harmful Algae 2019;82:19–25. [71] Holland DP, Pantorno A, Orr PT, Stojkovic S, Beardall J. The impacts of a high CO2 environment on a bicarbonate user: the cyanobacterium Cylindrospermopsis raciborskii. Water Res 2012;46(5):1430–1437. [72] Mihali TK, Kellmann R, Muenchhoff J, Barrow KD, Neilan BA. Characterization of the gene cluster responsible for cylindrospermopsin biosynthesis. Appl Environ Microbiol 2008;74(3): 716–722. [73] Rasmussen JP, Giglio S, Monis PT, Campbell RJ, Saint CP. Development and field testing of a real-time PCR assay for cylindrospermopsin-producing cyanobacteria. J Appl Microbiol 2008;104(5):1503–1515. [74] Orr PT, Rasmussen JP, Burford MA, Eaglesham GK, Lennox SM. Evaluation of quantitative real-time PCR to characterise spatial and temporal variations in cyanobacteria, Cylindrospermopsis raciborskii (Woloszynska) Seenaya et Subba Raju and cylindrospermopsin concentrations in three subtropical Australian reservoirs. Harmful Algae 2010;9(3):243–254. [75] Xiao M, Hamilton DP, O’Brien KR, Adams MP, Willis A, Burford MA. Are laboratory growth rate experiments relevant to explaining bloom-forming cyanobacteria distributions at global scale?. Harmful Algae 2020;92:101732. [76] Ganf GG, Horne AJ. Diurnal stratification, photosynthesis and nitrogen fixation in a shallow, equatorial Lake (Lake George, Uganda). Freshw Biol 1975;5(1):13–39. [77] Brookes JD, Ganf GG, Green D, Whittington J. The influence of light and nutrients on buoyancy, filament aggregation and flotation of Anabaena circinalis. J Plankton Res 1999;21 (2):327–341. [78] Oliver RL. Floating and sinking in gas-vacuolate cyanobacteria. J Phycol 1994;30(2):161–173. [79] Wood SA, Prentice MJ, Smith K, Hamilton DP. Low dissolved inorganic nitrogen and increased heterocyte frequency: precursors to Anabaena planktonica blooms in a temperate, eutrophic reservoir. J Plankton Res 2010;32(9):1315–1325. [80] Vaitomaa J, Rantala A, Halinen K, Rouhiainen L, Tallberg P, Mokelke L, Sivonen K. Quantitative real-time PCR for determination of microcystin synthetase E copy numbers for Microcystis and Anabaena in lakes. Appl Environ Microbiol 2003;69(12):7289–7297. [81] Ostermaier V, Kurmayer R. Application of real-time PCR to estimate toxin production by the cyanobacterium Planktothrix sp. Appl Environ Microbiol 2010;76(11):3495–3502. [82] Bonilla S, Aubriot L, Soares MCS, González-Piana M, Fabre A, Huszar VLM, Lürling M, Antoniades D, Padisák J, Kruk C. What drives the distribution of the bloom-forming cyanobacteria Planktothrix agardhii and Cylindrospermopsis raciborskii?. FEMS Microbiol Ecol 2012;79(3):594–607.

5 Potential effects of climate change on cyanobacterial toxin production

189

[83] Carraro E, Guyennon N, Hamilton D, Valsecchi L, Manfredi EC, Viviano G, Salerno F, Tartari G, Copetti D. Coupling high-resolution measurements to a three-dimensional lake model to assess the spatial and temporal dynamics of the cyanobacterium Planktothrix rubescens in a medium-sized lake. Hydrobiologia 2012;698(1):77–95. [84] Dokulil M, Teubner K. Deep living Planktothrix rubescens modulated by environmental constraints and climate forcing. Hydrobiologia 2012;698(1):29–46. [85] Posch T, Köster O, Salcher MM, Pernthaler J. Harmful filamentous cyanobacteria favoured by reduced water turnover with lake warming. Nat Clim Chang 2012;2(11):809–813. [86] Toporowska M, Pawlik-Skowronska B, Krupa D, Kornijow R. Winter versus summer blooming of phytoplankton in a shallow lake: effect of hypertrophic conditions. Pol J Ecol 2010;58(1): 3–12. [87] Crossetti L, de M. Bicudo C. Adaptations in phytoplankton life strategies to imposed change in a shallow urban tropical eutrophic reservoir, Garças Reservoir, over 8 years. Hydrobiologia 2008;614(1):91–105. [88] Gemelgo MCP, Mucci JLN, Navas-Pereira D. Population dynamics: seasonal variation of phytoplankton functional groups in Brazilian reservoirs (Billings and Guarapiranga, São Paulo). Braz J Biol 2009;69:1001–1013. [89] Walls JT, Wyatt KH, Doll JC, Rubenstein EM, Rober AR. Hot and toxic: Temperature regulates microcystin release from cyanobacteria. Sci Total Environ 2018;610:786–795. [90] Reynolds CS. The Ecology of Freshwater Phytoplankton. Cambridge; Cambridge University Press: 1984. [91] Oberhaus L, Briand JF, Leboulanger C, Jacquet S, Humbert J-F. Comparative effects of the quality and quantity of light and temperature on the growth of Planktothrix agardhii and P. rubescens. J Phycol 2007;43(6):1191–1199. [92] Briand E, Yéprémian C, Humbert J-F, Quiblier C. Competition between microcystin- and non-microcystin-producing Planktothrix agardhii (cyanobacteria) strains under different environmental conditions. Environ Microbiol 2008;10(12):3337–3348. [93] Briand E, Gugger M, François J-C, Bernard C, Humbert J-F, Quiblier C. Temporal variations in the dynamics of potentially microcystin-producing strains in a bloom-forming Planktothrix agardhii (Cyanobacterium) population. Appl Environ Microbiol 2008;74(12):3839–3848. [94] Komárek J, Anagnostidis K. Cyanoprokaryota 2 Teil: Oscillatoriales, in Susswasserflora von Mitteleuropa. Budel BB, et al. [ed.]. Gustav Fisher Verlag Jena: 2005. 750. [95] Cadel-Six S, Peyraud-Thomas C, Brient L, De Marsac NT, Rippka R, Méjean A. Different genotypes of anatoxin-producing cyanobacteria coexist in the Tarn River, France. Appl Environ Microbiol 2007;73(23):7605–7614. [96] Wood SA, Selwood AI, Rueckert A, Holland PT, Milne JR, Smith KF, Smits B, Watts LF, Cary CS. First report of homoanatoxin-a and associated dog neurotoxicosis in New Zealand. Toxicon 2007;50(2):292–301. [97] Heath MW, Wood SA, Ryan KG. Polyphasic assessment of fresh-water benthic mat-forming cyanobacteria isolated from New Zealand. FEMS Microbiol Ecol 2010;73(1):95–109. [98] Bouma-Gregson K, Kudela RM, Power ME. Widespread anatoxin-a detection in benthic cyanobacterial mats throughout a river network. PloS one 2018;13(5). [99] McAllister TG, Wood SA, Hawes I. The rise of toxic benthic Phormidium proliferations: A review of their taxonomy, distribution, toxin content and factors regulating prevalence and increased severity. Harmful Algae 2016;55:282–294. [100] Heath MW, Wood SA, Ryan KG. Spatial and temporal variability in Phormidium mats and associated anatoxin-a and homoanatoxin-a in two New Zealand rivers. Aquat Microb Ecol 2011;64(1):69–79.

190

Laura T. Kelly et al.

[101] Heath MW, Wood SA, Brasell KA, Young RG, Ryan KG. Development of habitat suitability criteria and in-stream habitat assessment for the benthic cyanobacteria Phormidium. River Res Appl 2013. Doi: 10.1002/rra.2722. [102] Trenberth KE. Changes in precipitation with climate change. Clim Res 2011;47(1–2):123–138. [103] Wood SA, Smith FMJ, Heath MW, Palfroy T, Gaw S, Young RG, Ryan KG. Within-mat variability in anatoxin-a and homoanatoxin-a production among benthic Phormidium (Cyanobacteria) strains. Toxins 2012;4(10):900–912. [104] Wood SA, Heath MW, Kuhajek J, Ryan KG. Fine-scale spatial variability in anatoxin-a and homoanatoxin-a concentrations in benthic cyanobacterial mats: implication for monitoring and management. J Appl Microbiol 2010;109(6):2011–2018. [105] Wood SA, Young RG. Cawthron Report No. 2217: Review of Benthic Cyanobacteria Monitoring Programme 2012. Nelson; Cawthron Institute: 2012. 42. [106] Kelly LT, Wood SA, McAllister TG, Ryan K. Development and application of a quantitative PCR assay to assess genotype dynamics and anatoxin content in Microcoleus autumnalisdominated mats. Toxins 2018;10(11):431. [107] Wood SA, Atalah J, Wagenhoff A, Brown L, Doehring K, Young RG, Hawes I. Effect of river flow, temperature, and water chemistry on proliferations of the benthic anatoxin-producing cyanobacterium Phormidium. Freshwater Sci 2017;36(1):63–76. [108] Wood SA, Rueckert A, Hamilton DP, Cary SC, Dietrich DR. Switching toxin production on and off: Intermittent microcystin synthesis in a Microcystis bloom. Environ Microbiol Rep 2011;3 (1):118–124. [109] Wood SA, Dietrich DR, Cary SC, Hamilton DP. Increasing Microcystis cell density enhances microcystin synthesis: A mesocosm study. Inland Waters 2012;2:17–22. [110] Nishizawa T, Ueda A, Asayama M, Fujii K, Harada K-I, Ochi K, Shirai M. Polyketide synthase gene coupled to the peptide synthetase module involved in the biosynthesis of the cyclic heptapeptide microcystin. J Biochem 2000;127(5):779–789. [111] Jang M-H, Ha K, Joo G-J, Takamura N. Toxin production of cyanobacteria is increased by exposure to zooplankton. Freshw Biol 2003;48(9):1540–1550. [112] Jang M-H, Ha K, Lucas MC, Joo G-J, Takamura N. Changes in microcystin production by Microcystis aeruginosa exposed to phytoplanktivorous and omnivorous fish. Aquat Toxicol 2004;68(1):51–59. [113] Rantala A, Fewer DP, Hisbergues M, Rouhiainen L, Vaitomaa J, Börner T, Sivonen K Phylogenetic evidence for the early evolution of microcystin synthesis. Proceedings of the National Academy of Sciences of the United States of America, 2004. 101(2): p. 568–573. [114] Rapala J, Sivonen K. Assessment of environmental conditions that favor hepatotoxic and neurotoxic Anabaena spp. strains cultured under light limitation at different temperatures. Microb Ecol 1998;36(2):181–192. [115] Codd GA, Poon GK. Cyanobacterial Toxins, in Proceedings of the Phytochemical Society of Europe. Gallon JG, Rogers LJ, [ed.]. Oxford; Oxford University Press: 1988. 283–296. [116] van der Westhuizen AJ, Eloff JN. Effect of temperature and light on the toxicity and growth of the blue-green alga Microcystis aeruginosa (UV-006). Planta 1985;163(1):55–59. [117] van der Westhuizen AJ, Eloff JN, Krüger GHJ. Effect of temperature and light (fluence rate) on the composition of the toxin of the cyanobacterium Microcystis aeruginosa (UV-006). Archiv fuer Hydrobiologie 1986;108:145–154. [118] Watanabe MF, Oishi S. Effects of environmental factors on toxicity of a cyanobacterium (Microcystis aeruginosa) under culture conditions. Appl Environ Microbiol 1985;49(5): 1342–1344.

5 Potential effects of climate change on cyanobacterial toxin production

191

[119] Amé M, Wunderlin D. Effects of iron, ammonium and temperature on microcystin content by a natural concentrated Microcystis aeruginosa population. Water Air Soil Pollut 2005;168(1): 235–248. [120] Sivonen K. Effects of light, temperature, nitrate, orthophosphate, and bacteria on growth of and hepatotoxin production by Oscillatoria agardhii strains. Appl Environ Microbiol 1990;56(9):2658–2666. [121] Kaebernick M, Neilan BA, Börner T, Dittmann E. Light and the transcriptional response of the microcystin biosynthesis gene cluster. Appl Environ Microbiol 2000;66(8):3387–3392. [122] Phelan RR, Downing TG. A growth advantage for microcystin production by Microcystis PCC7806 under high light. J Phycol 2011;47(6):1241–1246. [123] Geada P, Pereira RN, Vasconcelos V, Vicente AA, Fernandes BD. Assessment of synergistic interactions between environmental factors on Microcystis aeruginosa growth and microcystin production. Algal Res 2017;27:235–243. [124] Zilliges Y, Kehr J-C, Meissner S, Ishida K, Mikkat S, Hagemann M, Kaplan A, Börner T, Dittmann E. The cyanobacterial hepatotoxin microcystin binds to proteins and increases the fitness of Microcystis under oxidative stress conditions. PLoS ONE 2011;6(3):e17615. [125] Sevilla E, Martin-Luna B, Bes MT, Fillat M, Peleato ML. An active photosynthetic electron transfer chain required for mcyD transcription and microcystin synthesis in Microcystis aeruginosa PCC7806. Ecotoxicology 2012;21(3):811–819. [126] Straub C, Quillardet P, Vergalli J, de Marsac NT, Humbert J-F. A day in the life of Microcystis aeruginosa strain PCC 7806 as revealed by a transcriptomic analysis. PLoS ONE 2011;6(1): e16208. [127] Tonk L, Visser PM, Christiansen G, Dittmann E, Snelder EOFM, Wiedner C, Mur LR, Huisman J. The microcystin composition of the cyanobacterium Planktothrix agardhii changes toward a more toxic variant with increasing light intensity. Appl Environ Microbiol 2005;71(9): 5177–5181. [128] Song L, Sano T, Li R, Watanabe MM, Liu Y, Kaya K. Microcystin production of Microcystis viridis (cyanobacteria) under different culture conditions. Phycol Res 1998;46:19–23. [129] van der Westhuizen AJ, Eloff JN. Effect of culture age and pH of culture medium on the growth and toxicity of the blue green alga Microcystis aeruginosa. Zeitung Planzenphysiology 1983;110:157–163. [130] Jähnichen S, Ihle T, Petzoldt T, Benndorf J. Impact of inorganic carbon availability on microcystin production by Microcystis aeruginosa PCC 7806. Appl Environ Microbiol 2007;73(21):6994–7002. [131] van de Waal DB, Verspagen JMH, Lürling M, van Donk E, Visser PM, Huisman J. The ecological stoichiometry of toxins produced by harmful cyanobacteria: an experimental test of the carbon-nutrient balance hypothesis. Ecol Lett 2009;12(12):1326–1335. [132] Pereira DA, Giani A. Cell density-dependent oligopeptide production in cyanobacterial strains. FEMS Microbiol Ecol 2014;88(1):175–183. [133] Lehtimaki J, Moisander P, Sivonen K, Kononen K. Growth, nitrogen fixation and nodularin production by two Baltic Sea cyanobacteria. Appl Environ Microbiol 1997;63(5):1647–1656. [134] Teikari JE, Fewer DP, Shrestha R, Hou S, Leikoski N, Mäkelä M, Simojoki A, Hess WR, Sivonen K. Strains of the toxic and bloom-forming Nodularia spumigena (cyanobacteria) can degrade methylphosphonate and release methane. ISME J 2018;12(6):1619–1630. [135] Preußel K, Wessel G, Fastner J, Chorus I. Response of cylindrospermopsin production and release in Aphanizomenon flos-aquae (Cyanobacteria) to varying light and temperature conditions. Harmful Algae 2009;8(5):645–650.

192

Laura T. Kelly et al.

[136] Dyble J, Tester PA, Litaker RW. Effects of light intensity on cylindrospermopsin production in the cyanobacterial HAB species Cylindrospermopsis raciborskii. Afr J Mar Sci 2006;28(2):309–312. [137] Davis TW, Orr PT, Boyer GL, Burford MA. Investigating the production and release of cylindrospermopsin and deoxy-cylindrospermopsin by Cylindrospermopsis raciborskii over a natural growth cycle. Harmful Algae 2014;31:18–25. [138] Pierangelini M, Sinha R, Willis A, Burford MA, Orr PT, Beardall J, Neilan BA. Constitutive cylindrospermopsin pool size in Cylindrospermopsis raciborskii under different light and CO2 partial pressure conditions. Appl Environ Microbiol 2015;81(9):3069–3076. [139] Willis A, Adams MP, Chuang AW, Orr PT, O’Brien KR, Burford MA. Constitutive toxin production under various nitrogen and phosphorus regimes of three ecotypes of Cylindrospermopsis raciborskii ((Wołoszyńska) Seenayya et Subba Raju). Harmful Algae 2015;47:27–34. [140] Bormans M, Lengronne M, Brient L, Duval C. Cylindrospermopsin accumulation and release by the benthic cyanobacterium Oscillatoria sp. PCC 6506 under different light conditions and growth phases. Bull Environ Contam Toxicol 2014;92(2):243–247. [141] Carneiro RL, Dos Santos MEV, Pacheco ABF, Azevedo SMFDOE. Effects of light intensity and light quality on growth and circadian rhythm of saxitoxins production in Cylindrospermopsis raciborskii (Cyanobacteria). J Plankton Res 2009. [142] Castro D, Vera D, Lagos N, Garcı́a C, Vásquez M. The effect of temperature on growth and production of paralytic shellfish poisoning toxins by the cyanobacterium Cylindrospermopsis raciborskii C10. Toxicon 2004;44(5):483–489. [143] Dias E, Pereira P, Franca S. Production of paralytic shellfish toxin Aphanizomenon sp. LMECYA 31 (Cyanobacteria). J Phycol 2002;38(4):705–712. [144] Rapala J, Sivonen K, Luukkainen R, Niemelä S. Anatoxin-a concentration in Anabaena and Aphanizomenon under different environmental conditions and comparison of growth by toxic and non-toxic Anabaena-strains – A laboratory study. J Appl Phycol 1993;5(6):581–591. [145] Aráoz R, Nghiêm H-O, Rippka R, Palibroda N, de Marsac NT, Herdman M. Neurotoxins in axenic oscillatorian cyanobacteria: coexistence of anatoxin-a and homoanatoxin-a determined by ligand-binding assay and GC/MS. Microbiology 2005;151(4):1263–1273. [146] Taton A, Grubisic S, Brambilla E, De Wit R, Wilmotte A. Cyanobacterial diversity in natural and artificial microbial mats of Lake Fryxell (McMurdo Dry Valleys, Antarctica): A morphological and molecular approach. Appl Environ Microbiol 2003;69(9):5157–5169. [147] Jungblut AD, Lovejoy C, Vincent WF. Global distribution of cyanobacterial ecotypes in the cold biosphere. ISME J 2009;4(2):191–202. [148] Vincent WF. Microbial Ecosystems of Antarctica. Cambridge; Cambridge University Press: 1988. [149] Cavacini P. Soil algae from northern Victoria Land (Antarctica). Polar Biosci 2001;14:45–60. [150] Vaughan DG, Marshall GJ, Connolley WM, Parkinson C, Mulvaney R, Hodgson DA, King JC, Pudsey CJ, Turner J. Recent rapid regional climate warming on the Antarctic Peninsula. Clim Change 2003;60(3):243–274. [151] Turner J, Colwell SR, Marshall GJ, Lachlan-Cope TA, Carleton AM, Jones PD, Lagun V, Reid PA, Iagovkina S. Antarctic climate change during the last 50 years. Int J Climatol 2005;25(3): 279–294. [152] Quayle WC, Peck LS, Peat H, Ellis-Evans JC, Harrigan PR. Extreme responses to climate change in Antarctic lakes. Science 2002;295(5555):645. [153] Cowan DA, Ah Tow LA. Endangered Antarctica environments. Annu Rev Microbiol 2004;58(1): 649–690.

5 Potential effects of climate change on cyanobacterial toxin production

193

[154] Wood SA, Rueckert A, Cowan DA, Cary SC. Sources of edaphic cyanobacterial diversity in the Dry Valleys of Eastern Antarctica. ISME J 2008;2(3):308–320. [155] Hitzfeld BC, Lampert CS, Spaeth N, Mountfort D, Kaspar H, Dietrich DR. Toxin production in cyanobacterial mats from ponds on the McMurdo Ice Shelf, Antarctica. Toxicon 2000;38(12): 1731–1748. [156] Jungblut A-D, Hoeger SJ, Mountfort D, Hitzfeld BC, Dietrich DR, Neilan BA. Characterization of microcystin production in an Antarctic cyanobacterial mat community. Toxicon 2006;47(3): 271–278. [157] Vincent WF. Cyanobacterial Dominance in the Polar regions, in The Ecology of Cyanobacteria: Their Diversity in Time and Space. Whitton BA, Potts M, [Editors.]. Kluwer Academic Publishers: 2000. 321–340. [158] Chrapusta E, Węgrzyn M, Zabaglo K, Kaminski A, Adamski M, Wietrzyk P, Bialczyk J. Microcystins and anatoxin-a in Arctic biocrust cyanobacterial communities. Toxicon 2015;101:35–40. [159] Kleinteich J, Puddick J, Wood SA, Hildebrand F, Laughinghouse IV HD, Pearce DA, Dietrich DR, Wilmotte A. Toxic cyanobacteria in Svalbard: Chemical diversity of microcystins detected using a liquid chromatography mass spectrometry precursor ion screening method. Toxins 2018;10(4):147. [160] Trout-Haney J, Wood Z, Cottingham K. Presence of the cyanotoxin microcystin in Arctic lakes of Southwestern Greenland. Toxins 2016;8(9):256. [161] Kaasalainen U, Fewer DP, Jokela J, Wahlsten M, Sivonen K, Rikkinen J Cyanobacteria produce a high variety of hepatotoxic peptides in lichen symbiosis. Proceedings of the National Academy of Sciences, 2012. 109(15): p. 5886. [162] Velázquez D, Rochera C, Camacho A, Quesada A. Temperature effects on carbon and nitrogen metabolism in some Maritime Antarctic freshwater phototrophic communities. Polar Biol 2011;34(7):1045–1055. [163] Paul VJ. Global Warming and Cyanobacterial Harmful Algal Blooms, in Cyanobacterial Harmful Algal Blooms: State of the Science and Research Needs. Hudnell HK, [Editor]. Springer New York: 2008. 239–257. [164] Paerl HW, Huisman J. Climate change: a catalyst for global expansion of harmful cyanobacterial blooms. Environ Microbiol Rep 2009;1(1):27–37. [165] Velázquez D, López-Bueno A, De Cárcer DA, De Los Ríos A, Alcamí A, Quesada A. Ecosystem function decays by fungal outbreaks in Antarctic microbial mats. Sci Rep 2016;6(1):1–7.

Gustaaf M. Hallegraeff

6 Harmful marine algal blooms and climate change: progress on a formidable predictive challenge 6.1 Introduction In a strict sense, harmful algal blooms are completely natural phenomena that have occurred throughout recorded history (e.g., Exodus, Captain Vancouver in 1793). Whereas in the past four decades, unexpected new algal bloom phenomena have often been attributed to eutrophication [1] or ship ballast water introduction [2]; increasingly novel algal bloom episodes are now circumstantially linked to climate change. It is unfortunate that so few long-term records exist of algal blooms at any single locality; ideally we need at least 30 consecutive years. Whether the apparent global increase in harmful algal blooms represents a real increase or not is therefore a question that we will probably not be able to answer conclusively for some time to come. There is no doubt that our growing interest in using coastal waters for aquaculture is leading to a greater awareness of toxic algal species. People responsible for deciding quotas for pollutant loadings of coastal waters, or for managing agriculture and deforestation, should be made aware that a probable outcome of allowing polluting chemicals to seep into the environment will be an increase in harmful algal blooms. In countries that pride themselves on having disease and pollution-free aquaculture, every effort should be made to quarantine sensitive aquaculture areas against the unintentional introduction of nonindigenous harmful algal species. Nor can any aquaculture industry afford not to monitor for an increasing number of harmful algal species in water and for an increasing number of algal toxins in seafood products, or to use increasingly sophisticated analytical techniques such as liquid chromatography mass spectrometry. Last but not least, global climate change is now adding a new level of uncertainty to many seafood safety monitoring programs. Climate on our planet has been constantly changing, over scales of both millions of years (glacial to interglacial periods) and short-term oscillations of tens of years (El Niño–Southern Oscillation (ENSO), North Atlantic Oscillation (NAO)). The Earth’s climate in the distant past has at times been subject to much higher ultraviolet-B levels Acknowledgment: Our understanding of climate-driven impacts on phytoplankton processes is continuously increasing. This chapter represents a partial update on an earlier review in Journal of Phycology 2010 and benefitted from my attendance at an IOC-UNESCO meeting on HABs and climate change in Gothenburg in 2015. Gustaaf M. Hallegraeff, Institute for Marine and Antarctic Studies, University of Tasmania, Tasmania, Australia https://doi.org/10.1515/9783110625738-006

196

Gustaaf M. Hallegraeff

and CO2 concentrations than we are seeing at present. The first photosynthetic cyanobacteria evolved 3.5 billion years ago at CO2 levels 1,000 times those of the present, followed by green algae 1,000 million years ago (mya) at 500 times the present value and dinoflagellates 330–400 mya at eight times the present value, whereas more recently evolved diatoms and haptophytes operated under comparatively low CO2 environments (two to three times the present value) [3]. During the past 800,000 years, atmospheric CO2 has fluctuated between 180 ppm in glacial and 300 ppm in interglacial periods, but in the past 200 years, this has increased from 280 ppm to 412 ppm at present, with values of 750–1,000 ppm predicted by 2100. In the past 1,000 years, our planet has gone through episodes warmer than at present, such as the medieval warm period AD 550–1300, and colder than now, such as the little ice age AD 1300–1900. Global temperatures in the past 20–30 years have increased significantly with a further rise of 2–4 (6) °C predicted over the next 100 years [4]. From a geological perspective, there is nothing remarkable about the magnitude of climate change we are experiencing now, except that it appears to proceed at a faster pace and starts from a warmer baseline. The current rate of changes in carbonate chemistry referred to as “ocean acidification” (at least 10–100 times faster than the recent glacial transitions) however is unprecedented within the last 65 million years. Because of their short generation times and longevity, many phytoplankton are expected to respond to current climate change with only a very small time lag. They are expected to spread quickly with moving water masses into climatic conditions that match the temperature, salinity, land runoff and turbulence requirements of the species. However, our knowledge of marine microalgae’s potential to adapt is very limited. Collins and Bell [5] grew the freshwater green alga Chlamydomonas over 1,000 generations at almost three times the present atmospheric CO2 concentration. The cells acclimated to the change but did not show any genetic mutations that could be described as adaptation. Lohbeck et al. [6] exposed Emiliania huxleyi coccolithophorid cultures founded by single or multiple clones to increased concentrations of CO2 in 500 generation selection experiments. Compared with populations kept at ambient CO2 partial pressure, those selected at increased partial pressure exhibited higher growth rates suggestive of adaptive evolution. Organism’s limits to adaptive capacity exist but remain largely unexplored, but we also should not underestimate that microbial life in the oceans had some 3.5 billion years to evolve, thus representing an enormous genetic diversity and physiological plasticity [7]. While we can expect changes in distribution, performance and genetic diversity of individual species, complete extinction is unlikely.

6 Harmful marine algal blooms and climate change

197

6.2 Algal bloom range extensions and climate change Temperature defines the geographic distribution of many species and their responses to climate change. Shifting temperature means and extremes alter habitats and cause changes in abundance through local extinctions and latitudinal expansions or shifts. Vulnerability is thought to be greatest in polar organisms due to their narrow temperature ranges and in tropical species living close to upper thermal limits. The dinoflagellate Pyrodinium bahamense is presently confined to tropical, mangrove-fringed coastal waters of the Atlantic and Indo-West Pacific. A survey of cyst fossils (named Polysphaeridium zoharyii) going back to the warmer Eocene 50 mya indicates a much wider range of distribution in the past. For example, in the Australasian region at present, the alga is not found farther south than Papua New Guinea but, some 100,000 years ago in the Pleistocene, the alga ranged as far south as Sydney Harbour [8]. There is concern that, with increased greenhouse warming of the oceans, this toxin-producing species may one day return to Australian waters (Figure 6.1). Pyrodinium blooms in 1972 in Papua New Guinea coincided with the fatal food poisoning of three children in a seaside village, diagnosed as paralytic shellfish poisoning (PSP). Since then, toxic blooms have spread to Brunei and Sabah (1976), the central (1983) and northern Philippines (1987) and Indonesia (north Mollucas). There is strong circumstantial evidence of a coincidence between Pyrodinium blooms and unusual weather linked to the ENSO. In the Philippines alone, Pyrodinium has

IPWP

Recent plankton

Fossil

Recent cysts

NSW, 100.000 years ago

Figure 6.1: Global distribution of Pyrodinium bahamense in recent plankton (top left) and much wider distribution in the fossil cyst record (bottom left) (after [2]). Top right: the fossil dinocyst Dapsilidinium pastielsii, which became extinct in the Atlantic during cooling in the early Pleistocene but survived in the warm-water refuge of the Indo-Pacific Warm Pool (IPWP) (after [11]). Bottom right: Pyrodinium bahamense cyst from Port Moresby Harbor, Papua New Guinea.

198

Gustaaf M. Hallegraeff

now been responsible for more than 2,000 human illnesses and 100 deaths resulting from the consumption of contaminated shellfish as well as sardines and anchovies [9, 10]. Comparable examples of spreading of dinoflagellate cyst species distributions with increasing temperatures, and shrinkage of biogeographical zones with decreasing temperatures, are known. The fossil dinocyst Dapsilidinium pastielsii became extinct in the Atlantic during cooling in the early Pleistocene, but a warm-water refuge for this taxon was recently discovered in the Indo-Pacific Warm Pool (IPWP: Japan, Indonesia, Vietnam, Palau and Philippines) [11]. Ciguatera caused by the benthic dinoflagellate species complex Gambierdiscus toxicus is a tropical fish-food poisoning syndrome well-known in coral reef areas in the Caribbean, Australia and especially French Polynesia (Figure 6.2). Whereas, in a strict sense, this is a completely natural phenomenon; from being a rare disease two centuries ago, ciguatera has now reached epidemic proportions in French Polynesia. From 1960 to 1984, more than 24,000 patients were reported from this area, which is more than six times the average for the Pacific as a whole. Evidence is accumulating that reef disturbance by hurricanes, military and tourist developments as well as coral bleaching (linked to global warming) and perhaps future increasing coral damage and changing macrophyte cover due to ocean acidification are increasing the risk of ciguatera [12]. In the Australian region, Gambierdiscus is

Ciguatera endemic

Imported toxic fish

Ciguatera expansion area

Figure 6.2: Global distribution of ciguatera fish-food poisoning, indicating ciguatera endemic areas, countries where cases have resulted from imported fish, and areas where ciguatera is considered to have expanded, either by expansion of the causative dinoflagellate or migration of toxic fish or both (from ciguatera-online@ciguateraonline). Inset: the causative benthic dinoflagellate genus Gambierdiscus.

6 Harmful marine algal blooms and climate change

199

well known from the tropical Great Barrier Reef and southwards down to just north of Brisbane, but in the past five years ciguatera outbreaks have increasingly been caused by fish caught in northern New South Wales, either due to range expansion of the causative dinoflagellate or migration of toxic fish or both [13]. A similar apparent range expanse of Gambierdiscus has been reported in the Mediterranean and the Canary Islands [14] and the Caribbean and West Indies [15]. In the North Sea, an analogous shift of warm-water phyto- and zooplankton to the North Pole has occurred due to regional climate warming [16–18].

6.3 Range extensions further aided by ship ballast water transport Ballast water is seawater which has been pumped into a ship’s hold or dedicated ballast tanks to steady it by making it heavier and thus less likely to roll; the water is released when a ship enters port. Ballast water on cargo vessels was first suggested as a means of dispersing marine plankton 100 years ago [19]. However, it was only in the 1980s that the problem sparked considerable interest, after evidence was brought forward that nonindigenous toxic species such as the Paralytic Shellfish Toxin (PST) producing dinoflagellate Gymnodinium catenatum had been introduced into Australian waters in sensitive aquaculture areas, with disastrous consequences for commercial shellfish farms [20]. To prove that a particular species of microalga has been introduced into a particular location is much more complex than for example in the case of macroalgae or marine invertebrates. However, the dogma of widespread cosmopolitanism of marine microalgae is now increasingly rejected in favor of underestimated microalgal diversity. The implication is that human-mediated translocations and their impacts may have been seriously underestimated. For example, we now recognize that the Alexandrium tamarense species complex is comprised of six to eight different genotypes, now widely designated as distinct species [21, 22], some of which are always or mostly nontoxic (Tasmanian and European clades), while others (North American and Temperate Asian clade) contain potent toxic strains. The molecular detection in the Mediterranean ports of Sete and Barcelona of Alexandrium pacificum with a temperate Asian ribotype not found anywhere else in Europe [23, 24]; thus most likely reflects a human-assisted introduction. Similarly, Australian populations of Alexandrium hide introduced temperate Asian ribotypes among indigenous nontoxic strains [25]. However, a recent outbreak of Paralytic Shellfish Poisoning (PSP) in Tasmania (SE Australia) in 2012, which cost the local economy $23 M through contamination of shellfish, abalone and lobster, is most likely to have been caused by a previously cryptic genotype newly stimulated by climate-driven increases in the frequency and persistence of coastal stratification during winter and early spring [26].

200

Gustaaf M. Hallegraeff

It is by no means clear whether the harmful dinoflagellate Cochlodinium polykrikoides, which caused major fish kills and problems for the desalination industry in the Arabian Gulf in 2008, is a ballast water introduction [27]. It could also represent a climate-driven range expansion from 2005 Malaysian blooms or be a species that has always been there in low concentration but now stimulated by changing environmental conditions such as anthropogenic nutrient enrichment (Figure 6.3). A concerted effort looking for Cochlodinium cysts or their genetic fingerprints in dated sediment depth cores may help resolve this question.

Cochlodinium polykrikoides

1995

1978

1999

2004

2000

2008 2004 ?

East Asian Phillipines American Malaysian Unknown

Figure 6.3: Geographical distribution of the fish-killing dinoflagellate Cochlodinium polykrikoides with years of first bloom occurrences noted. The 2008–2009 Gulf bloom was a new bloom phenomenon for the Gulf region, represented by the American/Malaysian ribotype, which is distinct from the East Asian and Philippines ribotypes (after [31]). We cannot yet resolve whether this Gulf bloom event represents a ballast water introduction, a climate-driven range expansion or a response of a previously cryptic species to anthropogenic nutrient enrichment or perhaps a combination of all of the above.

A comparable scenario applies to the red-tide dinoflagellate Noctiluca scintillans in Australian waters. While native to the Sydney region (since 1860), in the 1980s, eutrophication caused it to expand its distribution; subsequently, climate-driven range extension moved it southwards into the Tasmanian region, and only very recently (2008) have we seen it expand against prevailing current systems to Port Esperance and Cairns, hence implying an additional role of domestic ballast water dispersal [28] (Figure 6.4). Ecosystems disturbed by pollution or climate change are more prone to ballast-water invasions than mature stable ecosystems [29]. Similarly, green Noctiluca blooms, containing green flagellate endosymbionts, appear to have

6 Harmful marine algal blooms and climate change

201

increased in the Arabian Sea putatively due to its competitive advantage under low oxygen conditions [30].

1860– 1950

10 S 1980– 1993

1994– 2005

20 S 2008–13 30 S 40 S 50 S 60 S 100 E

120 E

140 E

160 E

160 E

Figure 6.4: Australia-wide distribution of the red-tide dinoflagellate Noctiluca scintillans (inset), in the periods 1860–1950 (only known from Sydney Harbour), 1980–1993 (expanding along New South Wales coast in response to eutrophication), 1994–2005 (East Australian current driven range extension) and 2008–2013 (expanding against prevailing current systems to Port Esperance and Cairns, implying a role for domestic ballast water dispersal (updated after [28]).

6.4 The formidable challenge of predicting phytoplankton community responses While we made considerable progress in our understanding of the physics of climate change, our understanding of the impacts on biological communities is in its infancy. There will be winners and losers from climate change, but predicting how individual species will respond poses a formidable challenge [32, 33]. Increasing temperature, enhanced surface stratification, alteration of ocean currents, intensification or weakening of local nutrient upwelling, stimulation of photosynthesis by elevated CO2, ocean acidification and increased frequency of heavy precipitation and storm events causing changes in land runoff and micronutrient availability may all produce contradictory species or even strain-specific responses. Complex factor interactions exist and ecophysiological experiments rarely take into account genetic strain diversity and physiological plasticity [34]. An undoubted key driver for future phytoplankton changes will be increasing sea surface temperature and enhanced water column stratification (shallowing of the mixed layer). In open ocean environments, this may lead to more rapid depletion of surface nutrients, a decrease in replenishment from deep nutrient rich waters and therefore reduced phytoplankton biomass (“oligotrophication” [35]). By contrast, in high-latitude regions with relatively deep mixing and sufficient nutrients, decreasing mixing depth can result in higher phytoplankton biomass because of increased light

202

Gustaaf M. Hallegraeff

availability [36]. Cyanobacteria can dominate both marine and freshwater ecosystems under higher temperature, notably when combined with eutrophication [37]. Extreme weather events such as heavy rainfall (nutrients from land runoff), hurricanes and dust storms have well known impacts to marine phytoplankton. Winds influence the supply of iron to the surface ocean through aeolian transport of dust from land to sea. This contributes micronutrients such as iron, which stimulate Karenia brevis blooms off Florida [38]. Hurricanes have been claimed as being responsible for expanding the distribution of cyst-producing toxic dinoflagellates (e.g., A. catenella in New England [39]). In Hiroshima Bay, blooms of the fish-killing raphidophyte Chattonella marina followed typhoon-induced accretion of nutrientrich land runoff [40]. It is widely predicted that increasing CO2 will lead to ocean acidification (“the other CO2 problem”), a decrease in ocean pH (from 8.1 down to 7.7 by 2100) and associated changes in carbonate chemistry. Most of the tested harmful algal bloom species lack carbon-concentrating mechanisms and hence they may benefit from increased atmospheric CO2 [3]. While initial attention focused on potentially adverse impacts of ocean acidification on calcifying organisms such as the coccolithophorid E. huxleyi [41], a much greater impact may derive from how ocean acidification will alter the availability of micronutrients such as iron [42]. However, nobody could have predicted that ocean acidification would induce more subtle changes such as the swimming behavior of the fish-killing raphidophyte Heterosigma akashiwo [43]. Dominance and community structure of harmful bloom dinoflagellates can be profoundly altered by changing pCO2 [44], and both toxic dinoflagellates (Alexandrium catenella and Karlodinium veneficum) and diatoms (Pseudo-nitzschia multiseries) have been shown to produce higher cellular toxin concentrations under near-future levels of ocean acidification [45–47].

6.5 We can learn from the fossil record, long-term plankton records and decadal scale climate events The ecosystem response to natural climate variability in the past provides a glimpse into the climate-induced changes of the near future. We can learn important lessons from the dinoflagellate cyst fossil record [48] and from the few long-term data sets available (such as the Continuous Plankton Recorder surveys [16]; Figures 6.5 and 6.6). Data from the Continuous Plankton Recorder in the Northeast Atlantic confirm that warming from 1960 to 1995 enhanced phytoplankton growth [49]. In response to transient warming, phytoplankton distribution in the North Atlantic shifted toward the pole by hundreds of kilometers per decade since the 1950s. Phenology of plankton in

6 Harmful marine algal blooms and climate change

Anomaly

Anomaly

–0.7

–0.3

Anomaly

0.1

0.5

203

Anomaly

0.9

Figure 6.5: Decadal anomaly maps (difference between long-term 1960–1989 mean and the 1990–2002 period) for four common HAB species (from left to right) Prorocentrum, Tripos furca, Dinophysis and Noctiluca (insets) in the North Atlantic. Note the increase in Prorocentrum, Tripos furca and Dinophysis along the Norwegian coast and increase in Noctiluca in the Southern North Sea (after [17]).

the North Atlantic was also affected, with differences in sensitivity between groups. Hinder et al. [50] attributed a recent decline in North Sea dinoflagellates relative to diatoms to warming, increased summer windiness and thus water column turbulence. Seasonal timing of phytoplankton blooms is now occurring up to four to five weeks earlier in the North Sea in relation to regional climate warming (Figure 6.6). Similarly, Moore et al. [51] predict longer-lasting A. catenella blooms in Puget Sound under future climate change scenarios. Coccolithophore blooms of E. huxleyi in the Bering Sea were reported for the first time during the period 1997–2000, probably in response to a 4 °C warming, combined with a shallower mixed layer depth, higher light levels and low zooplankton grazing [52]. A similar range expansion and increase abundance since the 1990s of E. huxleyi in the Southern Ocean has been reported [53], but the underlying mechanism remains to be fully explained. Extreme weather events such as the anomalous warm 2015–2016 period in the entire Pacific can mimic future climate conditions and provide a “dress rehearsal” for understanding future frequency, intensity and geographic extent of harmful algal blooms (HABs). However, the differing HAB responses by Pseudo-nitzschia toxic diatoms in the US Pacific Northwest, A. catenella dinoflagellate and Pseudochattonella fish-killing flagellate blooms in Chile and A. catenella PSP outbreaks in Tasmania, Australia, were not foreseen [54]. It is also important to recognize that not all trophic levels are responding to the same extent, and where zooplankton or fish grazers are differentially impacted by ocean warming, this may have cascading impacts on the structure of marine food webs [55].

204

Gustaaf M. Hallegraeff

1950

1960

Years (1948–2001) 1970 1980

1990

2000

2

Month

4 6 8 10 12 0.0 0.8 1.3 1.5 1.7 1.9 2.1 2.4 2.8 3.3 Phytoplankton color index

Figure 6.6: Long-term monthly values of “phytoplankton color” in the central North Sea from 1948 to 2001. Circles denote >2 SD above the long-term monthly mean (after [49], with permission). Note an apparent shift toward earlier spring and autumn phytoplankton blooms.

6.6 Mitigation of the likely impact on seafood safety The greatest problems for human society will be caused by being unprepared for significant range extensions of HAB species or an increase of algal biotoxin problems in areas that are poorly monitored at present. While, for example, ciguatera contamination would be expected and monitored for in tropical coral reef fish, with the apparent range extension of the causative benthic dinoflagellate into warm-temperate sea grass beds of South-Eastern Australia, other coastal fisheries unexpectedly could be at risk. Similarly, incidences of increased surface stratification in estuaries, heavy precipitation or extreme storm events are all warning signs that call for increased vigilance of monitoring seafood products for algal biotoxins even in areas not currently known to be at risk. Only with improved global ocean observation systems, such as improved and expanded ocean sensor capabilities (e.g., argo floats, ocean gliders, coastal moorings and coastal radar, multiwavelength and variable fluorometers and optical sensors) in support of integrated satellite-derived “ocean color” maps and expanded biological and biogeochemical observations (continuous plankton recorder and ecogenomics) can we expect to define management options, forecast ocean-related risks to human health and safety and shed light on the impact of climate variability on marine life and humans in general. It is pleasing to see that a number of national (e.g. the US NSTC Joint Subcommittee on Ocean Science and Technology Ocean Observatories Initiative and the Australian Integrated Marine Observing System) and international programs (e.g., the Intergovernmental Oceanographic Commission of UNESCO’s GlobalHAB and IOC Global HAB Status Reports initiative [56]) are actively pursuing these ambitious goals.

6 Harmful marine algal blooms and climate change

205

References [1]

[2] [3] [4]

[5] [6] [7]

[8] [9] [10] [11]

[12] [13]

[14] [15] [16] [17] [18]

Smayda TJ. Novel and nuisance phytoplankton blooms in the sea: evidence for a global epidemic. In: Graneli E, Sundstrom B, Edler L, Anderson DM. (eds.) Toxic marine phytoplankton. New York; Elsevier: 1990. 29–40. Hallegraeff GM. A review of harmful algal blooms and their apparent global increase. Phycologia 1993;32:79–99. Beardall J, Raven JA. The potential effects of global climate change on microalgal photosynthesis, growth and ecology. Phycologia 2004;43:26–41. IPCC, 2019. Special Report on the Ocean and Cryosphere in a Changing Climate [Pörtner HO, Roberts DC, Masson-Delmotte V, Zhai P, Tignor M, Poloczanska E, Mintenbeck K, Nicolai M, Okem A, Petzold J, Rama B, Weyer [eds.]]. 2019. Collins S, Bell G. Phenotypic consequences of 1000 generations of selection at elevated CO2 in a green alga. Nature 2004;431:566–569. Lohbeck KT, Riebesell U, Reusch TBH. Adaptive evolution of a key phytoplankton species to ocean acidification. Nat Geosci 2012;5:346–351. Read BA, Kegel J, Klute MJ, Kuo A, Lefebvre SC, Maumus F, Mayer C, Miller J, Monier A, Salamov A, Aguila M, Claverie JM, Frickenhaus S, Gonzalez K, Herman EK, Lin Y-C, Napier J, Ogata H, Sarno AF, Shmutz J, Schroeder D, deVargas C, Verret F, von Dassow P, Valentin K, Van de Peer Y, Wheeler G. Emiliania huxleyi Annotation Consortium, Dacks JB, Delwiche CF, Dyhrman ST, Glöckner G, John U, Richards T, Worden AZ, Young J, Zhang X, Grigoriev I. Emiliania’s pan genome drives the phytoplankton’s global distribution. Nature 2013;499: 209–213. Late MA. Pleistocene dinoflagellate cysts from Botany Bay, New South Wales, Australia. Micropaleontology 1989;35:1–9. Hallegraeff GM, Maclean JL Biology, epidemiology and management of Pyrodinium red tides. International Centre for Living Aquatic Resources Management. Manila; Conf Proc: 1989 21. 286. Azanza RV, Taylor FJR. Are Pyrodinium blooms in Southeast Asian recurring and spreading? A view at the end of the millennium. AMBIO 2001;30:356–364. Mertens KN, Takano Y, Head MJ, Matsuoka K. Living fossils in the Indo-Pacific warm pool: A refuge for thermophilic dinoflagellates during glaciations. Geology 2014. Doi:10.1130/ G35456.1 Skinner MP, Brewer TD, Johnstone R, Fleming LE, Lewis RJ. Ciguatera fish poisoning in the Pacific Islands (1998 to 2008). PLoS Negl Trop Dis 2011;5(12):e1416. Farrell H, Zammit A, Harwood DT, McNabb P, Shadbolt C, Manning J, Turahui JA, van den Berg AJ, Szabo L. Clinical diagnosis and chemical confirmation of ciguatera fish poisoning in New South Wales, Australia. Commun Dis Intell 2016;40:E1–E6. Aligizaki K, Nikolaidis G, Fraga S. Is Gambierdiscus expanding to new areas?. Harmful Algae 2008;36:6–7. Tester PA, Feldman RL, Nau AW, Kibler SR, Litaker RW. Ciguatera fish poisoning and sea surface temperatures in the Caribbean Sea and the West Indies. Toxicon 2010;56:698–710. Hays GC, Richardson AJ, Robinson C. Climate change and marine plankton. Trends Ecol Evol (Amst.) 2005;20(6):337–344. Edwards M, Richardson AJ. The impact of climate change on the phenology of the plankton community and trophic mismatch. Nature 2004;430:881–884. Richardson AJ, Schoeman DS. Climate Impact on Plankton Ecosystems in the Northeast Atlantic. Science 2004;305:1609–1612.

206

Gustaaf M. Hallegraeff

[19] Ostenfeld CH. On the immigration of Biddulphia sinensis Grev. and its occurrence in the North Sea during 1903–1907 and on its use for the study of the direction and rate of flow of the currents. Meddelelser fra Kommissionen for Danmarks Fiskeriog Havundersøgelser: Serie Plankton 1908; 1: 1–44. [20] McMinn A, Hallegraeff GM, Thomson P, Jenkinson AV, Heijnis H. Cyst and radionucleotide evidence for the recent introduction of the toxic dinoflagellate Gymnodinium catenatum into Tasmanian waters. Mar Ecol Progr Ser 1998;161:165–172. [21] Scholin C, Hallegraeff GM, Anderson DM. Molecular evolution of the Alexandrium tamarense “species complex.” (Dinophyceae) and their dispersal in the North American and West Pacific regions. Phycologia 1995;34:472–485. [22] John U, Litaker RW, Montresor M, Murray S, Brosnahan ML, Anderson DM. Formal revision of the Alexandrium tamarense species complex (Dinophyceae) taxonomy: the introduction of five species with emphasis on molecular-based (rDNA) classification. Protist 2014;165: 779–804. [23] Lilly EL, Kulis DM, Gentien P, Anderson DM. Paralytic shellfish poisoning toxins in France linked to a human-introduced strain of Alexandrium catenella from the western Pacific: Evidence from DNA and toxin analysis. J Plankton Res 2002;24:443–452. [24] Vila M, Garces E, Maso M, Camp J. Is the distribution of the toxic dinoflagellate Alexandrium catenella expanding along the NW Mediterranean coast?. Mar Ecol Progr Ser 2001;222: 73–83. [25] Bolch CJS, de Salas MF. A review of the molecular evidence for ballast water introduction of the toxic dinoflagellates Gymnodinium catenatum and the Alexandrium tamarensis complex to Australasia. Harmful Algae 2007;6:465–485. [26] Condie SA, Oliver ECJ, Hallegraeff GM. Environmental drivers of unprecedented Alexandrium catenella dinoflagellate blooms off eastern Tasmania, 2012–2018. Harmful Algae 2019:87. [27] Richlen ML, Morton SL, Jamali EA, Rajan A, Anderson DM. The catastrophic 2008–2009 red tide in the Arabian Gulf region, with observations on the identification and phylogeny of the fish-killing dinoflagellate Cochlodinium polykrikoides. Harmful Algae 2010;9:163–172. [28] McLeod DJ, Hallegraeff GM, Hosie GW, Richardson AJ. Climate-driven range expansion of the red-tide dinoflagellate Noctiluca scintillans into the Southern Ocean. J Plankton Res 2012;34: 332–337. [29] Stachowicz JJ, Terwin JR, Whitlatch RB, Osman RW. Linking climate change and biological invasions: ocean warming facilitates non-indigenous species invasion. Proc Nat Acad Sc USA 2002;99:15497–15500. [30] Gomes HDR, Goes JI, Matondkar SGP, Buskey EJ, Basu S, Parab S, Thoppil P. Massive outbreaks of Noctiluca scintillans blooms in the Arabian Sea due to spread of hypoxia. Nat Commun 2014. Doi: 10.1038/ncomms5862. [31] Iwataki M, Kawami H, Mizushima K, Mikulski CM, Doucette GJ, Relox JJR, Anton A, Fukuyo Y, Matsuoka K. Phylogenetic relationships in the harmful dinoflagellate Cochlodinium polykrikoides (Gymnodiniales, Dinophyceae) inferred from LSU rDNA sequences. Harmful Algae 2008;7:271–277. [32] Hallegraeff GM. Ocean climate change, phytoplankton community responses, and harmful algal blooms: a formidable predictive challenge. J Phycol 2010;46:220–235. [33] Wells ML, Trainer VL, Smayda TJ, Karlson BSO, Trick CG, Kudela RM, Ishikawa A, Bernard S, Wulff A, Anderson DM, Cochlan WP. Harmful algal blooms (HABs) and climate change What do we know and where do we go from here?. Harmful Algae 2015;49:68–93.

6 Harmful marine algal blooms and climate change

207

[34] Feng Y, Warner ME, Zhang Y, Sun J, Fu FX, Rose JM, Hutchins DA. Interactive effects of increased pCO2, temperature and irradiance on the marine coccolithophore Emiliania huxleyi (Prymnesiophyceae). European J Phycol 2008;43:87–98. [35] Behrenfeld MJ, O’Malley RT, Siegel DA, McClain CR, Sarmiento JL, Feldman GC, Milligan AJ, Falkowski PG, Letelier RM, Boss ES. Climate-driven trends in contemporary ocean productivity. Nature 2006;444:752–755. [36] Doney SC. Plankton in a warmer world. Nature 2006;444:695–696. [37] O’Neil, Davis TW, Burford MA, Gobler CJ. The rise of harmful cyanobacteria blooms: The potential roles of eutrophication and climate change. Harmful Algae 2012;14:313–334. [38] Walsh JJ, Steidinger KA. Saharan dust and Florida red tides: The cyanophyte connection. J Geophys Research 2001;106:11597–11612. [39] Anderson DM. Bloom dynamics of toxic Alexandrium species in the northeastern US. Limnol Oceanogr 1997;42:1009–1022. [40] Kimura T, Mizokami A, Hashimoto T. The red tide that caused severe damage to the fishery resources in Hiroshima Bay: Outline of its occurrence and environmental conditions. Bull Plankton Soc Japan 1973;19:82–96. [41] Riebesell U, Zondervan I, Rost B, Tortell PD, Zeebe RE, Morel FMM. Reduced calcification of marine plankton in response to increased atmospheric CO2. Nature 2000;407:364–367. [42] Shi D, Xu Y, Hopkinson BM, Morel FM. Effect of ocean acidification on iron availability to marine phytoplankton. Science 2010;327:676–679. [43] Kim H, Spivack AJ, Menden-Deuer S. pH alters the swimming behaviors of the raphidophyte Heterosigma akashiwo: Implications for bloom formation in an acidified ocean. Harmful Algae 2013;26:1–11. [44] Fu FX, Tatters AO, Hutchins DA. Global change and the future of harmful algal blooms in the ocean. Mar Ecol Prog Ser 2012;470:207–233. [45] Fu FX, Place AR, Garcia NS, Hutchins DA. CO2 and phosphate availability control the toxicity of the harmful bloom dinoflagellate Karlodinium veneficum. Aquat Microb Ecol 2010;59: 55–65. [46] Sun J, Hutchins DA, Feng Y, Seubert EL, Caron DA, Fu FX. Effects of changing pCO2 and phosphate availability on domoic acid production and physiology of the marine harmful bloom diatom Pseudo-nitzschia multiseries. Limnol Oceanogr 2011;56:829–840. [47] Tatters AO, Flewelling LJ, Fu F, Granholm AA, Hutchins DA. High CO2 promotes the production of paralytic shellfish poisoning toxins by Alexandrium catenella from Southern California waters. Harmful Algae 2013;30:37–43. [48] Dale B. The sedimentary record of dinoflagellate cysts: looking back into the future of phytoplankton blooms. Sc Mar 2001;65:257–272. [49] Edwards M. Phytoplankton blooms in the North Atlantic: results from the Continuous Plankton Recorder survey 2001/2002. Harmful Algae News 2004;25:1–3. [50] Hinder SL, Hays GC, Edwards M, Roberts EC, Walne AW, Gravenor MB. Changes in marine dinoflagellate and diatom abundance under climate change. Nat Clim Chang 2012;2:271–275. [51] Moore SK, Mantua JM, Hickey B, Trainer VL. Recent trends in paralytic shellfish toxins in Puget Sound, relationship to climate, and capacity for prediction of toxic events. Harmful Algae 2008;8:463–477. [52] Merico A, Tyrrell T, Brown CW, Groom SB, Miller PI. Analysis of satellite imagery for Emiliania huxleyi blooms in the Bering Sea before 1997. Geophys Res Lett 2003;30:1337. Doi:10.1029/ 2002GL016648 [53] Winter A, Henderiks J, Beaufort L, Rickaby REM, Brown CW. Poleward expansion of the coccolithophore Emiliania huxleyi. J Plankton Res 2014;36:316–325.

208

Gustaaf M. Hallegraeff

[54] Trainer VL, Moore SK, Hallegraeff G, Kudela RM, Clement A, Mardones JI, Cochlan WP. Pelagic harmful algal blooms and climate change: Lessons from nature’s experiments with extremes. Harmful Algae 2019. doi.org/10.1016/j.hal.2019.03.009. [55] Engelhard GH, Righton DA, Pinnegar JK. Climate change and fishing: a century of shifting distribution in North Sea cod. Glob Chang Biol 2014. Doi: 10.1111/ gcb.12513 [56] Hallegraeff G, Bresnan E, Enevoldsen H, Schweibold L, Zingone A. Call to contribute to global Harmful Algal Bloom status reporting. Harmful Algae News 2017;58:1–3.

Elke S. Reichwaldt, Som Cit Sinang and Anas Ghadouani

7 Global warming, climate patterns and toxic cyanobacteria 7.1 Introduction “The world is getting warmer.” Most people who read this will think of an increase in the average yearly temperature. However, there is more to it. Heat waves showed unprecedented scales in the last decades [1–7]. More importantly, it is also predicted that their intensity, frequency and duration will increase rapidly under future scenarios of climate change (e.g., [2, 8, 9]). Heat waves can be defined as a number of consecutive days above a certain threshold temperature, usually calculated for a specific location [9]. For example, Australia experienced an extensive heat period in 2013 [4] that led to the introduction of a new color on the charts of the Bureau of Meteorology (Figure 7.1). While the effect of heat waves on human health has been studied extensively (e.g., [10]) and governments are starting to develop approaches to coordinate an integrated response to heat waves to support risk groups (e.g., [11]), the direct and indirect effects of heat waves on lake processes have received very little attention. What do extreme temperature events do to water bodies? Most studies investigated the effect of environmental factors, including temperature, on cyanobacteria on a seasonal or interannual scale. However, there is a tremendous lack of information on the immediate and long-term effect of heat waves on cyanobacterial dynamics. Some first information collected during a heat wave in Europe showed that surface temperature, thermal stability and hypolimnetic oxygen depletion increased [12]; the same heat wave was responsible for an increase in cyanobacterial biomass through a combination of higher temperature, reduced wind speed and reduced cloud cover, with the bloom exploding during the heat wave as soon as an artificial mixing process was switched off [13]. Heat waves have also been shown to affect climate patterns, such rainfall patterns [7], which indirectly impact cyanobacterial dynamics. While there are some consensus on the fact that “blooms like it hot” [14], this idea may have led to the generation of a simplistic explanation based solely on an average

Elke S. Reichwaldt, Anas Ghadouani, Aquatic Ecology and Ecosystem Studies, School of Civil, Environmental and Mining Engineering, The University of Western Australia, Australia Som Cit Sinang, Aquatic Ecology and Ecosystem Studies, School of Civil, Environmental, and Mining Engineering, The University of Western Australia, Australia and Faculty of Science and Mathematics, Sultan Idris Education University, Perak, Malaysia https://doi.org/10.1515/9783110625738-007

210

Elke S. Reichwaldt, Som Cit Sinang and Anas Ghadouani

Screen Temperature Valid 06UTC Mon 14 Jan 2013

ACCESS-Global t+162

95°E 100°E 105°E 110°E 115°E 120°E 125°E 130°E 135°E 140°E 145°E 150°E 155°E 160°E 165°E 170°E 175°E

5°S 10°S 15°S 20°S 25°S 30°S 35°S 40°S 45°S 50°S

°C 54 50 46 42 38 34 30 26 22 18 14 10 6 2 –2 –6 –10 –14 –18 –25 –40 –60

95°E 100°E 105°E 110°E 115°E 120°E 125°E 130°E 135°E 140°E 145°E 150°E 155°E 160°E 165°E 170°E 175°E

Figure 7.1: Temperature map of Australia during a heat period in January 2013. Note the purple color toward the middle, which represents the newly introduced color code for temperatures between 50 and 52 °C (picture courtesy of the Australian Bureau of Meteorology).

increase in water temperature. However, there is evidence that extreme events can trigger regime shifts in ecosystems [15–17] and this might also be true for the development of (toxic) algal blooms. This chapter aims to provide a synthesis of the combined effects of direct and indirect consequences of global warming on lake processes and on the occurrence of cyanobacteria; it also presents an outlook of some future scenarios and their implications for society that urgently need further consideration.

7.2 The effect of global warming on inland water bodies 7.2.1 Direct effects of global warming on inland water bodies Global average surface air temperature is predicted to increase between 1.8 and 4.0 °C by the end of this century [18] and this will have severe consequences for inland water bodies [19]. A higher temperature will directly affect lake temperature and this will be especially pronounced in shallow lake systems because their water temperature usually closely follows air temperature. A higher average temperature or a longer period of warm temperature can affect cyanobacterial growth, bloom

7 Global warming, climate patterns and toxic cyanobacteria

211

development and toxin production directly [13]. Furthermore, any change in temperature-driven lake processes, such as stratification, inflow rates into water bodies, evaporation and nutrient cycling processes will result in changes in cyanobacterial growth (7.2), species composition and toxin production. Higher temperatures that prevail for a long period will cause stronger stratifications leading to fast depletion of oxygen in the hypolimnion [12, 20] followed by phosphorous enrichment in the deeper layer [12, 20]. The phosphorous can then be used as an important nutrient source by buoyant cyanobacterial genera that are able to move within the water column (e.g., Microcystis, Anabaena) [13]. Stronger stratification might require events of higher energy to break it down, suggesting that a higher volume of inflow is needed to mix the water column. In addition, higher temperature can lead to a deeper epilimnion and, in temperate lakes, to an earlier onset of the stratification [21], which can lead to extended growth seasons for phytoplankton. Higher temperature in the hypolimnion might also facilitate mineralization of organic matter; this adds to the nutrient pool [22, 23]. A higher temperature will also lead to higher evaporation rates. This can lead to decreased water levels, higher nutrient concentrations and decreased surface inflow into inland water bodies resulting in lower flushing rates and longer residence times, all conditions that favor cyanobacteria [24–26] (Figure 7.2).

7.2.2 Indirect effects of global warming on inland water bodies Higher atmospheric temperatures will increase the water vapor capacity in the atmosphere [27]. This will have severe impacts on climate patterns such as the distribution of rainfall on the Earth and the frequency and the intensity of rainfall [27]. These predicted changes in precipitation patterns will strongly influence water quality [28]. The symptoms of increased temperature and changes in rainfall patterns are often the same (Table 7.1), making it impossible to identify which of the two is responsible for an actual change in cyanobacterial bloom dynamics (Figure 7.2) (e.g., [29–32]). With global warming, it is expected that the total amount of rainfall will increase; however, due to a high interdecadal variability, the observed trend for the global annual land mean precipitation depends on the period it is calculated from and on the region from which the data was sourced ([27], see also [34]). On a global scale, it is predicted that the frequency of extreme rainfall events will change more dramatically than the mean precipitation rate [27, 35]; heavy rainfall events are predicted to occur more often in the near future while the amount of total precipitation is predicted to change only slowly. The increase in frequency is even likely for regions where a reduction or no change in the total amount of rainfall is predicted [36–40]. This will lead to prolonged dry periods between events [41], and the probability for droughts is predicted to increase, especially in mid-continental areas during summer [18]. The mean intensity of events has also increased worldwide, and this trend is especially pronounced for very heavy and extreme days of rain [39, 42, 43].

212

Elke S. Reichwaldt, Som Cit Sinang and Anas Ghadouani

Global warming Temperature → Water vapor capacity Surface Evaporation runoff

Climate patterns Temperature UV radiation level

Evaporation

Rainfall frequency Rainfall amount

Surface inflow

Stratification Water residence time

Temperature Light

Toxin production

Mixing

Turbidity Conductivity

Cyanobacterial biomass

Phytoplankton Grazers

pH

Subsurface inflow

Nutrients

Figure 7.2: A conceptual diagram showing the different pathways of how global warming can potentially affect cyanobacterial bloom dynamics. Dark gray arrows depict direct pathways, light gray arrows indicate the indirect pathway through changes in climate patterns and blue arrows indicate the hydrologic pathway.

Such changes in rainfall patterns will impact the physicochemical conditions within water bodies, which in turn directly affect the ecology of the lake including the occurrence of cyanobacteria (Figure 7.2). Environmental conditions that are affected by changes in climate patterns include shifts in the concentration of micro- and macronutrients, salinity, turbidity, pH and the stratification of the water body. During rainfall events, water enters the water body directly by wet deposition and indirectly by surface runoff and subsurface flows (Figure 7.2). The fraction of direct wet deposition of rainwater into water bodies becomes more important the lower the input through groundwater or surface water flow (i.e., small catchment area or distant coastal areas) [44, 45]. While the effect of wet deposition is immediate as rainwater mixes with lake water during the rainfall, the impact of water entering the water body indirectly via groundwater or surface water might take hours to years (e.g., [46]), depending on the size of the catchment area and the geology [47]. The effect of rainfall events on inland water bodies will also differ, depending on whether the lake receives its water primarily through groundwater exchange or through surface streams (e.g., runoff, rivers) [48], because of differences in physical and biological processes in

7 Global warming, climate patterns and toxic cyanobacteria

213

Table 7.1: Correlation between the effects of global warming, specifically an increase in the atmospheric temperature, a change in rainfall intensity and the length of dry periods on water body conditions or mechanisms that affect these conditions and the direct relationships between these conditions/mechanisms and nutrient enrichment of water bodies, which will then indirectly affect the occurrence of cyanobacterial blooms. Atmospheric temperature

Rainfall intensity

Length of dry period

Condition/mechanism

Nutrient enrichment

+

−

+

Water residence time

+

+

−

+

Anoxic conditions

+

+

0/ −

+

Water temperature



+

−/+*

+

Conductivity



−

+

−

Flushing

−

+

−

+

Water column stability

+/−



+

+

Nutrient concentrations in the inflow after rainfall

+



+

−

Turbidity through resuspension of sediment





−

+

Turbidity through concentration of biomass



The table should be read as follows: water residence time is positively correlated to atmospheric temperature, negatively correlated to rainfall intensity and positively correlated to the length of the dry period; water residence time in turn is positively correlated with nutrient enrichment. Explanation for the respective correlations can be found in the text. 0 indicates no correlation. * In areas with highly saline groundwater (e.g., South Australia) [33].

groundwater and surface water that modify, for instance, nutrient species during transport after rainfall events [49]. The effect of rainfall events on water bodies is highly complex and depends mainly on the interplay between the quantity and quality of the inflowing water, the volume ratios of inflowing to receiving water, the seasonal timing of the event and in-lake conditions [34, 50]. The main processes that affect the quantity and quality of the inflowing water in relation to rainfall patterns are the following: the quality of inflowing water depends not only on the geology and land use in the catchment area, and the length of the preceding dry period, but also on the intensity of the rainfall event and the chemistry of the rainwater. The quantity of the inflowing water depends on the amount of rainfall, the length of the preceding dry period, the size of the catchment area and the land use in the area, and is an important driver of water residence time. Although rainfall volume and inflow volume are correlated, this correlation might only be visible if enough rain falls to saturate the catchment [51].

214

Elke S. Reichwaldt, Som Cit Sinang and Anas Ghadouani

7.2.2.1 Effect of rainfall events on lake temperature, stratification and turbidity Many rainfall events, especially the more extreme ones, occur concurrently with lower air temperatures and strong winds. A significant decrease in air temperature for a prolonged period will especially influence shallow lake systems, because their temperature usually closely follows air temperature. Strong wind events will increase the depth of the epilimnion by mixing and consequently decrease mean temperature in this layer. A high inflow of water can destabilize the stratification and cause mixing [52]. Turbidity, which describes the amount of suspended solids in the water, can increase significantly and for a long period through sediment transported in surface inflows or resuspension of sediment by storm events [52, 53]. A decrease in the penetration of radiation into the water column has been found after high rainfall [54–56] and during rainy seasons [57]. Furthermore, prolonged dry periods and high temperatures can gradually increase turbidity due to high biomass accumulation. On the other hand, prolonged no-rain periods can lead to lower turbidity and thus higher irradiance penetration in water reservoirs [58] due to a decrease in suspended particles. Another factor that will change due to variation in rainfall patterns is the dissolved organic carbon (DOC) input into aquatic systems. However, the trend is not clear with studies showing either higher or lower amounts of DOC in the water column after longer dry periods (reviewed in [59]). As DOC, and especially colored DOC, significantly influences the light regime in the water column [60], a change in the concentration of DOC will strongly influence the competition between primary producers and thus affect the occurrence of algal blooms.

7.2.2.2 Effect of rainfall events on water level and water residence time The water level of water bodies fluctuates naturally due to an imbalance of the water budget, which depends on meteorological and hydrological processes such as evaporation rate, precipitation rate, inflow and outflow [61, 62]. In general, prolonged periods without precipitation lead to a decreased water level [63–65] and a recent study indicated that the water level is correlated with the total amount of rainfall per year and the frequency of extreme rainfall events [66]. While the first correlation has a delayed reaction time of weeks to months due to subsurface inflows, the latter is a more immediate effect possibly through an increased surface runoff during high-intensity events [53, 66]. Therefore, the predicted increase in the frequency of extreme rainfall events will lead to more variable water levels in the future. This effect might be especially pronounced in closed lakes that rely on precipitation as their inflow, and in reservoirs as the natural outflow is artificially regulated. Lower lake levels generally favor cyanobacteria due to a concentration of nutrients. Water residence time is strongly negatively correlated to the intensity of rainfall [67] and positively to the length of dry periods [64].

7 Global warming, climate patterns and toxic cyanobacteria

215

7.2.2.3 Effect of rainfall events on lake nutrients Nutrient concentrations [e.g., phosphorous (P), nitrogen (N) and iron (Fe)] in water bodies can increase significantly during and after rainy periods [53, 56, 65, 68] leading to eutrophication of water bodies. In general, the processes involved are an input from the catchment (e.g., [69]), suspension of sediment by water inflow and associated wind [53] and an induction of upwelling events of nutrient-rich hypolimnic water [57, 70, 71]. However, there are different transport mechanisms for N and P with surface runoff and soil erosion being more important pathways for P [72] and the slower flushing of deeper soil waters is more important for the flush rate of nitrate–nitrogen (NO3− –N) [73]. It was also shown that the transport pathways of total P, dissolved P, particulate P, particulate N and ammonium–nitrogen (NH4+ –N) are different from those of total N, dissolved N and NO3− –N, with the first being mobilized quickly once a rainfall events occurs and the release of the latter being slower [74]. As some cyanobacteria seem to be able to take up NH4+ more efficiently than NO3 [75], this might promote an immediate response of cyanobacterial growth after rainfall events. Direct wet deposition of rainwater onto water bodies may have a direct effect on primary production through nutrient input [76, 77]. The addition of N by direct wet deposition to aquatic systems is mainly due to NH4+ [78]. It can be well above the critical loads to systems and can accelerate eutrophication in inland waters, estuaries and coastal waters [79, 80]. Direct deposition of N and P to Lake Taihu (China) by wet deposition alone was calculated as > 4,700 and 75 t year–1, respectively accounting for 16.5% and 7.3% of the annual N and P input [76, 81]. However, the chemistry of rainwater is highly variable on a regional and seasonal scale [76, 78, 79] with the regional chemical rainfall composition usually reflecting local anthropogenic emission (e.g., agriculture and industry) [45, 76, 79, 80, 82, 83]. The length of the preceding dry period is another important factor shaping the nutrient conditions in inland water bodies. Rainfall events that terminate long dry spells will lead to a comparably higher pulse of nutrients than rainfall events that occur on a regular basis, as nutrients and pollutants build up during the no-rain period (nonurban systems [84], urban systems [85]). Additionally, soil moisture, which is affected by the length of the preceding dry period, influences the discharge of N, although this correlation is N-species dependent: the highest concentrations of NO3− –N discharge occurs at relatively low soil moisture while NH4+ –N discharge is lowest at high soil moisture [74]. Also, water-soluble organic forms of P are mobilized more easily after longer preceding dry periods or after drying and rewetting due to processes such as microbial or biochemical that facilitate their release [73, 86]. The length of the preceding dry period also affects runoff volume with larger inflow from dry than from wet catchments [87]. Therefore, disproportionally more nutrients are added to water bodies due to higher runoff volume in dry catchments. Furthermore, phosphorous concentrations in lakes can decrease with longer droughts as the lake

216

Elke S. Reichwaldt, Som Cit Sinang and Anas Ghadouani

retention of nutrients might not be high enough to fully compensate the lower nutrient input due to the loss of inflow streams [60]. Compared to low-intensity rainfall events, high-intensity rainfall events are able to mobilize and dislodge larger particles and less water soaks into the soil leading to a higher fraction of rainwater input to runoff [88]. A storm with a higher intensity adds more nutrients to a water body than a less intense storm [84, 85, 89] and can lead to massive erosion resulting in very high nutrient input especially into artificial water bodies such as reservoirs [90, 91]. However, the nutrient composition added during high-intensity events will be biased toward particulate rather than soluble nutrient forms [52, 72, 89], with the former being less available for direct uptake by cyanobacteria. An initial dilution of nutrients in water bodies through the addition of large amounts of nutrient-poor rainwater is generally thought to be negligible as the subsequent input of nutrients by runoff and subsurface flows usually outweighs the dilution effect [74, 92].

7.2.2.4 Effect of rainfall events on lake conductivity and salinity The major processes regulating the conductivity in lakes are soil and rock weathering, atmospheric precipitation and evaporation [48]. This also includes the seasonality of rainfall (summer, winter), the ratio of water entering through surface runoff and groundwater flows and the conductivity of these flows. The interaction of all these processes, which is a function of catchment characteristics and meteorological parameters, determines the effect of rainfall on the conductivity of a water body [93, 94]. During wet periods, the conductivity of lakes has been shown to depend mainly on precipitation and groundwater inputs, while evaporation controls the conductivity during dry periods [33, 94]. Especially heavy rainfall events can lead to dilution and thus a reduction of the conductivity of the water body [95], because conductivity of rain is lower than that of most lakes [96]. Water loss from water bodies due to evaporation will lead to higher conductivity [64], as the evaporating water contains no or very little ions. However, in areas with high-salinity groundwater (e.g., South Australia), an increase in groundwater inflow to waters could potentially lead to a higher conductivity in the water body [33].

7.2.2.5 Effect of rainfall events on lake pH The effect of rainfall events on the pH of a water body depends on the respective acid neutralizing capacity of the water. The acid neutralizing capacity of poorly buffered systems can be reduced by increased deposition of NO3− and NH4+ [97] such as during rainfall events, leading to acidic periods during precipitation [98]. An increase in the intensity of rainfall will therefore lead to increased acidic

7 Global warming, climate patterns and toxic cyanobacteria

217

atmospheric deposition [99]. Rain that is in equilibrium with the CO2 in the air has a pH of 5.6 [100], however, emissions from industry can change the pH of rain significantly (“acid rain”). This effect is both local and long distance and can lead to an acidification of many ecosystems [100–103]. The pH of rainwater can also vary considerably on a seasonal basis (3.3–7.76) [104].

7.3 The ecology of cyanobacteria and toxin production 7.3.1 Environmental factors affecting cyanobacterial biomass Current studies have proposed a range of physicochemical factors that trigger the occurrence of cyanobacterial blooms in freshwater ecosystems (Figure 7.3). Excessive nutrient loading and increasing water temperatures are known as the basis for the presence of massive cyanobacterial biomass [14, 105]. As massive cyanobacterial blooms usually occur in eutrophic water bodies, high phosphorus and nitrogen concentrations are assumed to be the primary triggers for their growth [106, 107]. However, the environmental factors causing cyanobacterial blooms in water bodies still remain the subject of a long-standing debate as they seem to be site-specific [108–110] and species-specific (e.g., [111]). Furthermore, ecophysiological differences between cyanobacterial groups in mixed blooms have led to nonhomogenous behavior and responses to the natural environment [112]. In terms of phosphorus, Izydorczyk et al. [113] have suggested that high phosphorus concentrations promote high cyanobacterial biomass. On the other hand, de Figueiredo et al. [114] have reported that low phosphorus concentrations favor high cyanobacterial biomass. Similarly, nitrogen was also reported to have either positive [115, 116] or no correlation [117, 118] with cyanobacterial biomass. Evidence has also accumulated that the ratio of total nitrogen to total phosphorus (TN:TP ratio) is important for explaining phytoplankton and cyanobacterial growth [97, 119]; however, this relationship also differs between studies. The hypothesis of low TN:TP ratio favoring the dominance of cyanobacteria [120] has been challenged by studies which reported contrasting findings [121, 122]. The potential correlations between iron concentration in the water column and cyanobacterial biomass have been reported to be either positive [123, 124], negative [125] or have no clear correlation [126]. Higher water temperatures favor the formation of high cyanobacterial biomass through increased growth rates [127], the migration of biomass from sediment into the water column [128], increased stratification and reduced vertical mixing [14] and enhanced hypolimnetic phosphorus accumulation from sediment [22]. In addition, the effects of temperature on cyanobacterial dominance may be direct through

218

Elke S. Reichwaldt, Som Cit Sinang and Anas Ghadouani

Global warming Climate patterns Cyanobacteria-dominated community Species-rich phytoplankton community

ss Ecological factors ma Bio • Light • Temperature • Nutrients • Trace elements • Water residence time • Zooplankton grazing Microcystin production • Competition Genetic level mcy genes

Yes No Toxic strain Nontoxic strain Cellular level Figure 7.3: Conceptual model of the connection between global warming and climate patterns, and the environmental factors that affect cyanobacterial bloom development and microcystin production.

its effect on growth rates, or indirect [129, 130], for instance through its effect on water mixing and nutrient transport from sediments [14, 22]. Many studies show a positive relationship between water temperature and cyanobacteria [113, 114, 131, 132]. However, there is also evidence that this is highly species specific and that the length of the warm period is more important than the absolute temperature [133, 134]. Analogously, the correlations between temperature, light and cyanobacterial biomass are also contrasting [114, 130, 134, 135]. In terms of light, it has been suggested that low-light availability in the water column is responsible for initiating the presence of high cyanobacterial biomass [130] and especially buoyant cyanobacteria were shown to be favored under low light conditions [58, 107, 136]. However, it has also been suggested that low light conditions were not the triggers for the development of high cyanobacterial biomass, but rather caused by it [75]. The presence of these contrasting results clearly highlights that the development of cyanobacterial blooms is the result of the complex interaction between all environmental factors, especially temperature and nutrients [137, 138].

7 Global warming, climate patterns and toxic cyanobacteria

219

7.3.2 Environmental factors affecting microcystin production Microcystin is one of the most commonly detected cyanotoxin in inland water bodies worldwide, and its production has been studied extensively. The ability of cyanobacteria to produce microcystin is first determined at the genetic level with the presence of microcystin synthesis genes known as mcy genes [139–141] (Figure 7.3). Microcystin is synthesized nonribosomally by the thiotemplate functions of large multifunctional enzyme complexes. Gene clusters encoding the biosynthetic enzyme (mcyS) have been sequenced and characterized in Microcystis genera [139]. It is important to note that over 90 variants of microcystins were identified so far (e.g., microcystin-LR and microcystin-RR) [142], each of which exhibits a different toxicity to organisms [143]. In nature, cyanobacterial populations may consist of single or mixed species, and a single species may be a mixture of toxic and nontoxic strains [107, 144, 145]. Strains are specific genetic subspecies with slightly different traits [107]. As an example, 11 strains of Microcystis aeruginosa were isolated from a single cyanobacterial population [146]. The proportions of cyanobacterial strains capable of producing microcystin within a single population are highly variable, ranging from 12% [116, 147] to 80% [148]. Therefore, the concentration of microcystin produced during a bloom will to some extent depend on variations in the proportion of strains containing mcy genes [116] (Figure 7.3). In addition, microcystin content in the individual cell can vary considerably in any given strain due to different levels of expression of genes that are involved in microcystin biosynthesis [149]. There is accumulating evidence that toxic strains dominate under higher temperature (reviewed in [150, 151]). A range of environmental factors including temperature [152], light intensities [153–155], phosphorus [156], nitrogen [157], TN:TP ratio [158], iron [159] and the presence of other competing phytoplankton [160] have been suggested to correlate with the level of gene expression involved in microcystin biosynthesis. It is important to note that many studies emphasize the importance of synergistic interactions, specifically of eutrophication and temperature in shaping the toxin production (reviewed in [150]). Further, many studies indicate that toxin production is positively related to cell growth rate [152, 161–163], implying that a higher toxin production rate per cell can be found if growth conditions are improved. However, the role of these environmental factors on microcystin production still remains largely unclear and is inconsistent between studies, partly also due to site and species-specific responses [108–110, 164, 165]. Changes in temperature have been reported to cause up to threefold difference in cellular microcystin content [107]. For example, microcystin production increased from 300 to 900 µg g−1 phytoplankton dry mass as water temperature increased from 25 to 29 °C [166]. On the other hand, there are also reports suggesting that microcystin

220

Elke S. Reichwaldt, Som Cit Sinang and Anas Ghadouani

production was reduced from 2 to 1 mg g−1 dry mass when the water temperatures increased from 25 to 30 °C [167]. In general, high concentrations of nitrogen and phosphorus in the water column have been suggested to have a positive correlation with microcystin production [168, 169]. This may be due to extra energy required for toxin biosynthesis in toxic cyanobacteria [158]. For example, higher nitrogen availability was found to be associated with higher microcystin concentrations in nonnitrogen fixing cyanobacteria such as M. aeruginosa [170]. Nevertheless, studies also reported that microcystin production in M. aeruginosa is independent of nitrogen availability [115, 159]. Similar to the role of nitrogen, the effect of phosphorus on microcystin production appears to differ between studies. Wang et al. [171] and Rinta-Kanto et al. [156] have shown that the cellular microcystin content in M. aeruginosa increased with higher total phosphorus concentrations in the water column. In contrast, other studies suggest a negative correlation or no correlation between phosphorus and microcystin production [116, 172]. These conflicting findings are possibly due to nonlinear effects of nutrients on microcystin production [173]. Graham et al. [173] have demonstrated that microcystin concentrations exceeding 2 µg L − 1 occurred when the total amount of nitrogen in the water column varied between 1 to 4 mg L − 1 . In contrast, microcystin concentration decreased to below 1.5 µg L−1 when total nitrogen exceeded 8 mg L − 1 . Therefore, it can be speculated that nutrient regulation on microcystin production depends on the concentration range of nutrient present in the systems [109]. Complicating this issue further is the fact that the relative availability of N, P and carbon (C) can lead to the production of different toxins and microcystin variants, because they slightly differ in their stoichiometry [174, 175]. This was hypothesized to lead to water bodies exhibiting different toxicity depending on the nutrient and light regime [34]. Iron has been identified as an important micronutrient affecting microcystin production [115]. M. aeruginosa was found to produce 20–40% more microcystin when grown in media in the absence of or at low iron concentrations ( Phyto

CB biomass ↑

Toxin production ↑ or ↓

Conditions in water body

Higher water temperature

CB > Phyto

Table 7.3: Summary of the predicted effects of likely changes in water body conditions due to global warming on cyanobacterial blooms; ↑ = increase; ↓ = decrease; CB = cyanobacteria; Phyto = noncyanobacterial phytoplankton; CB > Phyto or Phyto > CB means that conditions favor cyanobacteria or phytoplankton, respectively (adapted from [34]).

Lower water level

Elke S. Reichwaldt, Som Cit Sinang and Anas Ghadouani

Low diversity, species-poor system; development of toxic CB bloom;

Possible consequences for ecosystem

226

Destabilized water column by water inflow and wind

Longer water residence time = stable water column (stratification) Higher toxin production under growth-favorable conditionsb Stable conditions favor CB that have lower growth rates than Phytoa, d, g, l, m

Toxin production ↑

CB > Phyto

Comparatively higher growth rate of other phytoplanktona

Phyto > CB

Eutrophicationf by nutrient input from nutrient-rich deep water layer can rapidly be utilized by CB, especially if epilimnion was nutrient limited beforeg; resuspension of CB dormant formsn Light limitation under high C, N supplyo

CB biomass ↑

Toxin production ↑

(continued )

Low diversity, species-poor system; development of toxic CB blooms;

Lower toxin production under unfavorable growth conditionsb

Toxin production ↓

Long term (after mixing conditions)

Unfavorable for CB due to their relatively slow growth High diversity system; toxin ratea concentration in the system low;

Low diversity, species-poor system; development of toxic CB bloom;

CB biomass ↓

Short term (during mixing conditions)

Nutrient depletion of the euphotic zone, CB have higher affinity to P, N and Fe than Phytoa; some buoyant CB species migrate to nutrient-rich hypolimnionk

CB biomass ↑

7 Global warming, climate patterns and toxic cyanobacteria

227

Eutrophicationf by nutrient input through surface and Low diversity, species-poor system; subsurface flows development of toxic CB blooms; higher biomass carrying capacity of the water body; Higher toxin production at higher growth ratesa, b; higher growth rates of toxic compared to nontoxic strainsc Eutrophicationf Nutrients are mobilized from the sedimentq → eutrophicationf ↑: Due to favorable growth conditionsb and increased c(Fesol)r, ↓: due to increase of c(Fesol)s Depends on species-specific optimum growth curves Depends on species-specific optimum growth curves CB are more salt tolerant than many Phytot

CB biomass ↑

Toxin production ↑

CB > Phyto

CB biomass ↑

Toxin production ↑ or ↓

Toxin production ↑ or ↓

CB > Phyto

Nutrient input

Increased salinity

Anoxic conditions

CB biomass ↑ or ↓

Dilutionp

CB and Phyto biomass ↓

Flushing

CB possibly dominant but it depends on the salinity level if blooms develop

Low diversity, species-poor system; development of CB bloom with high or low toxicity;

Low CB and Phyto biomass levels; toxin concentration in the system low;

Eutrophicationf

CB > Phyto

Possible consequences for ecosystem

Possible mechanisms

Conditions in water body

Predicted effects on CB biomass, toxin production and competition between CB and Phyto

Table 7.3 (continued )

228 Elke S. Reichwaldt, Som Cit Sinang and Anas Ghadouani

Buoyant CB avoid light limitation by forming surface scumh; CB can adapt to different light levelsg ↑: Light limitation under high C, N supplyo, ↓: nontoxic strains are better competitors for light under light limitationi Relatively higher growth rate of CB under low light intensitiesa Depends on species-specific optimum growth curves

Depends on species-specific optimum growth curves (Buoyant) CB are better competitors at higher pHk Buoyant CB avoid light limitation by forming surface scumh; CB can adapt to different light levelsg ↑: Light limitation under high C, N supplyo, ↓: Nontoxic strains are better competitors for light under light limitationi Relatively higher growth rate of CB under low light intensitiesa

CB biomass ↑

toxin production ↑ or ↓

CB > Phyto

CB biomass ↑ or ↓

Toxin production ↑ or ↓

CB > Phyto

CB biomass ↑

Toxin production ↑ or ↓

CB > Phyto

Citations are examples only: a [107], b [163], c [152], d [105], e [31], f [208], g [58], h [136], i [209], j [210], k [32], l [211], m [25], n [70], o [175], p [56], q [90], r [212], s [213], t [214].

High turbidity due to high biomass

Higher pH due to high photosynthesis levels

Increased turbidity due to sediment resuspension

Low diversity, species-poor system; development of CB bloom with high or low toxicity;

CB possibly dominant but it depends on the pH if blooms develop;

Low diversity, species-poor system; development of CB bloom with high or low toxicity;

7 Global warming, climate patterns and toxic cyanobacteria

229

230

Elke S. Reichwaldt, Som Cit Sinang and Anas Ghadouani

effects of other environmental factors (Tilman et al. 1986, as cited in [129]), making any prediction of bloom development and therefore management of the water body extremely complex. Global warming has been shown to be responsible for a wider geographical distribution of cyanobacteria worldwide. The best-known example is the invasion of Cylindrospermopsis raciborskii, a tropical and subtropical species into temperate water bodies (e.g., [216, 217]). It has been suggested that this species is able to extend its current distribution due to higher average water and sediment temperatures and due to an ongoing eutrophication of many water bodies [218]. More recently, it was also suggested that the distribution of Microcystis could expand under future climate scenarios [165]. The higher stability of the stratification during periods of higher temperature and no rain might favor buoyant cyanobacterial genera such as Anabaena, Aphanizomenon and Microcystis as they are promoted during nonmixing phases [111, 219, 220], although Oscillatoria growth was shown to be stimulated by mixing events [221]. In regions with distinctive dry seasons, blooms prevail longer if the start of the rain is delayed or the rainfall events are too weak to break up the stratification [58]. Higher temperature and long dry periods can also increase the water retention time, and cyanobacteria were shown to dominate during periods of low flushing [25, 26]. The inflow of high volumes of water during rain events can lead to a direct reduction of algal biomass due to high flushing rates [51, 54, 56, 57, 63, 204] (Table 7.2), and can significantly change the community composition possibly due to persistent high turbidity and turbulent conditions which favor diatoms and small-celled cyanobacteria over large-celled or filamentous cyanobacteria [53]. Additionally, heavy rainfall events can lead to a mixing of the water column [56, 63, 205], which might favor noncyanobacterial species [63]. Thus, in the short term, intense rainfall can lead to a lower total chlorophyll biomass with a higher diversity due to the absence of cyanobacterial dominance [64]. A movement of the thermocline by water inflow or winds associated with storms can also allow nutrients that are released from sediments to enter the photic zone, leading to an additional nutrient enrichment and can resuspend cells (e.g., Microcystis spp.) or dormant forms of cyanobacteria (e.g., akinetes from Cylindrospermopsis spp.) into the water column which encourages bloom development [65, 70].

7.4.2 Nutrients In general, increased atmospheric temperature can lead to eutrophication supported by lower water levels, longer water residence time, an increased release of nutrients (e.g., soluble P, NH4+ and Fe) from the sediment if anoxic hypolimnia develop [90] and increased mineralization rates at higher temperature [22, 23]. The nutrients released from the sediment can be utilized by buoyant cyanobacteria that are able to control their vertical movement in the water column [134] and additionally contribute to the eutrophication process during mixing events [107]. All these conditions will

7 Global warming, climate patterns and toxic cyanobacteria

231

strongly favor freshwater cyanobacteria [56, 63, 211] and it was demonstrated that cyanobacterial blooms dominate during drought and falling water levels [51] and that surface scums can occur during dry periods with low wind velocities [203]. Less intense rainfall events can immediately increase cyanobacterial biomass through nutrient enrichment if the event does not lead to destratification as is often the case for isolated rainfall events or in shallow, nonstratified lakes [58, 200]. The importance of nutrient addition to inland water bodies by direct wet deposition can be significant [76, 79, 81]. Especially in closed lakes, nutrient addition through precipitation is the main source for phytoplankton [222]. However, there is still little understanding of the magnitude of this impact, especially in the context of cyanobacterial bloom development. One study indicated a clear connection between the addition of N from wet deposition and the occurrence Microcystis [223]. Also, the TN:TP ratio of wet and dry deposition has been identified as an important factor to drive the TN:TP ratio and phytoplankton biomass in alpine lakes [224], and it was shown that dominance of nonnitrogen-fixing cyanobacterial species (e.g., M. aeruginosa] depend more on the TN:TP ratio than on TN alone [107, 225]. Therefore, an addition of N through precipitation will also affect systems that are not N limited by definition, such as many inland water bodies, and it has been emphasized that N can play a major role in eutrophic [226–228] and tropical freshwater systems [229]. This is supported by a study that suggested that atmospheric deposition of N and P might have promoted cyanobacterial blooms during periods that are in other respect optimal for their growth [230]. Prolonged periods of wind, which are related to more intense storms or longer periods of rain, can enhance the effect of nutrient input during rainfall by mixing events for up to three months (Bormans et al. 2001 in [58]), and the large nutrient input during intense events will lead to nutrient-rich conditions which generally favor cyanobacteria.

7.4.3 Salinity In general, higher atmospheric temperature can lead to increased salinity due to high evaporation rates. An increased salinity might selectively favor cyanobacteria, as many species are more salt tolerant than other phytoplankton species. For instance, long-term growth of Microcystis aeruginosa was not affected by salinity up to 9.8 g L − 1 [210] with short-term tolerance to salinities as high as 17.5 g L − 1 [214]. In contrast, rainfall events can decrease salinity [95]. Therefore, the first period after a rainfall event in which salinity is decreased only slightly while there is already a nutrient pulse might represent an important timeslot that gives cyanobacteria a competitive advantage.

232

Elke S. Reichwaldt, Som Cit Sinang and Anas Ghadouani

7.4.4 Turbidity and pH Cyanobacteria and specifically buoyant genera are highly adaptive to different light levels, and higher turbidity generally favors cyanobacteria [136]. Sudden low light availability after a rainfall event, in combination with turbulence, was shown to lead to a decrease of large-celled cyanobacterial species [53] (Table 7.2). Increased light availability was shown to correlate well with the biomass of a M. aeruginosa bloom in the presence of a sufficient nitrogen source [24], while a decrease in nitrogen was shown to favor N-fixing cyanobacteria [231]. If the higher turbidity is the result of algal growth, this can lead to a higher pH due to increased photosynthesis; in turn, this would limit dissolved inorganic carbon (DIC) in the water column. Both conditions would favor buoyant cyanobacterial species (e.g., Microcystis, Anabaena and Planktothrix) over other phytoplankton as their ability to dwell directly on the surface (scum) enables them to harvest light and CO2 from the atmosphere thus avoiding DIC limitation [136].

7.5 Direct and indirect effects of global warming on microcystin concentration There are no studies that directly investigate changes in cyanotoxin concentration in water bodies following rainfall events or prolonged periods of heat. However, plenty of information on the drivers of toxin production is available from laboratory studies. Furthermore, numerous field studies tried to correlate in-situ toxin concentrations to environmental conditions. All these results can be used to draw conclusions on how toxin concentrations might change with global warming. Cyanotoxin concentrations in a water body depend on the toxin content within each cyanobacterial cell, the biomass of toxic cyanobacteria and on toxin-degrading processes. Therefore, it is critical to understand the possible effects of global warming on all these processes. Cyanobacterial toxin concentration in the water is also a function of the succession of species and strains. Conditions leading to species-specific proliferation of cyanobacterial cells is fairly well understood; for instance, Oscillatoria can outcompete other phytoplankton and cyanobacteria under mixed, low light conditions, while Anabaena, Aphanizomenon and Microcystis will dominate under stable conditions with higher underwater irradiance [220]. In contrast, the factors that lead to the dominance of toxic or nontoxic strains are still relatively unknown [232]. It can be hypothesized that toxin concentration in the water will be higher under future rainfall scenarios and global warming due to an increase in cyanobacterial biomass alone [233]. However, higher biomass does not necessarily mean that growth conditions (and thus the toxin production rate) are optimal for cyanobacteria. Therefore, whether global warming will lead to higher or lower average toxin concentration in a water body will ultimately

7 Global warming, climate patterns and toxic cyanobacteria

233

depend on the effect of temperature and rainfall patterns on (variant)-specific toxin production rates [34]. As discussed previously, the toxin production per cell is a function of the environmentally influenced gene expression [107]. A range of environmental factors including temperature [152] have been suggested to correlate with the level of gene expression involved in microcystin biosynthesis. As with the effect of temperature on cyanobacterial growth, the direct effect of temperature on toxin production seems to be species and strain specific, and is intertwined with the effect of other factors (reviews [107, 161]). As changes in temperature, that have been reported, result in an increase [166] or decrease [167] of the microcystin production, no simple prediction can be made on the effect of increased water temperature on the toxin concentration in the water body. However, there is accumulating evidence that toxic strains dominate under higher temperature (reviewed in [150, 151]) emphasizing that the future might hold more toxic blooms. Decreased turbidity following a heavy rainfall event or due to high biomass during periods of high temperature and/or no rain can lead to higher light penetration; this might increase microcystin production [234]. The second important factor affecting the toxin concentration in the water is the biomass of toxic cyanobacteria [149]. Very often, total cyanobacterial biomass is used as an indicator of the toxin concentration in the water column [235]; this works well if a bloom consists of single species and strains. However, cyanobacterial blooms often contain a mix of toxic and nontoxic species. Even those species that are potentially toxic are comprised of toxic and nontoxic strains (genetic subgroups of species) [107, 236] that might have different optimum growth conditions. For instance, nontoxic strains of M. aeruginosa were shown to be better competitors for light than toxic strains [209], while higher water temperature and higher concentrations of inorganic N and P are likely to promote toxic strains ([228], reviewed in [150]). It has also been demonstrated that the succession of different strains was responsible for high variations in toxin concentration in the water over the lifetime of a bloom [237]. Therefore, any predictions of the effects of global warming on the biomass dynamics of toxic cyanobacteria have to be based on our understanding of toxic and nontoxic species and strain succession, especially as there is evidence that the toxin production rate per cell is often positively correlated with cyanobacterial growth [152, 161–163]. However, little is known in this respect and results are highly ambiguous, indicating that favorable growth conditions for cyanobacteria can lead to the dominance of nontoxic or toxic strains [148, 149, 152, 162]. The third mechanism that influences the total toxin concentration in the water includes processes by which the concentrations of dissolved toxins are decreased [238]. These processes include dilution, biodegradation by bacteria [239, 240], adsorption onto particulate organic matter and sediments [240–242], thermal decomposition and photolysis [243]. No direct study of the effects of rainfall events on these processes has been performed, but it can be inferred from the literature that increased water temperature through higher atmospheric temperature or prolonged

234

Elke S. Reichwaldt, Som Cit Sinang and Anas Ghadouani

dry periods with high evaporation enhances thermal biodegradation and increases bacterial biodegradation [244]. Heavy rainfall events that input large amounts of sediment and particulate organic matter and/or stirs up sediment could increase the amount of adsorbed toxins to sediment or alternatively release loosely bound toxins into the water column [245]. The input of large water volumes could dilute the dissolved toxin fraction, and changes in pH can lead to changes in the adsorption of toxins onto sediments, with a decreased sorption at higher pH [246]. Most of these processes would decrease the concentration of dissolved toxins in the water column. However, as the dissolved toxin fraction is usually small compared to the intracellular fraction, the impact of possible changes with global warming might be less relevant than mechanisms that influence the intracellular fraction.

7.6 Why should we care? How are we responding to the cyanobacterial problem in its new context of climate change? Cyanobacterial blooms have been around for a long time [247] and, since then, there has been an enormous scientific interest in understanding the dynamics

Society Policies

Policies Public health

Treatment costs

Treatment costs

Drinking water

Health Waste water

Infrastructure

Infrastructure Toxic cyanobacterial bloom

Reuse

Policies Public health Treatment Environment costs Infrastructure

Public health

Recreation

Food Infrastructure

Policies Treatment costs

Figure 7.4: Important implications of cyanobacterial blooms for the society.

7 Global warming, climate patterns and toxic cyanobacteria

235

of cyanobacteria and their harmful compounds. Cyanobacteria have been considered as a “nuisance” [211] that is harmful to aquatic organisms and humans [113], and as such represent a “global public health threat” [248]. This potential hazard is reflected in numerous guidelines adopting critical values for accepted toxin concentrations in drinking and recreational waters [232, 235, 249]. These guidelines are an important response to minimize cyanobacterial-related health issues, especially in the light of global change that is very likely to increase cyanobacterial occurrence. More recently, awareness has been raised that, apart from a health related impact (e.g., [250]), cyanobacterial blooms have implications for many other sectors, such as the water resource sector, infrastructure, policies and the social sector. It is important that this awareness is reflected in public documents to ensure an adequate reply to this increasing problem. An extensive series on the “Perspectives on water and climate change adaptation” by the World Water Council, the Cooperative Program on Water and Climate and the International Water Association in cooperation with other contributors was a valuable contribution to identify areas of impact on water by climate change with the aim to “define and distil the critical role of water in climate change adaptation and to lay out strategic and operational priorities for adaptation of water management and services” [251]. In the included papers, water quality was identified as a serious problem for food production [252], the water industry [253] and WASHservice delivery [254]. Interestingly, cyanobacteria were only mentioned once, indicating that the link between industry and science is not complete. This is of particular concern as “information and data collection and sharing” has been identified as an extremely important strategy to adapt to climate change in all water-use sectors [251]. Furthermore, the Intergovernmental Panel on Climate Change [99] acknowledges that adaptation procedures including risk management practices are being developed by the water sector in some countries, but this is mainly directed to hydrological changes [255]. The implications of the deterioration of water quality, including cyanobacterial blooms by climate change for the water sector, are described as being substantial but seem to be less developed. The implications of cyanobacteria on our life are manifold (Figure 7.4) and the intensity of the impacts can be expected to multiply in the future. While the direct implications on the water resource sector, including wastewater and drinking water treatment, are obvious, the impact of cyanobacteria on the infrastructure sector, the food sector and the society are less explicit. Cyanobacterial blooms are now regularly found in drinking water reservoirs (e.g., [256, 257], which increases the likelihood of the presence of toxins and taste and odor compounds in this water resource [26]). Their presence incurs high costs for infrastructure maintenance, treatment of the water and for provision of alternative drinking water to the community during blooms. New, cost-

236

Elke S. Reichwaldt, Som Cit Sinang and Anas Ghadouani

effective treatment methods for the removal of toxins (e.g., [258]) and taste and odor compounds are needed to counteract increasing costs in the future. Further, an increase in the occurrence of cyanobacterial blooms during wastewater treatment processes will hinder to achieve high water quality effluent that can be reused for multiple purposes, including irrigation. Similar to the problems with drinking water, costefficient treatment methods have to be developed (e.g., [259–261]) to allow the safe future use of this important water resource. The occurrence of cyanobacteria will certainly also impact our food security due to less clean water being available for irrigation and watering of animals. It will make agricultural products more expensive due to higher costs for treating the water before use or transporting good quality water. Aquaculture and fisheries will also be impacted by the future increase in cyanobacteria, both in inland water bodies and the ocean [262–266]. The presence of blooms will lead to shorter fishing seasons. This, in combination with the necessity to conduct more rigorous monitoring of toxin concentrations in seafood, can lead to enormous economic losses [266]. An often-forgotten ecosystem service that water delivers is its role as a recreation source [267]. The recreational value of lakes, rivers, estuaries and the coast will very likely suffer in the future. A recent study showed that in some countries the recreational use of lakes is at risk due to cyanobacteria and high nutrient levels in as many as 50% of the studied lakes [268]. This indicates not only a huge impact on recreation, but also large economic losses due to the cleaning of infrastructure and water equipment, such as boats and boating harbors’ infrastructure. This is not only expensive but potentially also harmful. Additionally, tourism and recreational fishing have been shown to be negatively affected by blooms (e.g., [250]) and this is likely to increase even further in the future. In summary, toxic cyanobacterial blooms are directly and indirectly impacting society, making this a global problem that is likely to worsen in the future. So, where do we go from here? There seems to be sufficient scientific evidence that the occurrence of toxic cyanobacterial blooms will very likely increase in many water bodies and regions worldwide due to a combination of climate change, eutrophication and man-made changes of the hydrology of water bodies. Although the exact bloom dynamics will certainly be site-specific, the general patterns seems to be clear. Therefore, it is now time to take the next logical step and focus on defining the implication of toxic cyanobacterial blooms for the public, for the governments, for businesses and, most importantly, for the water sector. This requires a close collaboration between social and natural sciences and will need substantial support by the private sector, politics and stakeholders. Just like the dynamics of cyanobacterial blooms is a complex issue, solving it will have to be a collective effort. And only when these implications have been clearly identified and perceived by the public, can we start developing clear strategies to adapt to a blue-green future.

7 Global warming, climate patterns and toxic cyanobacteria

237

References [1] [2] [3] [4] [5] [6] [7] [8]

[9] [10]

[11]

[12]

[13] [14] [15] [16] [17] [18]

[19]

Dole R, Hoerling M, Perlwitz J, Eischeid J, Pegion P, Zhang T, Quan XW, Xu TY, Murray D. Was there a basis for anticipating the 2010 Russian heat wave?. Geophys Res Lett 2011:38. Rahmstorf S, Coumou D Increase of extreme events in a warming world. Proceedings of the National Academy of Sciences of the United States of America 2011;108:17905–9. Luo LF, Zhang Y. Did we see the 2011 summer heat wave coming?. Geophys Res Lett 2012:39. Lewis SC, Karoly DJ. Anthropogenic contributions to Australia’s record summer temperatures of 2013. Geophys Res Lett 2013;40:3705–9. Sillmann J, Donat MG, Fyfe JC, Zwiers FW. Observed and simulated temperature extremes during the recent warming hiatus. Environ Res Lett 2014:9. Christidis N, Jones GS, Stott PA. Dramatically increasing chance of extremely hot summers since the 2003 European heatwave. Nat Clim Chang 2015;5:46–50. Beniston M. The 2003 heat wave in Europe: A shape of things to come? An analysis based on Swiss climatological data and model simulations. Geophys Res Lett 2004:31. Schneider SH, Semenov S, Patwardhan A, Burton I, Magadza CHD, Oppenheimer M, Pittock AB, Rahman A, Smith JB, Suarez A, Yamin F. Assessing key vulnerabilities and the risk from climate change. In: Parry ML, Canziani OF, Palutikof JP, Van der Linden PJ, Hanson CE [eds.] Climate Change 2007: Impacts, Adaptation and Vulnerability. Contribution of Working Group II to the Fourth Assessment Report of the Intergovernmental Panel on Climate Change. Cambridge, UK; Cambridge University Press: 2007. Meehl GA, Tebaldi C. More intense, more frequent, and longer lasting heat waves in the twenty-first century. Science 2004;305:994–7. Diaz J, Jordan A, Garcia R, Lopez C, Alberdi JC, Hernandez E, Otero A. Heat waves in Madrid 1986–1997: effects on the health of the elderly. Int Arch Occup Environ Health 2002;75: 163–70. Department of Health 2011. Heatwave plan for Victoria – Protecting health and reducing harm from heatwaves. State Government Victoria [ed.]. Available from: http://www.health.vic.gov. au/environment/heatwaves-plan.htm. Jankowski T, Livingstone DM, Buhrer H, Forster R, Niederhauser P. Consequences of the 2003 European heat wave for lake temperature profiles, thermal stability, and hypolimnetic oxygen depletion: Implications for a warmer world. Limnol Oceanogr 2006;51:815–9. Jöhnk KD, Huisman J, Sharples J, Sommeijer B, Visser PM, Stroom JM. Summer heatwaves promote blooms of harmful cyanobacteria. Glob Chang Biol 2008;14:495–512. Paerl HW, Huisman J. Blooms like it hot. Science 2008;320:57–8. Scheffer M, Carpenter SR. Catastrophic regime shifts in ecosystems: linking theory to observation. Trends Ecol Evol 2003;18:648–56. Mayer AL, Rietkerk M. The dynamic regime concept for ecosystem management and restoration. Bioscience 2004;54:1013–20. Jentsch A, Beierkuhnlein C. Research frontiers in climate change: Effects of extreme meteorological events on ecosystems. C R Geosci 2008;340:621–28. Meehl GA, Stocker TF, Collins WD, Friedlingstein P, Gaye AT, Gregory JM, Kitoh A, Knutti R, Murphy JM, Noda A, Raper SCB, Watterson IG, Weaver AJ, Zhao Z-C. Global climate projections. In: Solomon S, Qin D, Manning M, Chen Z, Marquis M, Averyt KB, Tignor M, Miller HL [eds.] Climate Change 2007: The Physical Science Basis. Contribution of Working Group I to the Fourth Assessment Report of the Intergovernmental Panel on Climate Change. Cambridge, United Kingdom and New York, NY, USA; Cambridge University Press: 2007. Murdoch PS, Baron JS, Miller TL. Potential effects of climate chance on surface-water quality in North America. J Am Water Resou Assoc 2000;36:347–66.

238

Elke S. Reichwaldt, Som Cit Sinang and Anas Ghadouani

[20] Wilhelm S, Adrian R. Impact of summer warming on the thermal characteristics of a polymictic lake and consequences for oxygen, nutrients and phytoplankton. Freshw Biol 2008;53:226–37. [21] Winder M, Schindler DE. Climatic effects on the phenology of lake processes. Glob Chang Biol 2004;10:1844–56. [22] Søndergaard M, Jensen JP, Jeppesen E. Role of sediment and internal loading of phosphorus in shallow lakes. Hydrobiologia 2003;506:135–45. [23] Gomez E, Fillit M, Ximenes MC, Picot B. Phosphate mobility at the sediment-water interface of a Mediterranean lagoon (etang du Mejean), seasonal phosphate variation. Hydrobiologia 1998;374:203–216. [24] Lehman PW, Marr K, Boyer GL, Acuna S, Teh SJ. Long-term trends and causal factors associated with Microcystis abundance and toxicity in San Francisco Estuary and implications for climate change impacts. Hydrobiologia 2013;718:141–58. [25] Cross ID, Mcgowan S, Needham T, Pointer CM. The effects of hydrological extremes on former gravel pit lake ecology: management implications. Fundam Appl Limnol 2014;185: 71–90. [26] Winston B, Hausmann S, Scott JT, Morgan R. The influence of rainfall on taste and odor production in a south-central USA reservoir. Freshwater Sci 2014;33:755–764. [27] IPCC. Climate Change 2007: The Physical Science Basis. In: Solomon S, Qin D, Manning M, Chen Z, Marquis M, Averyt KB, Tignor M, Miller HL [eds.] Contribution of Working Group I to the fourth Assessment Report of the Intergovernmental Panel on Climate Change. Cambridge, United Kingdom and New York, NY, USA; Contribution of Working Group I to the Fourth Assessment Report of the Intergovernmental Panel on Climate Change: 2007b. [28] Codd GA. Cyanobacterial toxins, the perception of water quality, and the prioritisation of eutrophication control. Ecol Eng 2000;16:51–60. [29] Desai AR, Austin JA, Bennington V, McKinley GA. Stronger winds over a large lake in response to weakening air-to-lake temperature gradient. Nat Geosci 2009;2:855–8. [30] Snucins E, Gunn J. Interannual variation in the thermal structure of clear and colored lakes. Limnol Oceanogr 2000;45:1639–46. [31] Elliott JA. Is the future blue-green? A review of the current model predictions of how climate change could affect pelagic freshwater cyanobacteria. Water Res 2012;46:1364–71. [32] Reynolds CS, Walsby AE. Water-Blooms. Biol Rev Camb Philos Soc 1975;50:437–81. [33] Tweed S, Leblanc M, Cartwright I. Groundwater-surface water interaction and the impact of a multi-year drought on lakes conditions in South-East Australia. J Hydrol 2009;379:41–53. [34] Reichwaldt ES, Ghadouani A. Effects of rainfall patterns on toxic cyanobacterial blooms in a changing climate: between simplistic scenarios and complex dynamics. Water Res 2012;46: 1372–93. [35] Allen MR, Ingram WJ. Constraints on future changes in climate and the hydrologic cycle. Nature 2002;419:224–33. [36] Alpert P, Ben-Gai T, Baharad A, Benjamini Y, Yekutieli D, Colacino M, Diodato L, Ramis C, Homar V, Romero R, Michaelides S, Manes A. The paradoxical increase of Mediterranean extreme daily rainfall in spite of decrease in total values. Geophys Res Lett 2002:29. [37] Easterling DR, Evans JL, Groisman PY, Karl TR, Kunkel KE, Ambenje P. Observed variability and trends in extreme climate events: A brief review. Bull Am Meteorol Soc 2000;81:417–25. [38] Groisman PY, Knight RW, Easterling DR, Karl TR, Hegerl GC, Razuvaev VAN. Trends in intense precipitation in the climate record. J Clim 2005;18:1326–50. [39] Brunetti M, Maugeri M, Monti F, Nanni T. Changes in daily precipitation frequency and distribution in Italy over the last 120 years. J Geophys Res Atmos 2004:109.

7 Global warming, climate patterns and toxic cyanobacteria

239

[40] Yu B, Neil DT. Long-term variations in regional rainfall in the south-west of Western Australia and the difference between average and high intensity rainfalls. Int J Climatol 1993;13: 77–88. [41] Groisman PY, Knight RW. Prolonged dry episodes over the conterminous united states: New tendencies emerging during the last 40 years. J Clim 2008;21:1850–62. [42] Karl TR, Knight RW. Secular trends of precipitation amount, frequency, and intensity in the United States. Bull Am Meteorol Soc 1998;79:231–41. [43] De Toffol S, Laghari AN, Rauch W. Are extreme rainfall intensities more frequent? Analysis of trends in rainfall patterns relevant to urban drainage systems. Water Sci Technol 2009;59: 1769–76. [44] Jassby AD, Reuter JE, Axler RP, Goldman CR, Hackley SH. Atmospheric deposition of nitrogen and phosphorus in the annual nutrient load of Lake Tahoe (California Nevada). Water Resour Res 1994;30:2207–16. [45] Zhang J, Liu MG. Observations on nutrient elements and sulfate in atmospheric wet depositions over the northwest pacific coastal oceans – Yellow Sea. Mar Chem 1994;47: 173–189. [46] Alvarez-Cobelas M, Cirujano S, Rojo C, Rodrigo MA, Pina E, Rodriguez-Murillo JC, Montero E. Effects of changing rainfall on the limnology of a Mediterranean, flowthrough-seepage chain of lakes. Int Rev Hydrobiol 2006;91:466–82. [47] Winter TC. Relation of streams, lakes, and wetlands to groundwater flow systems. Hydrogeol J 1999;7:28–45. [48] Wetzel RG. Limnology. San Diego; Academic Press: 2001. [49] Heathwaite AL, Johnes PJ. Contribution of nitrogen species and phosphorus fractions to stream water quality in agricultural catchments. Hydrol Process 1996;10:971–83. [50] Noges P, Noges T, Ghiani M, Sena F, Fresner R, Friedl M, Mildner J. Increased nutrient loading and rapid changes in phytoplankton expected with climate change in stratified South European lakes: sensitivity of lakes with different trophic state and catchment properties. Hydrobiologia 2011;667:255–70. [51] Harris GP, Baxter G. Interannual variability in phytoplankton biomass and species composition in a subtropical reservoir. Freshw Biol 1996;35:545–60. [52] Huang T, Li X, Rijnaarts H, Grotenhuis T, Ma W, Sun X, Xu J. Effects of storm runoff on the thermal regime and water quality of a deep, stratified reservoir in a temperate monsoon zone, in Northwest China. Sci Total Environ 2014b;485:820–7. [53] James RT, Chimney MJ, Sharfstein B, Engstrom DR, Schottler SP, East T, Jin KR. Hurricane effects on a shallow lake ecosystem, Lake Okeechobee, Florida (USA). Fundam Appl Limnol 2008;172:273–87. [54] Ahn CY, Chung AS, Oh HM. Rainfall, phycocyanin, and N: P ratios related to cyanobacterial blooms in a Korean large reservoir. Hydrobiologia 2002;474:117–24. [55] Hart RC. Cladoceran periodicity patterns in relation to selected environmental factors in two cascading warm-water reservoirs over a decade. Hydrobiologia 2004;526:99–117. [56] Jones GJ, Poplawski W. Understanding and management of cyanobacterial blooms in subtropical reservoirs of Queensland, Australia. Water Sci Technol 1998;37:161–8. [57] Figueredo CC, Giani A. Seasonal variation in the diversity and species richness of phytoplankton in a tropical eutrophic reservoir. Hydrobiologia 2001;445:165–74. [58] Shaw G, Garnett C, Moore MR, Florian P. The predicted impact of climate change on toxic algal (Cyanobacterial) blooms and toxin production in Queensland. Environ Health 2001:1. [59] Tranvik LJ, Downing JA, Cotner JB, Loiselle SA, Striegl RG, Ballatore TJ, Dillon P, Finlay K, Fortino K, Knoll LB, Kortelainen PL, Kutser T, Larsen S, Laurion I, Leech DM, McCallister SL, McKnight DM, Melack JM, Overholt E, Porter JA, Prairie Y, Renwick WH, Roland F, Sherman

240

[60] [61] [62] [63]

[64]

[65] [66] [67] [68] [69] [70]

[71]

[72] [73]

[74] [75]

[76] [77] [78] [79]

Elke S. Reichwaldt, Som Cit Sinang and Anas Ghadouani

BS, Schindler DW, Sobek S, Tremblay A, Vanni MJ, Verschoor AM, Von Wachenfeldt E, Weyhenmeyer GA. Lakes and reservoirs as regulators of carbon cycling and climate. Limnol Oceanogr 2009;54:2298–314. Schindler DW. Lakes as sentinels and integrators for the effects of climate change on watersheds, airsheds, and landscapes. Limnol Oceanogr 2009;54:2349–58. Hofmann H, Lorke A, Peeters F. Temporal scales of water-level fluctuations in lakes and their ecological implications. Hydrobiologia 2008;613:85–96. Anda A, Varga B. Analysis of precipitation on Lake Balaton catchments from 1921 to 2007. Idojaras 2010;114:187–201. Bouvy M, Nascimento SM, Molica RJR, Ferreira A, Huszar V, Azevedo SMFO. Limnological features in Tapacura reservoir (northeast Brazil) during a severe drought. Hydrobiologia 2003;493:115–30. Bouvy M, Molica R, De Oliveira S, Marinho M, Beker B. Dynamics of a toxic cyanobacterial bloom (Cylindrospermopsis raciborskii) in a shallow reservoir in the semi-arid region of northeast Brazil. Aquat Microb Ecol 1999;20:285–97. Kebede E, Belay A. Species composition and phytoplankton biomass in a tropical African lake (Lake Awassa, Ethiopia). Hydrobiologia 1994;288:13–32. Kuhn NJ, Baumhauer R, Schutt B. Managing the impact of climate change on the hydrology of the Gallocanta Basin, NE-Spain. J Environ Manage 2011;92:275–83. An KG, Jones JR. Temporal and spatial patterns in salinity and suspended solids in a reservoir influenced by the Asian monsoon. Hydrobiologia 2000b;436:179–89. Zaw M, Chiswell B. Iron and manganese dynamics in lake water. Water Res 1999;33:1900–10. Dillon PJ, Kirchner WB. Effects of geology and land-use on export of phosphorous from watersheds. Water Res 1975;9:135–148. Fabbro LD, Duivenvoorden LJ. Profile of a bloom of the cyanobacterium Cylindrospermopsis raciborskii (Woloszynska) Seenaya and Subba Raju in the Fitzroy River in tropical central Queensland. Mar Freshwater Res 1996;47:685–94. Figueredo CC, Giani A. Phytoplankton community in the tropical lake of Lagoa Santa (Brazil): Conditions favoring a persistent bloom of Cylindrospermopsis raciborskii. Limnologica 2009;39:264–72. Gentry LE, David MB, Royer TV, Mitchell CA, Starks KM. Phosphorus transport pathways to streams in tile-drained agricultural watersheds. J Environ Qual 2007;36:408–15. Stutter MI, Langan SJ, Cooper RJ. Spatial contributions of diffuse inputs and within-channel processes to the form of stream water phosphorus over storm events. J Hydrol 2008;350: 203–14. Kato T, Kuroda H, Nakasone H. Runoff characteristics of nutrients from an agricultural watershed with intensive livestock production. J Hydrol 2009;368:79–87. Presing M, Herodek S, Voros L, Kobor I. Nitrogen fixation, ammonium and nitrate uptake during a bloom of Cylindrospermopsis raciborskii in Lake Balaton. Archiv für Hydrobiologie 1996;136:553–62. Luo LC, Qin BQ, Song YZ, Yang LY. Seasonal and regional variations in precipitation chemistry in the Lake Taihu Basin, China. Atmos Environ 2007a;41:2674–9. Buijsman E, Erisman JW. Wet deposition of ammonium in Europe. J Atmos Chem 1988;6: 265–80. Kopacek J, Prochazkova L, Hejzlar J, Blazka P. Trends and seasonal patterns of bulk deposition of nutrients in the Czech Republic. Atmos Environ 1997;31:797–808. Rogora M, Mosello R, Arisci S, Brizzio M, Barbieri A, Balestrini R, Waldner P, Schmitt M, Stahli M, Thimonier A, Kalina M, Puxbaum H, Nickus U, Ulrich E, Probst A. An overview of

7 Global warming, climate patterns and toxic cyanobacteria

[80] [81] [82]

[83] [84]

[85] [86] [87] [88] [89]

[90]

[91]

[92] [93] [94] [95] [96] [97] [98]

[99]

241

atmospheric deposition chemistry over the Alps: Present status and long-term trends. Hydrobiologia 2006;562:17–40. Russell KM, Galloway JN, Macko SA, Moody JL, Scudlark JR. Sources of nitrogen in wet deposition to the Chesapeake Bay region. Atmos Environ 1998;32:2453–2465. Luo LC, Qin BQ, Yang LY, Song YZ. Total inputs of phosphorus and nitrogen by wet deposition into Lake Taihu, China. Hydrobiologia 2007b;581:63–70. Lara ER, Guardiola RM, Vasquez YG, Renteria IB, Alvarez HB, Echeverria RS, Alvarez PS, Jimenez AA, Torres MC, Kahl J. Chemical composition of rainwater in northeastern Mexico. Atmosfera 2010;23:213–24. Raper DW, Lee DS. Wet deposition at the sub-20 km scale in a rural upland area of England. Atmos Environ 1996;30:1193–1207. Kleinman PJA, Srinivasan MS, Dell CJ, Schmidt JP, Sharpley AN, Bryant RB. Role of rainfall intensity and hydrology in nutrient transport via surface runoff. J Environ Qual 2006;35: 1248–59. Davis AP, McCuen RH. Stormwater management for smart growth. New York; Springer: 2005. Turner BL, Haygarth PM. Biogeochemistry – Phosphorus solubilization in rewetted soils. Nature 2001;411:258–258. Chiew FHS, Whetton PH, McMahon TA, Pittock AB. Simulation of the impacts of climate change on runoff and soil moisture in Australian catchments. J Hydrol 1995;167:121–47. Trenberth KE, Dai A, Rasmussen RM, Parsons DB. The changing character of precipitation. Bull Am Meteorol Soc 2003;84:1205–17. Budai P, Clement A. Estimation of nutrient load from urban diffuse sources: experiments with runoff sampling at pilot catchments of Lake Balaton, Hungary. Water Sci Technol 2007;56: 295–302. Chorus I, Mur L. Preventive Measures. In: Chorus I, Bartram E [eds.] Toxic Cyanobacteria in Water: A guide to their public health consequences, monitoring and management. E & FN Spon: 1999. Huang T-L, Li X, Ma W-X, Qin C-H, Zhang Y-T. Dynamic characteristics of nutrients and causal analysis in eutrofic reservoir: a case study of Shibianyu reservoir. Desalin Water Treat 2014a;52:1624–35. Mez K, Hanselmann K, Preisig H. Environmental conditions in high mountain lakes containing toxic benthic cyanobacteria. Hydrobiologia 1998;368:1–15. Evans CD, Prepas EE. Potential effects of climate change on ion chemistry and phytoplankton communities in prairie saline lakes. Limnol Oceanogr 1996;41:1063–76. Pham SV, Leavitt PR, McGowan S, Peres-Neto P. Spatial variability of climate and land-use effects on lakes of the northern Great Plains. Limnol Oceanogr 2008;53:728–42. Badve RM, Kumaran KPN, Rajshekhar C. Eutrophication of Lonar Lake, Maharashtra. Curr Sci 1993;65:347–51. Gibson CE, Wu Y, Smith SJ, Wolfe-Murphy SA. Synoptic limnology of a diverse geological region – catchment and water chemistry. Hydrobiologia 1995;306:213–27. Rabalais NN. Nitrogen in aquatic ecosystems. Ambio 2002;31:102–12. Evans CD, Reynolds B, Hinton C, Hughes S, Norris D, Grant S, Williams B. Effects of decreasing acid deposition and climate change on acid extremes in an upland stream. Hydrol Earth Syst Sci 2008;12:337–51. IPCC. Climate Change 2007: Impacts, Adaptation and Vulnerability. In: Parry ML, Canziani OF, Palutikof JP, Van der Linden PJ, Hanson CE [eds.] Contribution of Working Group II to the Fourth Assessment Report of the Intergovernmental Panel on Climate Change. Cambridge. UK; Contribution of Working Group II to the Fourth Assessment Report of the Intergovernmental Panel on Climate Change: 2007a.

242

Elke S. Reichwaldt, Som Cit Sinang and Anas Ghadouani

[100] Krug EC, Frink CR. Acid-rain on acid soil – a new perspective. Science 1983;221:520–525. [101] Likens GE, Driscoll CT, Buso DC. Long-term effects of acid rain: Response and recovery of a forest ecosystem. Science 1996;272:244–6. [102] Schindler DW. Effects of acid-rain on fresh-water ecosystems. Science 1988;239:149–157. [103] Stoddard JL, Jefries DS, Lukewille A, Clair TA, Dillon PJ, Driscoll CT, Forsius M, Johannessen M, Kahl JS, Kellogg JH, Kemp A, Mannio J, Monteith DT, Murdoch PS, Patrick S, Rebsdorf A, Skjelkvale BL, Stainton MP, Traaen T, Van Dam H, Webster KE, Wieting J, Wilander A. Regional trends in aquatic recovery from acidification in North America and Europe. Nature 1999;401:575–578. [104] Basak B, Alagha O. The chemical composition of rainwater over Buyukcekmece Lake, Istanbul. Atmos Res 2004;71:275–88. [105] Persaud AD, Paterson AM, Ingram R, Yao H, Dillon PJ. Potential factors leading to the formation of cyanobacterial scums in a mesotrophic softwater lake in Ontario, Canada. Lake Reservoir Manage 2014;30:331–43. [106] Paerl HW, Xu H, McCarthy MJ, Zhu G, Qin B, Li Y, Gardner WS. Controlling harmful cyanobacterial blooms in a hyper-eutrophic lake (Lake Taihu, China): The need for a dual nutrient (N & P) management strategy. Water Res 2011b;45:1973–83. [107] Chorus I, Bartram J. Toxic Cyanobacteria in Water: A guide to their public health consequences, monitoring and management. E & FN Spon: 1999. [108] Sinang SC, Reichwaldt ES, Ghadouani A. Spatial and temporal variability in the relationship between cyanobacterial biomass and microcystins. Environ Monit Assess 2013;185:6379–95. [109] Sinang SC, Reichwaldt ES, Ghadouani A. Local nutrient regimes determine site-specific environmental triggers of cyanobacterial and microcystin variability in urban lakes. Hydrol Earth Syst Sci Discuss 2014;11:11109–36. [110] Sabart M, Pobel D, Briand E, Combourieu B, Salencon MJ, Humbert JF, Latour D. Spatiotemporal variations in microcystin concentrations and in the proportions of microcystin-producing cells in several Microcystis aeruginosa populations. Appl Environ Microbiol 2010;76:4750–9. [111] Rolland DC, Bourget S, Warren A, Laurion I, Vincent AF. Extreme variability of cyanobacterial blooms in an urban drinking water supply. J Plankton Res 2013;35:744–58. [112] Chorus I, Falconer IR, Salas HJ, Bartram J. Health risks caused by freshwater cyanobacteria in recreational waters. J Toxicol Environ Health Part B 2000;3:323–47. [113] Izydorczyk K, Jurczak T, Wojtal-Frankiewicz A, Skowron A, Mankiewicz-Boczek J, Tarczynska M. Influence of abiotic and biotic factors on microcystin content in Microcystis aeruginosa cells in a eutrophic temperate reservoir. J Plankton Res 2008;30:393–400. [114] De Figueiredo DR, Reboleira ASSP, Antunes SC, Abrantes N, Azeiteiro U, Goncalves F, Pereira MJ. The effect of environmental parameters and cyanobacterial blooms on phytoplankton dynamics of a Portuguese temperate lake. Hydrobiologia 2006;568:145–57. [115] Wilhelm SW, Farnsley SE, Lecleir GR, Layton AC, Satchwell MF, Debruyn JM, Boyer GL, Zhu G, Paerl HW. The relationships between nutrients, cyanobacterial toxins and the microbial community in Taihu (Lake Tai), China. Harmful Algae 2011;10:207–15. [116] Srivastava A, Choi -G-G, Ahn C-Y, Oh H-M, Ravi A, Asthana R. Dynamics of microcystin production and quantification of potentially toxigenic Microcystis sp. using real-time PCR. Water Res 2012;46:817–27. [117] Babanazarova OV, Lyashenko OA. Inferring long-term changes in the physical-chemical environment of the shallow, enriched Lake Nero from statistical and functional analyses of its phytoplankton. J Plankton Res 2007;29:747–56.

7 Global warming, climate patterns and toxic cyanobacteria

243

[118] Cho S, Lim B, Jung J, Kim S, Chae H, Park J, Park S, Park JK. Factors affecting algal blooms in a man-made lake and prediction using an artificial neural network. Measurement 2014;53: 224–33. [119] Smith VH, Tilman GD, Nekola JC. Eutrophication: impacts of excess nutrient inputs on freshwater, marine, and terrestrial ecosystems. Environ Pollut 1999;100:179–96. [120] Smith VH. Low nitrogen to phosphorous ratios favor dominance by blue-green algae in lake phytoplankton. Science 1983;221:669–71. [121] Downing JA, Watson SB, McCauley E. Predicting Cyanobacteria dominance in lakes. Can J Fish Aquat.Sci 2001;58:1905–8. [122] Barros LSS, De Souza FC, Tavares LHS, Amaral LA. Microcystin-LR in Brazilian aquaculture production systems. Water Environ Res 2010;82:240–8. [123] Molot LA, Li GY, Findlay DL, Watson SB. Iron-mediated suppression of bloom-forming cyanobacteria by oxine in a eutrophic lake. Freshw Biol 2010;55:1102–17. [124] Jiang Y, Ji B, Wong RNS, Wong MH. Statistical study on the effects of environmental factors on the growth and microcystins production of bloom-forming cyanobacterium – Microcystis aeruginosa. Harmful Algae 2008;7:127–36. [125] Sharma NK, Mohan D, Rai AK. Predicting phytoplankton growth and dynamics in relation to physico-chemical characteristics of water body. Water Air Soil Pollut 2009;202:325–33. [126] Yan HH, Pan HG, Zou HH, Song HL, Zhang HM. Effects of nitrogen forms on the production of cyanobacterial toxin microcystin-LR by an isolated Microcystis aeruginosa. J Environ Sci Health., Part A 2004;39:2993–3003. [127] Reynolds CS. The ecology of phytoplankton. Cambridge, UK; Cambridge University Press: 2006. [128] Schöne K, Jähnichen S, Ihle T, Ludwig F, Benndorf J. Arriving in better shape: Benthic Microcystis as inoculum for pelagic growth. Harmful Algae 2010;9:494–503. [129] Robarts RD, Zohary T. Temperature effects on photosynthetic capacity, respiration, and growth-rates of bloom-forming cyanobacteria. N Z J Mar Freshwater Res 1987;21:391–9. [130] Hyenstrand P, Blomqvist P, Petersson A. Factors determining cyanobacterial success in aquatic systems – a literature review. In: Forsberg C, Pettersson K [eds.] Advances in Limnology 51 – Lake Erken – 50 Years of Limnological Research. 1998. [131] Markenstein H, Moore K, Persson I. Simulated lake phytoplankton composition shifts toward cyanobacteria dominance in a future warmer climate. Ecol Appl 2010;20:752–67. [132] Wu S, Wang S, Yang H, Xie P, Ni L, Xu J. Field studies on the environmental factors in controlling microcystin production in the subtropical shallow lakes of the Yangtze River. Bull Environ Contam Toxicol 2008;80:329–34. [133] Galvao HM, Reis MP, Valerio E, Domingues RB, Costa C, Lourenco D, Condinho S, Miguel R, Barbosa A, Gago C, Faria N, Paulino S, Pereira P. Cyanobacterial blooms in natural waters in southern Portugal: a water management perspective. Aquat Microb Ecol 2008;53:129–40. [134] Wagner C, Adrian R. Cyanobacteria dominance: Quantifying the effects of climate change. Limnol Oceanogr 2009;54:2460–68. [135] Lehman JT. Nuisance cyanobacteria in an urbanized impoundment: interacting internal phosphorus loading, nitrogen metabolism, and polymixis. Hydrobiologia 2011;661:277–87. [136] Paerl HW, Hall NS, Calandrino ES. Controlling harmful cyanobacterial blooms in a world experiencing anthropogenic and climatic-induced change. Sci Total Environ 2011a;409: 1739–45. [137] Kosten S, Huszar VLM, Becares E, Costa LS, Van Donk E, Hansson LA, Jeppesenk E, Kruk C, Lacerot G, Mazzeo N, De Meester L, Moss B, Lurling M, Noges T, Romo S, Scheffer M. Warmer climates boost cyanobacterial dominance in shallow lakes. Glob Chang Biol 2012;18:118–26.

244

Elke S. Reichwaldt, Som Cit Sinang and Anas Ghadouani

[138] Taranu ZE, Zurawell RW, Pick F, Gregory-Eaves I. Predicting cyanobacterial dynamics in the face of global change: the importance of scale and environmental context. Glob Chang Biol 2012;18:3477–90. [139] Tillett D, Dittmann E, Erhard M, Von Dohren H, Borner T, Neilan BA. Structural organization of microcystin biosynthesis in Microcystis aeruginosa PCC7806: an integrated peptidepolyketide synthetase system. Chem Biol 2000;7:753–64. [140] Tooming-Klunderud A, Rohrlack T, Shalchian-Tabrizi K, Kristensen T, Jakobsen KS. Structural analysis of a non-ribosomal halogenated cyclic peptide and its putative operon from Microcystis: implications for evolution of cyanopeptolins. Microbiology 2007;153:1382–93. [141] Kehr J-C, Picchi DG, Dittmann E. Natural product biosyntheses in cyanobacteria: A treasure trove of unique enzymes. Beilstein J Org Chem 2011;7:1622–35. [142] Hotto AM, Satchwell MF, Berry DL, Gobler CJ, Boyer GL. Spatial and temporal diversity of microcystins and microcystin-producing genotypes in Oneida Lake, NY. Harmful Algae 2008;7:671–81. [143] Sivonen K, Jones G. Toxic cyanobacterial toxins. In: Chorus I, Bartram J [eds.] Toxic cyanobacteria in water: A guide to their public health consequences, monitoring and management. London; E & FN Spon: 1999. [144] Lopes VR, Ramos V, Martins A, Sousa M, Welker M, Antunes A, Vasconcelos VM. Phylogenetic, chemical and morphological diversity of cyanobacteria from Portuguese temperate estuaries. Mar Environ Res 2012;73:7–16. [145] Wood SA, Kuhajek JM, De Winton M, Phillips NR. Species composition and cyanotoxin production in periphyton mats from three lakes of varying trophic status. FEMS Microbiol Ecol 2012;79:312–26. [146] Rico M, Alatmirano M, LÃpez-Rodas V, Costas E. Analysis of polygenic traits of Microcystis aeruginosa (Cyanobacteria) strains by Restricted Maximum Likelihood (REML) procedures: 1. Size shape colonies cells Phycologia 2006;45:237–42. [147] Xu Y, Wang GX, Yang WB, Li RH. Dynamics of the water bloom-forming Microcystis and its relationship with physicochemical factors in Lake Xuanwu (China). Environ Sci Pollut Res 2010;17:1581–90. [148] Briand E, Yepremian C, Humbert JF, Quiblier C. Competition between microcystin- and nonmicrocystin-producing Planktothrix agardhii (cyanobacteria) strains under different environmental conditions. Environ Microbiol 2008b;10:3337–48. [149] Briand E, Gugger M, Francois J-C, Bernard C, Humbert J-F, Quiblier C. Temporal variations in the dynamics of potentially microcystin-producing strains in a bloom-forming Planktothrix agardhii (Cyanobacterium) population. Appl Environ Microbiol 2008a;74:3839–48. [150] El-Shehawy R, Gorokhova E, Fernandez-Pinas F, Del Campo FF. Global warming and hepatotoxin production by cyanobacteria: What can we learn from experiments?. Water Res 2012;46:1420–29. [151] Joung SH, Oh HM, Ko SR, Ahn CY. Correlations between environmental factors and toxic and non-toxic Microcystis dynamics during bloom in Daechung Reservoir, Korea. Harmful Algae 2011;10:188–93. [152] Davis TW, Berry DL, Boyer GL, Gobler CJ. The effects of temperature and nutrients on the growth and dynamics of toxic and non-toxic strains of Microcystis during cyanobacteria blooms. Harmful Algae 2009;8:715–25. [153] Kaebernick M, Neilan BA, Borner T, Dittmann E. Light and the transcriptional response of the microcystin biosynthesis gene cluster. Appl Environ Microbiol 2000;66:3387–92. [154] Renaud SL, Pick FR, Fortin N. Effect of light Intensity on the relative dominance of toxigenic and nontoxigenic strains of Microcystis aeruginosa. Appl Environ Microbiol 2011;77:7016–22.

7 Global warming, climate patterns and toxic cyanobacteria

245

[155] Phelan RR, Downing TG. A growth advantage for microcystin production by Microcystis PCC7806 under high light. J Phycol 2011;47:1241–6. [156] Rinta-Kanto JM, Konopko EA, Debruyn JM, Bourbonniere RA, Boyer GL, Wilhelm SW, Erie L. Microcystis: Relationship between microcystin production, dynamics of genotypes and environmental parameters in a large lake. Harmful Algae 2009;8:665–673. [157] Long BB, Jones BG, Orr BP. Cellular microcystin content in N-limited Microcystis aeruginosa can be predicted from growth rate. Appl Environ Microbiol 2001;67:278–83. [158] Vezie C, Rapala J, Vaitomaa J, Seitsonen J, Sivonen K. Effect of nitrogen and phosphorus on growth of toxic and nontoxic Microcystis strains and on intracellular microcystin concentrations. Microb Ecol 2002;43:443–54. [159] Sevilla E, Martin-Luna B, Vela L, Bes MT, Peleato ML, Fillat MF. Microcystin-LR synthesis as response to nitrogen: transcriptional analysis of the mcyD gene in Microcystis aeruginosa PCC7806. Ecotoxicology 2010;19:1167–73. [160] Engström-Öst J, Repka S, Mikkonen M. Interactions between plankton and cyanobacterium Anabaena with focus on salinity, growth and toxin production. Harmful Algae 2011;10:530–5. [161] Zurawell RW, Chen HR, Burke JM, Prepas EE. Hepatotoxic cyanobacteria: A review of the biological importance of microcystins in freshwater environments. J Toxicol Environ Health Part B 2005;8:1–37. [162] Briand JF, Jacquet S, Flinois C, Avois-Jacquet C, Maisonette C, Leberre B, Humbert JF. Variations in the microcystin production of Planktothrix rubescens (Cyanobacteria) assessed from a four–year survey of Lac du Bourget (France) and from laboratory experiments. Microb Ecol 2005;50:418–28. [163] Rivasseau C, Martins S, Hennion M-C. Determination of some physicochemical parameters of microcystins (cyanobacterial toxins) and trace level analysis in environmental samples using liquid chromatography. J Chromatogr A 1998;799:155–69. [164] Anderson DM, Glibert PM, Burkholder JM. Harmful algal blooms and eutrophication: Nutrient sources, composition, and consequences. Estuaries 2002;25:704–26. [165] Pitois F, Thoraval I, Baures E, Thomas O. Geographical patterns in cyanobacteria distribution: climate influence at regional scale. Toxins 2014;6:509–22. [166] Amé MV, MDP D, Wunderlin DA. Occurrence of toxic cyanobacterial blooms in San Roque Reservoir (Córdoba, Argentina): A field and chemometric study. Environ Toxicol 2003;18: 192–201. [167] Rapala J, Sivonen K, Lyra C, Niemela SI. Variation of microcystins, cyanobacterial hepatotoxins, in Anabaena spp as a function of growth stimuli. Appl Environ Microbiol 1997;63:2206–12. [168] Oh HH, Lee HS, Jang HM, Yoon HB. Microcystin production by Microcystis aeruginosa in a phosphorus-limited chemostat. Appl Environ Microbiol 2000;66:176–79. [169] Takahashi E, Yu Q, Eaglesham G, Connell DW, McBroom J, Costanzo S, Shaw GR. Occurrence and seasonal variations of algal toxins in water, phytoplankton and shellfish from North Stradbroke Island, Queensland, Australia. Mar Environ Res 2007;64:429–42. [170] Li S, Xie P, Xu J, Zhang X, Qin J, Zheng L, Liang G. Factors shaping the pattern of seasonal variations of microcystins in Lake Xingyun, a subtropical plateau lake in China. Bull Environ Contam Toxicol 2007;78:226–30. [171] Wang XX, Parkpian XP, Fujimoto XN, Ruchirawat XK, Delaume XR, Jugsujinda XA. Environmental conditions associating microcystins production to Microcystis aeruginosa in a reservoir of Thailand. J Environ Sci Health A Tox Hazard Subst Environ Eng 2002;37:1181–207. [172] Wu SS, Xie SP, Liang SG, Wang SS, Liang SX. Relationships between microcystins and environmental parameters in 30 subtropical shallow lakes along the Yangtze River, China. Freshw Biol 2006;51:2309–19.

246

Elke S. Reichwaldt, Som Cit Sinang and Anas Ghadouani

[173] Graham JL, Jones JR, Jones SB, Downing JA, Clevenger TE. Environmental factors influencing microcystin distribution and concentration in the Midwestern United States. Water Res 2004;38:4395–404. [174] Van de Waal DB, Smith VH, Declerck SAJ, Stam ECM, Elser JJ. Stoichiometric regulation of phytoplankton toxins. Ecol Lett 2014;17:736–742. [175] Van de Waal DB, Verspagen JMH, Lürling M, Van Donk E, Visser PM, Huisman J. The ecological stoichiometry of toxins produced by harmful cyanobacteria: an experimental test of the carbon-nutrient balance hypothesis. Ecol Lett 2009;12:1326–35. [176] Lukac MM, Aegerter MR. Influence of trace metals on growth and toxin production of Microcystis aeruginosa. Toxicon 1993;31:293–305. [177] Alexova R, Fujii M, Birch D, Cheng J, Waite TD, Ferrari BC, Neilan BA. Iron uptake and toxin synthesis in the bloom–forming Microcystis aeruginosa under iron limitation. Environ Microbiol 2011;13:1064–77. [178] Jiang C, Fan X, Cui G, Zhang Y. Removal of agricultural non-point source pollutants by ditch wetlands: implications for lake eutrophication control. Hydrobiologia 2007;581:319–27. [179] Amano Y, Sakai Y, Sekiya T, Takeya K, Taki K, Machida M. Effect of phosphorus fluctuation caused by river water dilution in eutrophic lake on competition between blue-green alga Microcystis aeruginosa and diatom Cyclotella sp. J Environ Sci-China 2010;22:1666–73. [180] Lang DS, Brown EJ. Phosphorous-limited growth of a green-alga and a blue-green-alga. Appl Environ Microbiol 1981;42:1002–9. [181] De Figueiredo DR, Azeiteiro UM, Esteves SM, Goncalves FJM, Pereira MJ. Microcystinproducing blooms – a serious global public health issue. Ecotoxicol Environ Saf 2004;59: 151–63. [182] Huisman J, Hulot FD. Population dynamic of harmful cyanobacteria. In: Huisman J, Matthijs HCP, Visser PM [eds.] Harmful cyanobacteria. Netherlands; Springer: 2005. [183] Davies JM, Nowlin WH, Matthews B, Mazumder A. Temporal discontinuity of nutrient limitation in plankton communities. Aquat Sci 2010;72:393–402. [184] Boyer GL, Gillam A, Trick C. Iron chelation and uptake. In: Fay P, Baalen CV [eds.] The Cyanobacteria. Netherlands; Elsevier: 1987. [185] Wilhelm S. Ecology of iron-limited cyanobacteria: A review of physiological responses and implications for aquatic systems. Aquat Microb Ecol 1995;9:295–303. [186] Lee W, Van Baalen M, Jansen VAA. An evolutionary mechanism for diversity in siderophoreproducing bacteria. Ecol Lett 2011;15:119–25. [187] Nagai T, Imai A, Matsushige K, Fukushima T. Growth characteristics and growth modeling of Microcystis aeruginosa and Planktothrix agardhii under iron limitation. Limnology 2007;8: 261–70. [188] Ormerod JG. Physiology of the photosynthetic prokaryotes. In: Mann NH, Carr NG [eds.] Photosynthetic prokaryotes. New York; Plenum Press: 1992. [189] Van Liere L, Mur LR. Growth kinetics of Oscillatoria agardhii Gomont in continuous culture, limited in its growth by the light energy supply. J Gen Microbiol 1979;115:153–60. [190] Fogg GE, Thake B. Algal cultures and phytoplankton ecology. London; University of Wisconsin Press: 1987. [191] Xiao Y, Gan N, Liu J, Zheng L, Song L. Heterogeneity of buoyancy in response to light between two buoyant types of cyanobacterium Microcystis. Hydrobiologia 2012;679:297–311. [192] Carey CC, Ibelings BW, Hoffmann EP, Hamilton DP, Brookes JD. Eco-physiological adaptations that favour freshwater cyanobacteria in a changing climate. Water Res 2012;46:1394–407. [193] Schatz D, Keren Y, Vardi A, Sukenik A, Carmeli S, Borner T, Dittmann E, Kaplan A. Towards clarification of the biological role of microcystins, a family of cyanobacterial toxins. Environ Microbiol 2007;9:965–70.

7 Global warming, climate patterns and toxic cyanobacteria

247

[194] Babica P, Blaha L, Marsalek B. Exploring the natural role of microcystins – A review of effects on photoautotrophic organisms. J Phycol 2006;42:9–20. [195] Camacho FA. Macroalgal and cyanobacterial chemical defenses in freshwater communitites. In: Amsler CD [ed.] Algal chemical ecology. Heidelberg, Germany; Springer: 2008. [196] Carmargo JA, Alonso A. Ecological and toxicological effects of inorganic nitrogen pollution in aquatic ecosystems: A global assessment. Environ Int 2006;32:831–49. [197] Guildford SJ, Hecky RE. Total nitrogen, total phosphorus, and nutrient limitation in lakes and oceans: Is there a common relationship?. Limnol Oceanogr 2000;45:1213–23. [198] McQueen DJ, Leam DRS. Influence of water temperature and nitrogen to phosphorus ratios on the dominance of blue-green-algae in Lake St-George, Ontario. Can J Fish Aquat.Sci 1987;44:598–604. [199] An KG, Jones JR. Factors regulating bluegreen dominance in a reservoir directly influenced by the Asian monsoon. Hydrobiologia 2000a;432:37–48. [200] Toth LG, Padisak J. Meteorological factors affecting the bloom of Anabaenopsis-Raciborskii Wolosz (Cyanophyta, Hormogonales) in the shallow Lake Balaton, Hungary. J Plankton Res 1986;8:353–63. [201] Vasconcelos VM. Species composition and dynamics of the phytoplankton in a recently/ commissioned reservoir (Azibo – Portugal). Archiv für Hydrobiologie 1991;121:67–78. [202] Robson BJ, Hamilton DP. Summer flow event induces a cyanobacterial bloom in a seasonal Western Australian estuary. Mar Freshwater Res 2003;54:139–51. [203] Soranno PA. Factors affecting the timing of surface scums and epilimnetic blooms of bluegreen algae in a eutrophic lake. Can J Fish Aquat.Sci 1997;54:1965–75. [204] Jacobsen BA, Simonsen P. Disturbance events affecting phytoplankton biomass, composition and species diversity in a shallow, eutrophic, temperate lake. Hydrobiologia 1993;249:9–14. [205] Tryfon E, Moustakagouni M, Nikolaidis G, Tsekos I. Phytoplankton and physical-chemical features of the shallow Lake Mikri-Prespa, Macedonia, Greece. Archiv für Hydrobiologie 1994;131:477–94. [206] Khondker M, Kabir MA. Phytoplankton primary production in a mesotrophic pond in subtropical Bangladesh. Hydrobiologia 1995;304:39–47. [207] Angeler DG, Alvarez-Cobelas M, Rojo C, Sanchez-Carrillo S. The significance of water inputs to plankton biomass and trophic relationships in a semi-arid freshwater wetland (central Spain). J Plankton Res 2000;22:2075–93. [208] Prepas EE, Charette T. Worldwide eutrophication of water bodies: Causes, concerns, controls. In: Holland HD, Turekian KK [eds.] Treatise on Geochemistry. Elsevier: 2005. [209] Kardinaal WEA, Tonk L, Janse I, Hol S, Slot P, Huisman J, Visser PM. Competition for light between toxic and nontoxic strains of the harmful cyanobacterium Microcystis. Appl Environ Microbiol 2007;73:2939–46. [210] Orr PT, Jones GJ, Douglas GB. Response of cultured Microcystis aeruginosa from the Swan River, Australia, to elevated salt concentration and consequences for bloom and toxin management in estuaries. Mar Freshwater Res 2004;55:277–83. [211] Paerl HW. Nuisance phytoplankton blooms in coastal, estuarine, and inland waters. Limnol Oceanogr 1988;33:823–47. [212] Utkilen H, Gjolme N. Iron-stimulated toxin production in Microcystis aeruginosa. Appl Environ Microbiol 1995;61:797–800. [213] Sevilla E, Martin-Luna B, Vela L, Bes MT, Fillat MF, Peleato ML. Iron availability affects mcyD expression and microcystin-LR synthesis in Microcystis aeruginosa PCC7806. Environ Microbiol 2008;10:2476–83. [214] Tonk L, Bosch K, Visser PM, Huisman J. Salt tolerance of the harmful cyanobacterium Microcystis aeruginosa. Aquat Microb Ecol 2007;46:117–23.

248

Elke S. Reichwaldt, Som Cit Sinang and Anas Ghadouani

[215] Lürling M, Eshetu F, Faassen EJ, Kosten S, Huszar VLM. Comparison of cyanobacterial and green algal growth rates at different temperatures. Freshw Biol 2013;58:552–9. [216] Padisak J. Cylindrospermopsis raciborskii (Woloszynska) Seenayya et Subba Raju, an expanding, highly adaptive cyanobacterium: worldwide distribution and review of its ecology. Archiv für Hydrobiologie Supplement 1997;107:563–93. [217] Briand JF, Leboulanger C, Humbert JF, Bernard C, Dufour P. Cylindrospermopsis raciborskii (Cyanobacteria) invasion at mid-latitudes: Selection, wide physiological tolerance, or global warming?. J Phycol 2004;40:231–8. [218] Paerl HW, Huisman J. Climate change: a catalyst for global expansion of harmful cyanobacterial blooms. Environ Microbiol Rep 2009;1:27–37. [219] Huisman J, Sharples J, Stroom JM, Visser PM, Kardinaal WEA, Verspagen JMH, Sommeijer B. Changes in turbulent mixing shift competition for light between phytoplankton species. Ecology 2004;85:2960–70. [220] Havens KE. Chapter 33: Cyanobacteria blooms: effects on aquatic ecosystems. In: Hudnell HK [ed.] Proceedings of the Interagency, International Symposium on Cyanobacterial Harmful Algal Blooms (ISOC-HAB): State of the Science and Research Needs. Springer: 2007. [221] Reynolds CS, Wiseman SW, Clarke MJO. Growth-rate and loss-rate responses of phytoplankton to intermittent artificial mixing and their potential application to the control of planktonic algal biomass. J Appl Ecol 1984;21:11–39. [222] Magnuson JJ, Webster KE, Assel RA, Bowser CJ, Dillon PJ, Eaton JG, Evans HE, Fee EJ, Hall RI, Mortsch LR, Schindler DW, Quinn FH. Potential effects of climate changes on aquatic systems: Laurentian great lakes and precambrian shield region. Hydrol Process 1997;11: 825–71. [223] Moisander PH, Ochiai M, Lincoff A. Nutrient limitation of Microcystis aeruginosa in northern California Klamath River reservoirs. Harmful Algae 2009;8:889–97. [224] Morales-Baquero R, Pulido-Villena E, Recher I. Atmospheric inputs of phosphorus and nitrogen to the southwest Mediterranean region: Biogeochemical responses of high mountain lakes. Limnol Oceanogr 2006;51:830–7. [225] Lehmann EM, McDonald KE, Lehmann JT. Whole lake selective withdrawal experiment to control harmful cyanobacteria in an urban impoundment. Water Res 2009;43:1187–98. [226] Elser JJ, Marzolf ER, Goldman CR. Phosphorus and nitrogen limitation of phytoplankton growth in the fresh-waters of North-America – a review and critique of experiemental enrichments. Can J Fish Aquat.Sci 1990;47:1468–77. [227] Zhang M, Xu J, Xie P. Nitrogen dynamics in large shallow eutrophic Lake Chaohu, China. Environ Geol 2008;55:1–8. [228] Davis TW, Harke MJ, Marcoval MA, Goleski J, Orano-Dawson C, Berry DL, Gobler CJ. Effects of nitrogenous compounds and phosphorus on the growth of toxic and non-toxic strains of Microcystis during cyanobacterial blooms. Aquat Microb Ecol 2010;61:149–62. [229] Downing JA, McClain M, Twilley R, Melack JM, Elser J, Rabalais NN, Lewis WM, Turner RE, Corredor J, Soto D, Yanez-Arancibia A, Kopaska JA, Howarth RW. The impact of accelerating land-use change on the N-cycle of tropical aquatic ecosystems: Current conditions and projected changes. Biogeochemistry 1999;46:109–48. [230] Zhai SJ, Yang LY, Hu WP. Observations of atmospheric nitrogen and phosphorus deposition during the period of algal bloom formation in northern Lake Taihu, China. Environ Manage 2009;44:542–51. [231] Agawin NSR, Rabouille S, Veldhuis MJW, Servatius L, Hol S, Van Overzee HMJ, Huisman J. Competition and facilitation between unicellular nitrogen-fixing cyanobacteria and nonnitrogen-fixing phytoplankton species. Limnol Oceanogr 2007;52:2233–48. [232] WHO. Guidelines for safe recreational water environments. Geneva; WHO: 2003.

7 Global warming, climate patterns and toxic cyanobacteria

249

[233] Graham JL, Jones JR. Microcystin in Missouri reservoirs. Lake Reservoir Manage 2009;25: 253–63. [234] Hesse K, Kohl JG. Effects of light and nutrient supply on growth and microcystin content of different strains of Microcystis aeruginosa. In: Chorus I [ed.] Cyanotoxins: Occurrence Causes Consequences. Berlin; Springer: 2001. [235] ANZECC 2000. Australian and New Zealand Guidelines for Fresh and Marine Water Quality, Australian and New Zealand Environment and Conservation Council and Agriculture and Resource Management Council of Australia and New Zealand. [236] Neilan BA, Pearson LA, Moffitt MC, Mihali KT, Kaebemick M, Kellmann R, Pomati F. Chapter 17: The genetics and genomics of cyanobacterial toxicity. In: Hudnell HK [ed.] Proceedings of the Interagency, International Symposium on Cyanobacterial Harmful Algal Blooms (ISOC-HAB): State of the Science and Research Needs. New York, USA; Springer: 2008. [237] Blackburn SI, Bolch CJ, Jones GJ, Negri AP, Orr PT. Cyanobacterial blooms: why are they toxic?. In: Davis JRD [ed.] Managing Algal Blooms: Outcomes from the CSIRO Blue-Green Algal Research Program. Canberra; CSIRO Land and Water: 1997. [238] Song H, Reichwaldt ES, Ghadouani A. Contribution of sediments in the removal of microcystin-LR from water. Toxicon 2014;83:84–90. [239] Cousins IT, Bealing DJ, James HA, Sutton A. Biodegradation of microcystin-LR by indigenous mixed bacterial populations. Water Res 1996;30:481–5. [240] Chen W, Song LR, Peng L, Wan N, Zhang XM, Gan NQ. Reduction in microcystin concentrations in large and shallow lakes: Water and sediment-interface contributions. Water Res 2008;42:763–73. [241] Rapala J, Lahti K, Sivonen K, Niemala SI. Biodegradability and adsorption on lake-sediments of cyanobacterial hepatotoxins and anatoxin-a. Lett Appl Microbiol 1994;19:423–8. [242] Tsuji K, Masui H, Uemura H, Mori Y, Harada K. Analysis of microcystins in sediments using MMPB method. Toxicon 2001;39:687–92. [243] Harada K, Tsuji K. Persistence and decomposition of hepatotoxic microcystins produced by cyanobacteria in natural environment. J Toxicol Toxin Rev 1998;17:385–403. [244] Park HD, Sasaki Y, Maruyama T, Yanagisawa E, Hiraishi A, Kato K. Degradation of the cyanobacterial hepatotoxin microcystin by a new bacterium isolated from a hypertrophic lake. Environ Toxicol 2001;16:337–43. [245] Wu XQ, Wang CB, Xiao BD, Wang Y, Zheng N, Liu JS. Optimal strategies for determination of free/extractable and total microcystins in lake sediment. Anal Chim Acta 2012;709:66–72. [246] Wu XQ, Xiao BD, Li RH, Wang CB, Huang JT, Wang Z. Mechanisms and factors affecting sorption of microcystins onto natural sediments. Environ Sci Technol 2011;45:2641–7. [247] Francis G. Poisonous Australian lake. Nature 1878;18:11–12. [248] Otten T, Paerl H. Health effects of toxic cyanobacteria in U.S. drinking and recreational waters: our current understanding and proposed direction. Curr Environ Health Rep 2015: 1–10. [249] WHO. Cyanobacterial toxins: microcystin-LR. In: Guidelines for drinking water quality. Geneva, Switzerland; World Health Organization: 1998. 95–110. [250] Falconer IR. Toxic cyanobacterial bloom problems in Australian waters: risks and impacts on human health. Phycologia 2001;40:228–33. [251] World Water Coucil, The Co-operative Programme on Water and Climate & the International Water Association 2009. Introduction, summaries and key messages. Perspectives on water and climate change adaptation. World Water Council, Cooperative Program on Water and Climate, International Water Association, IUCN.

250

Elke S. Reichwaldt, Som Cit Sinang and Anas Ghadouani

[252] Burke J, Kuylenstierna J The Water Variable – Producing enough food in a climate insecure world. Perspectives on water and climate change adaptation. World Water Council, Cooperative Program on Water and Climate, International Water Association, IUCN: 2009. [253] Zwolsman G, Vanham D, Fleming P, Davis C, Lovell A, Nolasco D, Thorne O, De Sutter R, Fülöp B, Staufer P, Johannessen A Climate change and the water industry – Practical responses and actions. Perspectives on water and climate change adaptation. World Water Council, Cooperative Program on Water and Climate, International Water Association, IUCN: 2009. [254] Batchelor C, Schouten T, Smits S, Moriarty P, Butterworth J Climate change and WASH services delivery – Is improved WASH governance the key to effective mitigation and adaptation? Perspectives on water and climate change adaptation. World Water Council, Cooperative Program on Water and Climate, International Water Association, IUCN: 2009. [255] Kundzewicz ZW, Mata LJ, Arnell NW, Döll P, Kabat P, Jiménez B, Miller KA, Oki T, Sen Z, Shiklomanov IA. Freshwater resources and their management. In: Parry ML, Canziani OF, Palutikof JP, Van der Linden PJ, Hanson CE [eds.] Climate Change 2007: Impacts, Adaptation and Vulnerability. Contribution of Working Group II to the Fourth Assessment Report of the Intergovernmental Panel on Climate Change. Cambridge, UK; Cambridge University Press: 2007. [256] Yen HK, Lin TF, Tseng IC, Su YT. Cyanobacteria toxins and toxin producers in nine drinking water reservoirs in Taiwan. Drinking Water Treatment, Supply and Management in Asia (IWAASPIRE 2005) 2006;6:161–7. [257] Liu YM, Chen W, Li DH, Huang ZB, Shen YW, Liu YD. Cyanobacteria-/cyanotoxin/ contaminations and eutrophication status before Wuxi Drinking Water Crisis in Lake Taihu, China. J Environ Sci-China 2011;23:575–81. [258] Lawton LA, Welgamage A, Manage PM, Edwards C. Novel bacterial strains for the removal of microcystins from drinking water. Water Sci Technol 2011;63:1137–42. [259] Barrington DJ, Ghadouani A. Application of hydrogen peroxide for the removal of toxic cyanobacteria and other phytoplankton from wastewater. Environ Sci Technol 2008;42: 8916–21. [260] Barrington DJ, Ghadouani A, Ivey GN. Environmental factors and the application of hydrogen peroxide for the removal of toxic cyanobacteria from waste stabilization ponds. J Environ Eng 2011:137. Doi: 10.1061/(ASCE)EE.1943-7870.0000401 [261] Barrington DJ, Ghadouani A, Ivey GN. Cyanobacterial and microcystins dynamics following the application of hydrogen peroxide to waste stabilisation ponds. Hydrol Earth Syst Sci 2013;17:2097–2105. [262] Kanoshina I, Lips U, Leppanen JM. The influence of weather conditions (temperature and wind) on cyanobacterial bloom development in the Gulf of Finland (Baltic Sea). Harmful Algae 2003;2:29–41. [263] Sipiä VO, Kankaanpaa HT, Flinkman J, Lahti K, Merilouto JAO. Time-dependent accumulation of cyanobacterial hepatotoxins in flounders (Platichthys flesus) and mussels (Mytilus edulis) from the northern Baltic Sea. Environ Toxicol 2001;16:330–6. [264] Hallegraeff GM. Harmful algal blooms in the Australian region. Mar Pollut Bull 1992;25: 186–90. [265] Hallegraeff GM. A review of harmful algal blooms and their apparent global increase. Phycologia 1993;32:79–99. [266] Shumway SE. A review of the effects of algal blooms on shellfish and aquaculture. J World Aquacult Soc 1990;21:65–104. [267] Costanza R, Darge R, Degroot R, Farber S, Grasso M, Hannon B, Limburg K, Naeem S, Oneill RV, Paruelo J, Raskin RG, Sutton P, Van den Belt M. The value of the world’s ecosystem services and natural capital. Nature 1997;387:253–60.

7 Global warming, climate patterns and toxic cyanobacteria

251

[268] Carvalho L, McDonald C, De Hoyos C, Mischke U, Phillips G, Borics G, Poikane S, Skjelbred B, Solheim AL, Van Wichelen J, Cardoso AC. Sustaining recreational quality of European lakes: minimizing the health risks from algal blooms through phosphorus control. J Appl Ecol 2013;50:315–23. [269] Golterman HL. Phosphate release from anoxic sediments or `What did Mortimer really write?’. Hydrobiologia 2001;450:99–106.

Panagiota Katikou

8 Human impact in Mediterranean coastal ecosystems and climate change: emerging toxins 8.1 Introduction An increase in the occurrence of harmful or toxic algal incidents, in both frequency and geographic distribution, in many parts of the world is being observed over the past decades [1–3]. The predicted changes in the oceans have both direct and indirect impact on interactions between humans and the oceans and, over the years, numerous studies have suggested possible relationships between climate and the magnitude, frequency and duration of harmful algal blooms (HABs) [4]. Marine and freshwater ecosystems are warming, acidifying and deoxygenating as a consequence of climate change, while, in parallel, the impacts of HABs on these ecosystems are intensifying. Many eutrophic habitats that host recurring HABs already experience thermal extremes, low dissolved oxygen and low pH, making these locations potential sentinel sites for conditions that will become more common in larger scale systems as climate change accelerates. In fact, the relationship between HABs and climate change is bidirectional, with the former actually considered as one of the important climate change costressors (Figure 8.1) [5]. Dense phytoplankton blooms develop in response to favorable conditions for cell growth and accumulation [6]. These blooms of autotrophic algae and some heterotrophic protists are increasingly frequent in coastal waters around the world. Undoubtedly, HABs are occurring in more locations than ever before and new sightings are regularly reported. Certain researchers argue that this trend is due to increasing eutrophication throughout the world [7] but, generally, phytoplankton blooms have regional, seasonal and species-specific aspects that should be considered [8–10]. Recent research indicates that although both water temperature and eutrophication are critical factors affecting phytoplankton production, the combined effects of warming and nutrient changes on phytoplankton production in coastal waters are not yet well understood [11]. Under nonenriched experimental conditions, the effect of warming on phytoplankton production was significantly positive in some months, significantly negative in others or had no effect. On the other hand, under enriched conditions, warming affected phytoplankton production positively in all months except one, when the salinity was as low as 6.5 practical salinity units (psu). It is thus

Panagiota Katikou, Directorate of Research, Innovation and Education, Ministry of Rural Development and Food, Hapsa & Karatasou 1, Thessaloniki, Greece https://doi.org/10.1515/9783110625738-008

Panagiota Katikou

Aquaculture

Warming, increased stratification

Nutrients

HABs Increased organic matter

Increased respiration

CO2

O2 Deoxygenation

Atmospheric CO2

Global temperatures

Acidification

254

Hypoxia, acidification

HABs, vertical migration

Benthos and submerged aquatic vegetation

Figure 8.1: The co-occurrence of climate change stressors and HABs in coastal ecosystems. Figure depicts how successive increases in temperature, carbon dioxide and nutrient loading since the twentieth century independently and interactively promote HABs. Reprinted from Harmful Algae, vol. 91, 101590, 2020. DOI: https://doi.org/10.1016/j.hal.2019.03.008, Griffith AW, Gobler CJ, Harmful algal blooms: A climate change costressor in marine and freshwater ecosystems, Copyright (2019), with permission from Elsevier.

suggested that nutrient conditions can alter the effects of warming on phytoplankton production [11], further supporting the bidirectional character of the relationships between all costressors involved in the development of climate change [5]. Generally, drivers of HAB expansion are grouped into three broad classes: (i) climatic, (ii) anthropogenic and (iii) hydrodynamic [12]. Τhere is an increasing concern that accelerated environmental change attributed to human-induced warming of the planet may substantially alter the patterns, distribution and intensity of HABs. Changes in temperature, ocean acidification, precipitation, nutrient stress or availability and the physical structure of the water column all influence the productivity, composition and global range of phytoplankton assemblages, but large uncertainty remains about how integration of these climate drivers might shape future HABs [13]. On the other hand, nonnutrient pollutants including oil/hydrocarbon and its congeners should also be recognized as potential triggers of HABs.

8 Human impact in Mediterranean coastal ecosystems and climate change

255

Currently, most efforts for HAB control are heavily focused on N&P input reduction, while it is recognized that a transition from nutrient-related to nonnutrient-related triggers of HABs in future may be possible if both factors do not simultaneously attract adequate attention [12]. In contrast to large-scale blooms that are dominated by mesoscale circulation, Mediterranean HABs are a more localized phenomenon commonly present in areas of constrained dynamism, such as bays, lagoons, ports, beaches and estuaries. In such areas, enhanced growth of phytoplankton can cause an observable water discoloration along the shoreline as well as deterioration in water quality. In recent years, certain unprecedented ecological effects on the Mediterranean area, such as fish kills and risks to human health, have also been attributed to toxic algal proliferations. Taking into account that a bloom represents a deviation from the normal biomass cycle and despite the fact that in some cases algae proliferation may be of a natural origin, coastal blooms are considered as an emerging problem possibly related to nutrient enrichment of coastal waters. Furthermore, intensive urbanization and recreational use of coastal watersheds has led to a remarkable increase in nutrient sources along the Mediterranean coasts. This cultural eutrophication creates a contrast between coastal waters and the open ocean where, due to summer stratification and nutrient depletion, oligotrophic conditions prevail in the upper layer. Nutrient-rich coastal environments of the Mediterranean Sea and, in particular, semienclosed areas with low turbulence levels constitute a new and unique environment in which a number of phytoplankton species with harmful effects can become dominant [14, 15] and, in fact, the increased eutrophication index in urban Mediterranean marine environments was found to be directly linked to persistent phytoplankton blooms and conspicuous red tides [16]. Despite the fact that most of the factors implicated in the Mediterranean-nearshore algal outbreaks are known, the mechanisms underpinning their occurrence are not yet well established [17]. Along North African coasts, the spatial distribution of chlorophylls and carotenoids is attributed to the human-altered patterns of the physical structure and the nutrient concentrations, as well as to the Modified Atlantic Water (MAW). The physical forcing resulting from advection of the MAW could confront distinct water masses and result in potential mixing of water from coastal and/ or open-ocean origin. This water mixing may affect phytoplankton populations, which, in North Africa, present large variations in terms of abundance, composition and size structure due to the dynamic nature of their environment [18]. In the Black Sea, since the late 1970s, anthropogenic nutrient enrichment has been recognized as a key ecological problem for this basin, especially its northwestern and western parts, which are mostly subjected to the influence of freshwater nutrient inputs. Nutrient input and dissolved organic matter in the northwest shelf of the Black Sea by the Danube, the Dnieper and the Dniester rivers have showed an increase of about 10 times in the period 1950–1980. A rise in phytoplankton blooms frequency, involved species, duration, timing and area has been documented, provoking substantial perturbations of the entire food web structure and functioning. Changes in

256

Panagiota Katikou

zooplankton communities structure and deterioration of benthic coenoses, culminating during the 1980s (period of intensive eutrophication in the Black Sea), were to a large extent connected with the dramatic alterations in phytoplankton communities and recurrent hypoxic conditions. Microalgal blooms were therefore recognized as one of the major issues for the Black Sea’s ecological health [8]. Long-term (15 years) observations in the northeast Black Sea indicate that the seasonal dynamic of dominant species changes as a function of N, P, Si concentrations and of their stoichiometry. In turn, nutrient concentrations depend on physical factors (circulation of water masses), which are largely determined by wind speed and direction [19]. Similar eutrophication problems have been identified in the Eastern Mediterranean Sea in several Aegean and Ionian coastal areas, affected by urban and industrial wastewaters and/or nutrient inputs from rivers and agricultural activities [16]. In this context, phytoplankton, as primary producer, became the first target of the anthropogenic-induced stress, resulting in dramatic alterations in species composition, abundance and biomass, seasonal dynamics and succession in the two basins [8, 10].

8.2 Mediterranean coastal ecosystems The Mediterranean Sea constitutes the crucial environmental factor of the Mediterranean region. The presence of a large marginal and almost completely closed sea on the western side of a large continental area is geographically unique. Its size is substantial, with an area, excluding the Black Sea, of about 2.5 million km2, an extent of about 3,700 km in longitude and 1,600 km in latitude and an average depth of 1,500 m. The Strait of Gibraltar, connecting the Mediterranean Sea to the Atlantic, has a width of only 14.5 km and a depth of less than 300 m at the shallowest sill. These morphological characteristics make the Mediterranean Sea a large source of moisture and a heat reservoir with a significant capacity for the surrounding land areas (considering the annual average, it acts as a moderate source of heat). The Strait of Gibraltar has a particular role in the Mediterranean Sea environment. The fluxes through the strait compensate for the mass deficit due to the large evaporation in the basin, supply comparatively freshwater masses to one of the saltiest seas on the Earth, and also provide a relatively small supply of heat, because the Mediterranean water (MW) outflow is cooler than the Atlantic water inflow. The region’s complicated morphology – with the presence of many sharp orographic features, often close to the coastlines, and of distinct basins and gulfs, islands, and peninsulas – has a strong influence on atmospheric circulation. Moreover, it is responsible for several cyclogenetic areas, local winds, many mesoscale processes and intense air–sea interactions, such as those responsible for dense-water formation processes driving the Mediterranean thermohaline cells. Furthermore, the shape of

8 Human impact in Mediterranean coastal ecosystems and climate change

257

the Mediterranean Sea bottom, with deep basins linked through much shallower straits, strongly constrains the Mediterranean Sea circulation [20]. Despite the fact that the Mediterranean is a semienclosed sea with a relatively small area compared to other large marine ecosystems worldwide, its waters and coasts are characterized by a disproportionately huge diversity from the environmental and socioeconomical points of view. Three continents (Europe, Asia, Africa) and 21 countries surround this basin, a number which on its own indicates the great variety in human culture and socioeconomic development found in Mediterranean coastal zones that have been inhabited for millennia. Marine environmental and ecological conditions as well as habitats present a great variation: a wide range of depths is covered, and primary productivity displays a clear west–east and north–south decrease, similarly to fishery landings, a spatial pattern opposite to that of temperature and salinity [21–24]. The natural balance of Mediterranean coast ecosystems is severely affected by major disturbances, which could cause an extensive loss of biodiversity. Such disturbances include habitat destruction and alien species introductions, overfishing and pollution; these phenomena are tightly related to the increase of anthropogenic pressure due to urbanization on the northwestern shores, on the one hand and outstanding growth population on the southern and eastern shores on the other. The Mediterranean basin is therefore of major interest for studying the climate change effects [25]. In fact, the Mediterranean Sea is identified as an important anthropogenic carbon pool where the column inventory is much higher than in the Atlantic [26, 27]. This is due to its intrinsic physicochemical characteristics, in which warm and highly alkaline waters are prone to absorb high amounts of CO2 from the atmosphere and transport it to deep waters via a number of convective areas [28]. Conversely, the ongoing increase of water temperature is influencing the Mediterranean marine biota by a combination of direct (survival, reproduction, recruitment, etc.) and indirect effects, mediated by biotic interactions (predation, parasitism, diseases, etc.) or the altered emphasis of marine current fluxes between adjacent basins. Species distribution within the Mediterranean Sea is modifying in response to climate change, with tropical taxa taking over communities, thus triggering – together with sea water warming – a process of “tropicalization” [29, and references thereof]. Warming of the Mediterranean Sea surface is currently estimated at 0.4 °C per decade for the period between 1985 and 2006 (+0.3 °C per decade for the western basin and +0.5 °C per decade for the eastern basin). The increases in temperature are not constant throughout the year but occur primarily during May, June and July. Maximum increases of 0.16 °C per year were found in June in the Tyrrhenian, Ligurian and Adriatic Seas and close to the African coast. The Aegean Sea shows maximum change in sea surface temperature during August. The projections for 2100 vary between +1.8 and +3.5 °C in average compared to the period between 1961 and 1990. The Balearic Islands, the northwest Ionian, the Aegean and Levantine Seas have been identified as the regions with maximum increase of sea surface temperature [30, and references thereof].

258

Panagiota Katikou

Similar to worldwide trends caused by warming and loss of glacial ice, sea level in the Mediterranean has risen between 1945 and 2000 at a rate of 0.7 mm per year and between 1970 and 2006 at the level of 1.1 mm per year. There has been a sharp increase during the last two decades as sea level rise reached about 3 mm per year. Moreover, ocean pH has decreased by 0.1 pH units since the preindustrial period, which is unprecedented during at least the last 65 million years. Globally, CO₂ uptake by the oceans is expected to lead, by 2100, to acidification of 0.15–0.41 pH units below 1870–1899 levels. Similar rates are expected for the Mediterranean, which is currently estimated to decrease by 0.018 to 0.028 pH units per decade [30, and references thereof]. In addition, local anthropogenic stressors, such as habitat degradation or destruction, pollution, sedimentation and overfishing, together with a gradual history of changes in environmental conditions (climate, habitat quality, resource availability), can potentially exacerbate biodiversity decline and habitat loss, thus eroding the ability of ecosystems to absorb and recover from additive or synergic affecters [29, and references thereof]. Indeed, a large number of coastal areas in the Mediterranean Sea are eutrophic, a situation more intensively present in semienclosed areas [31]. The use of satellite imagery on chlorophyll distribution has shown that the highest concentrations are located close to river deltas and estuaries or near urban agglomerations, especially in the estuary of Nile (from Alexandria to Gaza), Gulfs of Antalya and Alexandretta (Turkey), Northern Aegean, Thermaikos Gulf (Greece), the Adriatic Sea, the Gulf of Lions (France), Valencia-Barcelona (Spain) and the Gulf of Gabes (Tunisia). Increased temperatures associated with eutrophication can enhance the occurrence of HABs and can negatively impact aquaculture production (especially filter feeders farming), and increase human health risks. In fact, human consumption of seafood contaminated with harmful biotoxins can result in a variety of commonly known intoxications with varying degrees of severity, including amnesic shellfish poisoning (ASP), shellfish poisoning from lipophilic toxins (such as diarrhetic shellfish poisoning, DSP) and paralytic shellfish poisoning (PSP), whereas the increased presence of new/emerging marine biotoxins of unknown human toxicity further complicates the problem. Marine biotoxins primarily target the nervous or digestive systems, but can also result in potentially fatal acute respiratory distress and other chronic neurological and immunological illnesses. The danger of inadvertently consuming biotoxins is compounded by the fact that they are odorless and tasteless and are unaffected by food preparation procedures [32].

8.2.1 Human impact A worldwide proliferation of HABs (toxic, food web altering, hypoxia-generating) has been linked to human nutrient [phosphorus (P) and nitrogen (N)] overenrichment. Human activities are probably the strongest drivers of change in marine

8 Human impact in Mediterranean coastal ecosystems and climate change

259

biodiversity at all levels of organization; hence, future trends will depend largely on human-related threats [33]. Most human activities present a local impact, whereas others and, above all, the overall total acting synergistically may have global impact through cumulative processes. Building of the Aswan High Dam in 1968 may not only have deleteriously affected the productivity, biochemistry and food web structure in the Nile delta and Eastern Mediterranean, but also the hydrological functioning and structure of the Mediterranean as a whole, which itself will influence the chemical and biological characteristics in a feedback loop [34–35]. Similarly, one of the most important anthropogenic effects was the opening of the Suez Canal in Egypt in November 1869, which was the start of a massive invasion of hundreds of marine alien species from Indo-Pacific origin [36–39]. With the opening of the Suez Canal in 1869, two markedly different zoogeographical areas were joined: the subtropical Mediterranean Sea, which connects with the Atlantic, and the tropical Red Sea, the most northern extension of the Indian Ocean. In order to pass between these areas, organisms must be able to bridge the difference in adaptive requirements, and also withstand the extreme conditions in the Canal itself [40]. In this context, the term ‘‘Lessepsian migration’’ named after Ferdinand de Lesseps, the French diplomat in charge of the canal’s construction, has been introduced to characterize a new phenomenon of unidirectional and successful biotic advance from the Red Sea to the Eastern Mediterranean (Figure 8.2), while the term ‘‘Lessepsian migrant’’ refers to the Red Sea species that have passed through the Suez Canal and settled in the Eastern Mediterranean [41]. The Lessepsian migrant Lagocephalus sceleratus (Gmelin, 1789), also known as silverstripe blaasop, is an Indo-Pacific-originated puffer fish of the family Tetraodontidae, which is now very well established in the Eastern Mediterranean. Similar to its congeneric tropical species, L. sceleratus maybe a source of food poisoning with a high-associated risk of mortality, as it commonly contains tetrodotoxin (TTX), a toxin that can cause death by muscular paralysis, respiratory depression and circulatory failure [42–44]. The presence of this alien species and its associated toxin (TTX) in the Mediterranean Sea in the last 10–15 years, therefore, constitutes a new and highly important emerging risk that has to be considered by the authorities of the affected countries. Furthermore, around half of all the extra CO2 produced so far by human activities has dissolved in the oceans. The anthropogenic increase in atmospheric carbon dioxide concentration results in a reduction of seawater pH on a global scale (a process termed as “ocean acidification”), changing the chemistry and biogeochemical cycling of carbon and carbonate. A growing number of studies have demonstrated the adverse effects of acidification on marine organisms. The combination of elevated temperature and acidification has been proved detrimental to the calcification process; hence, marine organisms with calcareous skeletons, shells or plates are expected to experience problems. These include major bioconstructing organisms, such as scleractinian

260

Panagiota Katikou

Figure 8.2: Geography of the Mediterranean Sea with the main routes of species range expansion. Bold capital abbreviations correspond to the main Mediterranean subregions (ALB: Alboran Sea; NWM: North Western Mediterranean; TYR: Tyrrhenian Sea; ADR: Adriatic Sea; ION: Ionian Sea; AEG: Aegean Sea; LEV: Levantine Basin) and adjacent seas (ATL: Atlantic Ocean; BLA: Black Sea; RED: Red Sea). Italic abbreviations correspond to some remarkable Mediterranean locations (Gib: Gibraltar Straits; GoL: Gulf of Lions; Sue: Suez Canal). Reported temperatures correspond to winter–summer mean sea-surface temperatures. Arrows represent main routes of species range expansion according to their origin: Mediterranean natives (orange), Atlantic migrants (green) and Lessepsian migrants (red). Reprinted from Trends in Ecology & Evolution, Vol 25(4), Lejeusne C, Chevaldonné P, PergentMartini C, Boudouresque CF, Pérez T, Climate change effects on a miniature ocean: The highly diverse, highly impacted Mediterranean Sea, Pages 250–260, Copyright (2010), with permission from Elsevier.

corals, bryozoans or red coralline algae, but also mollusks, crustaceans, echinoderms, foraminifera and some calcifying phytoplankton (mainly coccolithophorides) as well as zooplankton (pteropods and larvae of many groups) [35]. On the other hand, several planktonic organisms are affected by acidification with possible negative impacts on fish populations, whereas acidification also threatens iconic and invaluable Mediterranean ecosystem-building species (such as sea grass meadows, coralligene reefs and vermetid snail reefs), which create rich key habitats and homes to thousands of species, and also protect shores from erosion as well as offer a source of food and natural products to society [45]. Mediterranean coastal ecosystems were and are still undergoing heavy human-derived impacts [46]. Large concentrations of contaminants have accumulated in sediments of coastal lagoons due to past industrial activity, thus one of the important issues to be faced is the fate of these contaminants with changing environmental conditions, including those related to climate change [25, 47].

8 Human impact in Mediterranean coastal ecosystems and climate change

261

8.2.2 Socioeconomical implications of climate change Coastal marine systems are among the most ecologically and socioeconomically vital at a global level. There is a strong scientific consensus for the fact that these marine ecosystems, along with the resources and services they provide, are significantly threatened by anthropogenic climate change [48, 49], with HABs and invasive marine species being within the major issues creating problems, especially in terms of presence of emerging toxins. The HAB designation is mostly a societal concept rather than a scientific definition –blooms are considered to fit the HAB criterion if they cause injury to human health or socioeconomic interests, or to components of aquatic ecosystems. Some HAB species are toxigenic and produce blooms that cause illness and death of fish, seabirds, mammals and other marine life, often via toxin transfer through the food web. Human consumers of seafood contaminated by these toxins may also be poisoned, suffering acute toxic symptoms and even fatalities in extreme cases. Further, toxic threats to human health are posed by toxic aerosols and waterborne compounds that cause respiratory and skin irritation when released from toxic cells [50]. The economic effects of HABs arise from public health costs including morbidities and mortalities, commercial fishery closures and fish kills, declines in coastal and marine recreation and tourism as well as the costs of monitoring and management. Aggregating economic effects both within and across these categories can be problematic, as the measures of effects are rarely the changes in economic surpluses sought by economists [51]. A rough estimate of the economic effects of HABs in the United States is $100 million per year (at the 2012 value of the dollar). Anderson et al. [52] estimated back in 2000 the proportional breakdown of costs related to HAB impacts to be 45% for public health costs, 37% in terms of the costs of closures and losses experienced by commercial fisheries, 13% to the impact on lost recreation and tourism, and 4% to monitoring and management costs. Where the European Union (EU) is concerned, the socioeconomic impact of HABs for three evaluated Mediterranean countries – Italy, Greece and France – had been estimated at around 329 million euro per year in 2006 [53]; however, approximately two-thirds of this had been associated with the noxious, but nontoxic, effects of macroalgal (and some microalgal, e.g., Phaeocystis) blooms affecting the human uses of the coast [54, 55]. A recent report of the EU Joint Research Centre indicated a general shortage of data on the economic effects of HABs, with their socioeconomic effects grouped in four main impacts: (1) human health; (2) fishery; (3) tourism and recreation; and (4) monitoring and management costs (Figure 8.3), with the monitoring and management ones ranging from $300,000 to more than $1,000,000 in Norway, Denmark, Portugal, France and Spain (Galicia), at an annual basis [56].

262

Panagiota Katikou

Human health impacts

Medical expenses Hospitalization expenses Costs of transportation to hospitals Lost of productivity (lost wages and work days)

Fish mortality Fishing closures applied to commercial fishers Increase of fish price Reduction in consumer demand of fish and fish products

Commercial fishery impacts

Economic impacts of HABs

Water sampling

Monitoring and management impacts

Water treatments to remove toxins

Tourism/ recreation impacts

Fishing closures applied to recreational fishers Economic damage to the tourist industry Reduction in recreational experiences of visitors near the beaches

Actions to identify factor causing blooms Strategies adopted to destroy HAB

Figure 8.3: Economic impact of harmful algal blooms (HABs). Figure depicts the four sectors affected by HAB episodes and lists, for each of them, the main causes of economic losses. Reprinted from Sanseverino I, Conduto D, Pozzoli L, Dobricic S, Lettieri T. Algal bloom and its economic impact, 2016. JRC Technical Reports, Publications Office of the European Union, EUR 27905 EN; DOI: 10.2788/ 660478. Reproduction is authorized provided the source is acknowledged. On the other hand, the significance of invasive species (also known as alien, exotic, nonindigenous, introduced or nonnative species) in marine ecosystems worldwide has been highlighted and discussed intensively in recent years from ecological, environmental and economic points of view [57–61]. Some of these species can become economically harmful and sometimes even threaten human health [38, 62], like the L. sceleratus puffer fish that has been ranked among the 100 “worst” Invasive Alien Species (IAS) in the Mediterranean Sea with profound social and ecological impacts [63].

8.2.3 Effect to ecosystem from extreme events of climate change Impacts of climate change on the Mediterranean environment particularly relate to water, via a change of its cycle due to a rise in evaporation and a decrease in rainfall. Extreme events, such as heat waves, droughts or floods, are likely to be more frequent and violent. A significant decrease in rainfall, ranging between −4% and −27% for the countries of Southern Europe and the Mediterranean region is

8 Human impact in Mediterranean coastal ecosystems and climate change

263

expected (while the countries of Northern Europe will report a rise between 0 and 16%) [64]. Concomitantly, an increase in drought periods related to high frequency of days during which the temperature would exceed 30 °C is also predicted [65, 66]. Owing to climate change (enhanced evapotranspiration and reduced rainfall) alone, fresh water availability is likely to decrease substantially (by 2–15% for 2 °C of warming), among the largest decreases in the world, with significant increases in the length of meteorological dry spells and droughts. River flow will generally be reduced, particularly in the south and the east where water is in critically short supply. The median reduction in runoff almost doubles from about 9% (likely range: 4.5–15.5%) at 1.5 °C to 17% (8–28%) at 2 °C. Water levels in lakes and reservoirs will also probably decline [67]. This water problem is expected to be of crucial importance with regard to the issue of sustainable development in the region. For instance, floods (resulting from altered rainfall patterns) will affect nutrient loads in the coastal aquaculture areas. High inorganic sediment loads can reduce or arrest the filtration rates of bivalves. Elevated nutrient levels can also stimulate the evolution of HABs. For coastal and offshore aquaculture, more frequent and intense storms result in increased physical damage and stock losses, both of which are costly to operations. Many coastal processes, such as sediment transport, happen mostly during highenergy events (storms). An increase in storm activity may therefore initiate erosion. Any severe flooding event could result in mass mortalities of animals in aquaculture ponds, open-water rafts and lines or cages in coastal and offshore areas. Regarding drought, over much of the Mediterranean basin the general tendency is toward decreasing rainfall. As indicated earlier, the predicted water stress is thought to result in decreasing water availability in the major Mediterranean freshwater systems, areas where there are important aquaculture activities [32, 35]. The effects of climate change on seafood safety are a relatively new topic. Currently, the issue of seafood vulnerability to climate change is scarcely considered both at national and at international levels, despite the fact that climate change is already affecting the biology and ecology of some organisms, as well as several chemical pathways. Seafood security and safety are related issues because unacceptable standards of food safety that render food unfit for human consumption will also impair seafood security, possibly forcing people to consume seafood that are of lower quality or contaminated, or having higher (bio)availability of chemical contaminants. For chemical contaminants, including toxins, a systematic change in marine hydrographical conditions due to a change toward warmer temperatures, reduced salinity and hypoxia may directly affect seafood safety at several levels [68, 69]: (a) increase the input of chemical contaminants to marine systems and consequently the exposure level, particularly due to flood events; (b) change their chemical forms to more toxic ones and thus the exposure level; (c) increase resuspension processes of sedimentbound chemical contaminants; (d) increase their bioavailability, especially with regard to metals, with contaminants being converted to more bioavailable forms (e.g.,

264

Panagiota Katikou

increases in temperature enhance the methylation rate of mercury); (e) diminish the ability of species to deal with toxic substances and the different physiological regulation processes involved in the detoxification of hazardous substances; and (f) modification of contaminant transport pathways to marine systems. The significance of a particular pathway or process depends on the underlying properties (e.g., hydrophobicity, solubility, volatility) and form of the contaminant/toxin (particulate, particle associated, dissolved, etc.) [70, 71]. Conversely, climate impacts on plankton can be direct, such as through the effects of temperature and solar radiation on their physiology and growth substrates, or indirect, through caloric and kinetic energy inputs and freshwater inputs, which determine the availability of essential elements, light and reducing power [72]. In any case, predicting the health of Mediterranean marine ecosystems in response to global warming and other anthropogenic phenomena inevitably leads to understanding and predicting the adaptations of the planktonic community to the changing environment [73].

8.2.4 Ecological response to climate change Susceptibility of organisms to diseases can be affected by three interlinked factors: the organism itself, contaminants and the environment. If any of these three factors are altered, changes in the progression of a disease epidemic can occur. Climate change may impact these factors in various ways, such as by exacerbating the presence of biological contaminants in the marine environment (e.g., toxins produced by HABs) and increasing pathogenic microorganisms’ populations [69]. Climate change on its turn encompasses changes in a number of environmental factors including pH, water level, salinity and temperature [72] and in a number of related changes, for example, oxygen and food availability, which ultimately modify organism performances and adaptation capability [73]. According to climatic models, the Mediterranean basin is bound to be one of the regions most affected by the ongoing warming trend and by an increase in extreme events. There are reasons to believe that the Mediterranean is already one of the most impacted seas in the world, since climate change interacts synergistically with many other influences [74]. Climate change combines with Atlantic influx, Lessepsian migration and the introduction of exotic species by humans to the establishment of tropical marine biota in the Mediterranean Sea [75], which increase the establishment and range extension of tropical exotic species, such as the puffer fish containing the strongest neurotoxin known to date, TTX [76]. Present-day warming ultimately favors the spread of warm water species through direct and indirect effects, and especially by changing water circulation [25]. Furthermore, climate change seems to be responsible for a change in the geographical distribution of certain harmful benthic dinoflagellates. A number of them

8 Human impact in Mediterranean coastal ecosystems and climate change

265

were until recently mainly restricted to circumtropical areas, but have lately spread to temperate regions, including the Mediterranean Sea. Indeed, records of toxic benthic dinoflagellates have dramatically increased along the Mediterranean coast over the past decades and the list of new/emerging harmful species is continuously growing [35, 77–79]. In particular, Ostreopsis ovata, a species considered as tropical, has bloomed in the Mediterranean region in recent years, with increasing frequency, intensity and distribution, in both western and eastern coasts of the Mediterranean, causing benthic organisms’ mortality and human health problems. Health problems also include the formation of toxic aerosols due to wave action, leading to respiratory asthma-like symptoms in humans, such as those reported in the Ostreopsis ovata blooms in the Ligurian coast of Italy in 2005 [80]. In the same context, the tropical genus Gambierdiscus responsible for the production of ciguatera fish poisoning (CFP) ciguatoxins (CTX), has been detected in the Mediterranean Sea, as well as in northeastern Atlantic Ocean, Canary Islands and Madeira [81]. The presence of the rather recently described benthic dinoflagellate species Vulcanodinium spp., responsible for the production of emerging toxins (pinnatoxins, PnTXs), has been repeatedly reported in the Mediterranean region along with the associated toxins in bivalve mollusks [82–85; Greek NRL Marine Biotoxins, unpublished data]. Climate change subjects marine ecosystems to multifactorial stressors such as increased temperature, enhanced surface stratification, alteration of ocean currents, intensification or weakening of nutrient upwelling, stimulation of photosynthesis by elevated CO2, reduced calcification from ocean acidification and changes in land runoff and micronutrient availability. The topic of HABs and climate change is usually addressed by researchers by taking into account only single environmental factors (e.g., CO2, temperature increase, stratification), single biological properties (photosynthesis, calcification, nutrient uptake) or selected “pet” species categories. Complex factor interactions, on the other hand, are rarely covered by simulated ecophysiological experiments [50]. Nevertheless, it seems there are early signs that some parts of the Mediterranean may become less productive in response to climate-driven increased sea surface temperatures and associated reduced nutrient availability [86]. Ecosystems disturbed by pollution or climate change also tend to be more prone to ballast water invasions [87]. Some HAB phenomena, such as toxic dinoflagellates benefiting from land runoff and/or water column stratification as well as benthic dinoflagellates responding to coral reef disturbance may increase, while others may diminish in areas currently impacted [80]. Temperature is probably the most widely recognized component of climate change and also plays a crucial role in determining potential algal growth rates. Consequently, temperature can be of significant influence to the community dynamics of HAB species relatively to their competitors and grazers. In diatoms, for instance, nitrate uptake and reduction rapidly decline at elevated temperatures, thus potentially favoring competing algae. The benthic/epiphytic dinoflagellate genus

266

Panagiota Katikou

Gambierdiscus spp. respond to warming sea surface temperatures and habitat transformation by concurrent spreading of the marine macroalgae with which they are associated [88–90]. Where Ostreopsis spp. is concerned, a number of laboratory experiments examining temperature suggested that Ostreopsis grow more efficiently at high temperatures, but are more toxic at lower temperatures [91–94]. More recently though, a study on the environmental factors and their influence on the toxin production of Ostreopsis cf. ovata during bloom events in the northern Adriatic Sea indicated that the highest cellular toxin content was recorded during the seasonal maximum of water temperature (27 °C) and, although Ostreopsis cells were also detected with water temperatures around 16 °C, no toxins were detected under 20 °C. A significant positive correlation was thus found between cellular toxin content and water temperature values, clearly pointing out the potential impacts of sea temperature warming on toxicity events in the Mediterranean [95]. Climatic changes in conjunction with deteriorated ecosystems near ports and lagoons have also resulted in significant changes of biodiversity due to the introduction and establishment of exotic/invasive species. The majority of exotics are found in the eastern basin (Levantine) of the Mediterranean Sea, obviously due to the opening and periodic widening of the Suez Canal, which caused an increase of their arrival. However, fouling of ships, ballast water exchange, aquaculture and the aquarium trade are also considered responsible for the introduction of invasive species into this region. [96, and references thereof]. The introduction of exotic species (more than 600 records in 2004) is a dynamic nonstop process with approximately 15 new species reported each year. It is noteworthy that in the twenty-first century, 64 new species have been reported in the Mediterranean, with 23 of them recorded in 2004 [53]. In 2008, it was estimated that the number of recorded alien species in the Mediterranean Sea was continuing to increase at a rate of one new record every nine days. Latest reports (2010–2016), considering only multicellular alien species, indicate the rate of introductions is 11 species per year. The Mediterranean Sea can thus be considered an early-warning system for other European marine environments. In fact, it has been estimated that 76% of the first marine introductions of invasive species across Europe were reported first from the Mediterranean Sea, with 54% of them first reported in the eastern Aegean-Levantine Sea [96, 97, and references thereof].

8.3 Emerging toxins in the Mediterranean Sea Marine biotoxins detected worldwide, but particularly in European waters, were originally classified based on their acute symptomatic effect in humans following intoxification. The three main groups monitored in the EU, regulated since 1991

8 Human impact in Mediterranean coastal ecosystems and climate change

267

by the Directive 91/492/EEC [98] were: (a) Paralytic Shellfish Poisoning (PSP) toxins; (b) DSP toxins; and (c) ASP. However, due to the progress of alternative detection methods, classification of DSP toxins has changed to focus more on the chemical structures and properties of the toxins. DSP toxins have therefore become known as lipophilic toxins incorporating okadaic acid, dinophysistoxins, azaspiracids, pectenotoxins and yessotoxins, with the last two not proved to cause diarrheic symptoms following intoxication. For each of these three main toxin groups and subgroups, the occurrence of the toxins, their chemical characteristics, toxicokinetic evaluations, human-exposure assessments and detailed review of potential methods of analysis have in recent years been published by the European Food Safety Authority (EFSA) as scientific opinions [99–104]. The diversity of the numerous analogues or natural enzymatic metabolites of marine biotoxins has been described [105]. However, this series of reviews also included prospective emerging toxins to European waters such as cyclic imines, palytoxin (PLTX), TTX, maitotoxin, CTXs and the neurotoxin-poisoning brevetoxins (BTXs), as their occurrence could have severe implications with regard to seafood safety [106]. Although EFSA had by 2010 already published the relevant scientific opinions for the emerging toxin groups of PLTXs, CTXs, cyclic imines and BTXs [107–110], and recently for the group of TTXs [111], still none of these toxins are regulated by the current EU legislation [112, 113].

8.3.1 Identified emerging toxins and climate change effects 8.3.1.1 Tetrodotoxin TTX is one of the most potent low-molecular-weight marine neurotoxins (319 Da). Its chemical structure was described as a cage-like polar molecule with a cyclic guanidinium moiety fused to a dioxy-adamantane skeleton embellished by six hydroxyl groups. To date, at least 30 different analogues of TTX have been reported. Their degree of toxicity varies among analogues, although there is still significant uncertainty regarding their relative toxicity [114–116]. An important recognized feature is that the deoxy analogues of TTX are less toxic than TTX, while the hydroxyl analogues are more toxic than TTX. For instance, 11-oxoTTX is four to five times more toxic than the parent TTX, while 5-deoxyTTX, trideoxyTTX, 4-CysTTX and anhydroTTX have negligible toxicity [115]. TTX is considered to be produced by certain endosymbiotic bacteria, such as Vibrio sp., Pseudomonas sp. and Alteromonas [117], although recently a contribution of the microalgae Prorocentrum minimum in the occurrence of TTX in shellfish has been suggested [118,119]. Similarly to saxitoxin (STX), TTX consumption induces severe symptomatology in humans starting from numbness and mild gastrointestinal effects to respiratory paralysis, and even death of human consumers. Consequently,

268

Panagiota Katikou

bioaccumulation of TTX in seafood and subsequent entrance in the human food chain poses a real and very significant risk to human safety [111]. TTX intoxication is most commonly associated with the consumption of puffer fish and sometimes by the ingestion of gastropods or crabs. The close relationship between TTX and the tropical environment, where for a long time most of the species bearing this toxin inhabited, explains the fact that the large majority of the reported poisoning cases from this toxin were until recently confined to the southwest Asian area, especially in Japan where the regular consumption of Fugu-related cuisine results in most of the registered events. Despite this initial geographical limitation, a visible increase of TTX intoxication cases in MWs, where such cases should be unlikely, is occurring [120]. Since 2007, TTX has been gradually detected in various marine organisms collected from Mediterranean countries and specifically in three different puffer fish species (L. sceleratus, L. suezensis, Torquigener flavimaculosus) [44, 121–126], in the marine gastropod Charonia lampas lampas [122, 127, 128], and quite recently in bivalve mollusks, such as cultured Mytilus galloprovincialis mussels and infaunal Venus verrucosa clams in Greece [118] and M. galloprovincialis mussels in the north and south of Italy [129–131]. Significant TTX levels have also been found in bivalves from other European, non-Mediterranean, countries located in more northern latitudes, such as England and the Netherlands [132–134], while low TTX levels have also been detected in Spanish and Portuguese shellfish harvested in the Atlantic coasts [135–138], indicating a quite widespread presence of this toxin at a European level and a new potential source of risk for public health [139]. This new phenomenon of displaced TTX detection cases and the global increase of water temperatures can be linked to both the increase of the TTX-vectors presence in MWs, as well as to the increase of TTX contents in the actual vectors. A number of researchers attribute the new occurrence of TTX in European regions to the so-called Lessepsian migration (Figure 8.2). In 1869, the opening of Suez Canal caused the migration of many Red Sea species through the new waterway, which have settled and have been well established in the Eastern Mediterranean. These migrations are a growing phenomenon accompanying closely the increase in global temperature [140]. Among the most significant migrations in the Mediterranean Sea is that of puffer fish species of the Tetraodontidae family, known vectors for TTX, including L. sceleratus [43, 44, 123, 141, 142], L. suezensis [125, 143], L. lagocephalus [144] and Torquigener flavimaculosus [124]. Concerning the case of L. sceleratus, one of the most poisonous Lessepsian species discovered, it has already been related to cases of human intoxication [142, 145] along the Mediterranean shore. Its increased occurrence is currently considered to cause ecological and economic damages and is viewed as a pest by fishermen as it is capable of reducing the local stocks of important commercial cephalopod species, damaging fishing gears, deterring customers from buying fish and introducing additional effort to discard the fish [146, and references thereof]. These impacts have led scientists to classify L. sceleratus invasion among the 100 worst marine invasions in the Mediterranean basin [63].

8 Human impact in Mediterranean coastal ecosystems and climate change

269

Published studies are rather inconsistent when relating a water temperature increase to the increase of TTX content in its vectors, although certain studies supporting this argument exist. Matsumoto et al. [147] clearly relate an increased intake rate of TTX into the liver tissue of Takifugu rubripes with the increase in environmental temperature; however, this relation between temperature variation and TTX toxicity potential remains obscure as is the case with the mechanisms involving excretion and accumulation of the toxin on puffer fish [120]. The seasonal rise in environmental temperatures is considered to be a determining factor in the toxicity variation of TTX vectors. Several such studies have attempted to establish a valid correlation between the two. For instance, tests performed in specimens of Taiwanese gobies were performed and showed that a small percentage of the animals had a measurable TTX content between August 1996 to July 1998 being observable at both regional and seasonal variation of the toxin, with higher values from March to November [148]. These test results differ with those from another study performed during an annual period (August 2000 to August 2001) in Indonesia on the puffer fish Lagocephalus lunaris, which showed that the animals remained toxic through nine months (March to November), with the toxicity peak being observed in August with 100% of the test animals possessing the toxin [120, 149]. Still there is an obvious lack of these systematic studies and more research is required to establish any kind of definitive hypothesis on the seasonal variance of TTX vector toxicity, taking also into account that higher TTX contents are also associated with the spawning period of puffer fish, a period which coincides with the summer months in the Mediterranean Sea [150, 151]. Despite the fact that the epidemiologic perspective of TTX is well studied, its molecular and cellular mechanisms are still not clear enough. As indicated earlier, certain bacteria of the microflora are thought to be responsible for TTX production but no studies on their kinetics are available to understand their possible interactions with the current changing climate patterns. One of the scarce studies to relate changes in water temperature with changes in the bacterial content of TTX vectors (in this case the puffer fish Fugu niphobles) demonstrated that the bacterial content of the skin, gills and intestines of the fish was actually affected by temperature variations. Specifically, the identification of bacteria of the genus Vibrio, known TTXproducers, was positive in temperatures of 20 and 29 ºC, but negative at 10 ºC. These results were further confirmed in laboratory conditions when the same bacteria were cultivated, when again all strains were able to grow at 20 and 29 ºC, but very few were able to do so at 10 ºC suggesting their preference for higher water temperatures [120, 152]. Ballast waters on the other hand can also be responsible for the transfer of TTX-containing organisms from Asian waters to European waters. Over the last 20 years, spreading of marine mucilage in the Mediterranean Sea has also been observed due to sea surface warming, which helps the survival of migrated species in this marine environment [115, 153]. Increased TTX incidence has been associated with higher water temperatures also in the case of bivalve mollusks and shellfish in general. For instance, in northern

270

Panagiota Katikou

Italy the highest TTX concentrations (max. 541 μg kg–1) were found in mussels harvested in shallow estuarine waters (1–2 m depth) with water temperatures ranging between 18 and 20 °C [129, 130]; similarly the TTX-contaminated mussels in Sicily, southern Italy, were harvested from the harbor area, where they were found attached at ca. 1 m depth [131]. Higher temperatures are thought to be related to the quite consistent seasonal occurrence of the TTX toxic episodes reported from the Netherlands. TTX-positive samples were generally present in the summer months (June–August) with an increase in TTX concentrations being observed during late June (max. 253 μg kg–1), followed by a rapid decline in July [134]. This was also the case with English shellfish, where strong indications of increased toxicity during the summer months were noted in all six areas sampled within the study. Water temperature recordings obtained during sampling were compared to the data on TTX concentrations, resulting in a general indication that TTXs were typically present in shellfish collected from areas with water temperatures of ≥15 °C, with few exceptions. Despite the facts that no statistical correlations were established between water temperature and total TTX contents and also that no presence of TTXs was detected in a large number of shellfish samples harvested in water above 15 °C, this temperature seemed to be the threshold above which TTXs were more likely to occur in shellfish tissue. Moreover, water depth seemed to also play an important role, taking into account that 51 out of the 55 TTX-positive samples (93%) originated from intertidal or shallow water environments (0–5 m depth), while the four remaining ones were all harvested from the same medium-depth location (5–20 m depth) [133]. The Japanese government has long ago established a regulatory limit of 2 mg kg–1 of TTX equivalents in food, due to “fugu” consumption. In contrast, no regulatory limits for TTX were set in the EU as this type of poisoning had not ever before been considered as a potential problem. However, back in October 2007, the first toxic European episode was reported in Malaga (Spain), caused by the ingestion of a trumpet shell of the species Charonia lampas lampas. The product was purchased in a Malaga market, but was originally caught in the south coast of Portugal [128]. Again, in 2005 and 2008, puffer fish L. sceleratus present in the Mediterranean Sea due to the Lessepsian migration phenomenon have been incriminated for several severe poisonings and deaths (respiratory distress few minutes after the meal by sudden paralysis due to the brutal decrease of the neuromuscular transmission) following consumption of this highly toxic invader in Israel and Lebanon [142, 143, 154]. This fish is also largely present in Turkey, Greece and Cyprus [44, 144, 155, 156], whereas despite its rather recent introduction, L. sceleratus has already largely expanded in the whole Mediterranean Sea, with proven presence in Tunisia [157], Libya [158] Italy [159, 160], Malta [161], Croatia [162], Algeria [163] and Spain [126, 164]. An equally toxic indigenous puffer fish species L. lagocephalus is also sporadically observed along the North African Mediterranean coast from Morocco to Libya, but this pelagic species from the Atlantic is rarely caught [165].

8 Human impact in Mediterranean coastal ecosystems and climate change

271

Publication of the first reports on TTX presence in bivalve mollusks in different European countries, starting from 2015 [118, 132, 139], together with the reported intoxication cases and the known very toxic nature of the compound, raised concerns about TTX as a food safety hazard in Europe, due to the lack of a maximum permitted level (MPL) for TTXs content in seafood within the EU and the fact that this toxin group is not monitored on a regular basis [166, 167]. Indeed, the only relevant requirement foreseen in the current EU legislative framework is that fishery products derived from poisonous fish of the family Tetraodontidae must not be placed on the market [168–170]. As a result, national health measures were introduced in the Netherlands [171] and subsequently the European Commission requested the “EFSA Panel on Contaminants in the Food Chain” to deliver a scientific opinion on the risks related to the presence of TTXs in marine bivalves and gastropods, whose opinion was adopted on March 2017 and published later on April [111]. The risk assessment conducted by the CONTAM Panel within the framework of the EFSA scientific opinion resulted in the introduction of a proposed safe concentration of lower than 44 μg TTX eq. kg–1 of shellfish meat, which was considered as not expected to result in adverse effects in humans (NOAEL), with calculations based on a large portion size (400 g), an adult body weight of 70 kg and a group acute reference dose (ARfD) of 0.25 μg kg–1 b.w. The opinion, however, identified a series of uncertainty factors and several limitations due to inadequacy of the data available at the time, which resulted in a number of recommendations for necessary information, required in order to provide a more refined exposure assessment [111]. Despite the fact that a significant amount of research has been undertaken during the time period following the adoption of the relevant EFSA scientific opinion, there are many aspects of the EFSA recommendations that have not been yet addressed, before the possible introduction of a legislative maximum acceptable level in shellfish, or at least the adoption of a uniform regulatory management approach for this toxin group at EU level [139]. The fact that the TTX levels detected in shellfish in the EU are often higher than the NOAEL of 44 μg TTX eq. kg–1 of shellfish meat, with the maximum value so far reported (541 μg kg–1 in Italy) being at least tenfold, indicates that Mediterranean seafood is endangered of being contaminated with this hazardous toxin and that measures are possibly required to protect human health.

8.3.1.2 Palytoxin PLTXs have been originally detected in marine zoanthids (soft corals) of the genus Palythoa, but thereafter the production of several analogues was also confirmed in benthic dinoflagellates of the genus Ostreopsis (e.g., O. siamensis, O. mascarenensis, O. ovata). PLTXs were first reported in Hawaii and Japan, in warm waters where the soft corals naturally occur, but are currently known to be distributed worldwide

272

Panagiota Katikou

[172,173], especially after the uncovering of the broad distribution of Ostreopsis spp. Actually during the past two decades, blooms of Ostreopsis spp. are increasingly being reported in Mediterranean countries, such as France, Greece, Italy, Spain, Cyprus, Croatia, Tunisia, Lebanon and Egypt [174–177]. On the other hand, a number of different PLTX-like compounds have been identified in the Mediterranean strains of O. cf. ovata and in the newly described species O. fattorussoi [178], such as putative PLTX [175] later renamed as isobaric PLTX [179], ovatoxins-a [175], -b, -c, -d, -e [180] -f [181], -g [179], -h [182], -I, -j1, -j2, -k [176] and –l [183]. A bloom of Ostreopsis spp. on the coast of Algarve (south of Portugal) and the presence of both O. cf. ovata and O. cf. siamensis in Lagos Bay (south coast of Portugal), the latter also found in Lisbon Bay (west coast of Portugal), indicate that species capable of producing PLTX analogues such as ovatoxin may be spreading from the Mediterranean to the north Atlantic [184–186]. Temperature is a crucial factor determining both growth potential and toxin production of the genus Ostreopsis. Clonal laboratory cultures of O. lenticularis exposed to elevated temperatures (30–31 °C) for 33 and 54 days showed significant increase in the quantity of extractable toxin they produced as compared to their toxicities versus cells grown at temperatures of 25–26 °C. Furthermore, O. lenticularis samples collected directly from the field following exposure to elevated temperatures for comparable periods of time also showed significant increases in extractable toxin [187]. The increased toxicity of both field-sampled and laboratory-grown O. lenticularis exposed to elevated temperatures is considered to result from the effects of elevated temperatures on their metabolism and/or the symbiotic bacterial found associated with these microalgae. The number of bacteria associated with cultured O. lenticularis exposed to elevated temperatures was significantly reduced. Increased toxin recovery from O. lenticularis exposed to elevated temperatures, on the other hand, may have resulted from the direct effect of temperature on toxin production and/or the reduction of Ostreopsis-associated bacterial flora that consume toxin in the process of their growth. This reduction in the quantity of associated bacterial flora in temperaturetreated cultures, in its turn, may result in increased toxin recovery from O. lenticularis due to a reduction in the consumption of toxin by these symbiotic bacteria [187]. Alternatively, the optimum temperatures for growth and toxicity of O. ovata were found to be inversely related. High water temperatures (26–30 °C) stimulated O. ovata cells growth rate and biomass accumulation and low toxicities while lower temperatures (20–22 °C) induced higher toxicity per cell and lower cell numbers. Based on these results it was suggested that increased sea surface temperature, which can result from global warming, may play a crucial role inducing the geographical expansion and biomass increase, blooms of O. ovata, in future [93]. In another study, it was also concluded that environmental conditions seem to play a key role in influencing the abundance of Ostreopsis spp. High cell densities of an Adriatic O. cf. ovata isolate were generally recorded in concomitance with relatively high temperature and salinity and low hydrodynamic conditions. The highest growth rates of that

8 Human impact in Mediterranean coastal ecosystems and climate change

273

Adriatic strain were recorded for cultures grown at 20 °C and at salinity values of 36 and 40 psu, in accordance with natural bloom surveys. Toxicity was also affected by growth conditions, with the highest toxin content on a per cell basis being measured at 25 °C and salinity of 32 psu. However, the highest total toxin content on a per liter basis was recorded at 20 °C and salinity of 36 psu, since under such conditions the growth yield was the highest [188]. Similarly, O. cf. ovata from Japanese waters was found to grow faster than other benthic toxic-dinoflagellates [Coolia monotis, Gambierdiscus toxicus and Prorocentrum lima (Dinophyceae)] in waters of high temperature and salinity. This physiological feature was considered to confer an ecological advantage on O. cf. ovata in the bloom development during warmer seasons and could be responsible for outbreaks of PTX-like poisoning, especially during the warmer seasons [189]. The relationship of toxin concentrations with environmental parameters throughout an O. cf. ovata bloom was recently studied in the northern Adriatic Sea. It was found that high temperature and balanced nutrient conditions were the optimal environmental conditions to start and sustain blooms as well as to maximize toxin production. On the other hand, Ostreopsis showed a gradual decrease of toxin content throughout the bloom ascribed to the occurring of the same nonoptimal conditions that led to the bloom decline. It was additionally indicated that the toxin fraction released during a natural bloom could be higher than that released in batch culture and that the first bloom phase is potentially the most dangerous to human health [95]. The PLTX group constitutes one of the most poisonous nonprotein marine toxins known to date. They present high acute toxicity in animals by the intravenous or intraperitoneal route (e.g., an i.v. lethal dose (LD50) ranging from 0.15 to 0.73 μg kg–1 in mice), but the oral route has been reported as the least sensitive [190]. Despite this fact, a number of acute toxicity cases and deaths have been reported from human outbreaks but still reliable quantitative data on acute toxicity in humans are unavailable. In view of the acute toxicity reports and the lack of chronic toxicity data for the PLTX-group toxins, the EFSA Panel on Contaminants was able to derive an oral ARfD of only 0.2 μg kg–1 b.w. for the sum of PLTX and its analog ostreocin-D. In order for a 60 kg adult to avoid exceeding the ARfD, a 400 g portion of shellfish meat should not contain more than 12 μg of the sum of PLTX and ostreocin-D, corresponding to 30 μg kg–1 shellfish meat [107]. To date, there are increasing records of PLTX presence largely above these levels in many edible marine organisms from the Mediterranean Sea. For instance, in the Mediterranean coasts of France, maximum levels of PLTXs contamination in tissues of edible organisms tested by LC-MS/MS were higher than the EFSA proposed level of 30 μg kg–1 in the sea urchin Paracentrotus lividus, the redmouthed rock shell Stramonita haemastoma, the warty crab Eriphia verrucosa and the flathead mullet Mugil cephalus, with the last one presenting the highest toxin level of the study being 392.2 μg kg–1 for the sum of OVTX-a and PLTX per kg of digestive tube [191]. In subsequent studies in the same area, PLTXs levels in contaminated sea urchins ranged from 103 to 423 μg kg–1 of whole flesh, with an average of

274

Panagiota Katikou

223 μg kg–1 revealing a high variability of contamination levels between individual organisms, while in the case of contaminated sea breams, PLTX levels ranged between 33 and 152 μg kg–1 in the whole flesh [192], while PLTX levels above 30 μg kg–1 were also found in the spinous spider crab Maja squinado (51.3 μg PLTX eq. kg–1) and the banded dye-murex Hexaplex trunculus (40.4 μg PLTX eq. kg–1) [193]. Despite these data though, the EU still has not adopted a maximum permissible limit to confront the risk of PLTX poisoning of European consumers [10, 194, 195].

8.3.1.3 Cyclic imines (Gymnodimine, Spirolides, Pinnatoxins) The emerging toxin group of cyclic imines (CIs) consists of spirolides (SPXs), gymnodimines (GYMs), pinnatoxins (PnTXs) and pteriatoxins (PtTXs) and is a family of marine biotoxins largely present in shellfish and other marine organisms. They are macrocyclic compounds with imine (carbon–nitrogen double bond) and spirolinked ether moieties. They have been grouped together because of their common imine group as a part of a cyclic ring, which confers the pharmacological and toxicological activity, and due to their similar acute “fast acting toxicity” in mice [109, 196]. In addition to SPXs, GYMs, PnTXs and PtTXs, the CI group comprises prorocentrolides and spiroprorocentrimines, which to date have not been reported in European shellfish. Although SPXs, GYMs, PnTXs and PtTXs are now known to occur in microalgae and/or shellfish in several parts of the world (Canada, Denmark, New Zealand, Norway, Scotland, Tunisia, USA and Japan), no information has been so far reported linking these toxin groups to poisoning events in humans [197–201]. This fact explains the current absence of regulatory limits and official analysis methods for CIs in shellfish at a global level. Furthermore, the relevant EFSA opinion on CIs concluded that at the time it was issued (2010) the estimated exposure, at least to SPXs, did not raise concern for the health of the consumer, although it was stressed out that this conclusion for SPXs was based on very limited toxicity data, whereas available data on the other CI toxins were at that time inadequate to provide a risk assessment [109]. However, the working group of the EU-RLMB back in 2005 had proposed a guidance level for the sum of SPXs of 400 μg kg–1 in shellfish meat [202, 203]. Very recently, the French Agency for Food, Environmental and Occupational Health & Safety (Agence nationale de sécurité sanitaire de l’alimentation, de l’environnement et du travail – ANSES) has issued a scientific opinion with regard to assessment of the health risks associated with PnTXs in shellfish. In this opinion, the ANSES working group has concluded that there may be a health concern related to the consumption of shellfish contaminated with PnTXs and has proposed a provisional acute benchmark value of 0.13 μg PnTX-G kg–1 bw for humans and a provisional maximum tolerable concentration of 23 μg PnTX-G kg–1 of total shellfish meat [204].

8 Human impact in Mediterranean coastal ecosystems and climate change

275

Cyclic imines are produced by different dinoflagellates to a different extent: SPXs are mainly produced by Alexandrium ostenfeldii also known as A. peruvianum [205, 206], GYMs are produced by Karenia selliformis, also known as Gymnodinium selliforme [207], but also by certain strains of Alexandrium ostenfeldii [208–209]. The PnTXs-producing organism has been rather recently been identified as Vulcanodinium rugosum [82], prorocentrolides have been isolated from Prorocentrum lima [210], spiroprorocentrimines are suggested to be produced by Prorocentrum species [211] and PtTXs are suggested to be biotransformation products of PnTXs in shellfish, so no specific producing organism has been identified [109]. Spirolides have been identified in a number of European countries bordering the Mediterranean Sea, Atlantic coast and the North Sea. SPXs have been found in their producer dinoflagellate A. ostenfeldii/peruvianum in Scotland [212], Italy [213], Denmark [214] and Ireland [206]. They have also been found in shellfish from Norway [215], Spain [216], Italy [203], Croatia [217–219], Portugal and Slovenia [85]. SPXs are also commonly detected in Greek shellfish since 2008 [220], at concentrations ranging from trace levels up to the highest value of 118 µg kg–1 found in hairy mussels (Modiolus barbatus), with increased levels producing positive MBA assays (Greek NRL Marine Biotoxins, unpublished data). Gymnodimines have been first found in shellfish from New Zealand, specifically in greenshell mussel, blue mussel, scallop, cockle, surf clam, oyster and abalone [221, 222]. To date, there are limited records for the presence of GYMs in shellfish produced in European countries, such as in Croatia, where GYM-A has been detected in M. galloprovincialis mussels (range 5–15 μg kg–1), in Acanthocardia tuberculata cockles (range 2.65–15.77 μg kg–1) and in Callista chione smooth clams (range 1. 17–6.14 μg kg–1) [217–219]. In addition, we have repeatedly determined GYMs in Greek shellfish (mussels, venus clams and hard clams) at concentrations ranging from trace levels up to 66 µg kg–1 within the framework of the Greek HAB monitoring program, with the highest concentrations coinciding also with positive MBA tests (Greek NRL Marine Biotoxins, unpublished data). With regard to other Mediterranean countries, GYM has also been identified in Ruditapes decussatus clams from the Gulf of Gabès in Tunisia [223–225]. The low acute toxicity of gymnodimine when ingested with food (>7,500 µg kg–1 in mice) suggests that this compound is of low risk to humans, a conclusion consonant with anecdotal evidence for the absence of harmful effects in individuals who had consumed shellfish contaminated with gymnodimine [226]. Despite this fact, the chronic toxicity of GYM remains unclear as its role in the development of neurodegenerative illnesses like Alzheimer or Parkinson’s diseases has been debated [227, 228]. PnTXs were identified for the first time in shellfish in Europe during the last decade. They have so far been found in Norwegian mussels [229], French mussels and clams [83], Spanish mussels and oysters [84, 201], as well as in Italian and Slovenian mussels [85]. The concentrations found in Spanish, Italian and Slovenian shellfish were generally low, usually not exceeding 5 μg PnTX-G kg–1 shellfish [84, 85, 201].

276

Panagiota Katikou

However, the levels detected in French mussels derived from the Ingril lagoon (Mediterranean Sea) were extremely high, with concentrations reaching a maximum of 1,244 μg PnTX-G kg–1 shellfish tissue in 2010 and 1,143 μg PnTX-G kg–1 shellfish tissue in 2015, largely above the ANSES proposed provisional maximum tolerable concentration of 23 μg PnTX-G kg–1 of total shellfish meat [83, 204]. PnTX-G is also commonly detected in Greek shellfish, with increased concentrations occurring in summer and autumn months and a highest value of 116 μg kg–1 found in V. verrucosa clams. PnTXE and PnTX-F have also been detected in low levels, whereas there is also evidence for the presence of PnTX fatty ester metabolites (Greek NRL Marine Biotoxins, unpublished data). The recent findings suggest that PtTXs are transformed from PnTXs in shellfish, while their presence in Europe has only been reported so far at nonquantifiable traces in French shellfish [83]. Alexandrium ostenfeldii has been widely observed in temperate waters of Europe [230], North America [205], the Russian Arctic [231] and Eastern Siberian Seas [232]. There are also records of the occurrence of A. ostenfeldii from the coast of Spain [233], the Mediterranean [234], New Zealand [235], Peru [236] and Japan [237]. However, for a long time, A. ostenfeldii has been considered mainly as a background species, occurring at low cell concentrations mixed with other bloom forming dinoflagellates [212, 230, 238]. Only in the past decade it has gained increasing attention when dense blooms of this species (or its synonym A. peruvianum) were reported, for example, from South America [236], the Northern Baltic Sea [239], along the Adriatic coast of Italy [213], the estuaries of the US East coast [240] and the Netherlands [241], where annually recurrent dense blooms up to 4.5 × 106 cells L–1 have been recorded since 2012 [242]. It is not clear whether the recent increase in bloom events is due to anthropogenic spreading or changing environmental conditions favoring bloom formation. Most of the recent blooms occurred during summer in coastal areas and were associated with warm water periods (e.g., [243]). Experimental studies indicate that increased water temperature has a favorable effect on A. ostenfeldii bloom populations and it has been suggested that changing climate conditions promote bloom formation [242, 244]. The species produces PSP toxins [245], spirolides [205] and gymnodimines [246], and all compounds may even occur together in one strain [240]. Thus, an increase of A. ostenfeldii bloom events with several potent toxins involved may represent a new risk to the environment that is associated with climate change [247]. The situation is further perplexed by the fact that toxin composition of A. ostenfeldii was consistently altered by elevated temperature and increased CO2 supply in the strains tested by Kremp et al. [244], resulting in an overall promotion of saxitoxin production, with potentially severe consequences for the coastal ecosystem. In a study regarding growth and toxicity of a cultured strain of Karenia selliformis, where different temperature and salinity combinations were tested, it was concluded that growth rates were similar for the different salinities tested, but showed an increasing trend with water temperature increasing from 15–17 °C to 20–21 °C. Toxicity

8 Human impact in Mediterranean coastal ecosystems and climate change

277

of the culture, on the other hand, seemed to be dependent on culture age and growth phase, with the declining phase coinciding with higher GYM production [248]. Vulcanodinium rugosum [82] was first described from water samples of Mediterranean lagoons, with a water temperature of 23.3 °C and a salinity of 36.5 psu. In another study, V. rugosum cells were isolated and cultured from the tropical Mexican Pacific, where the local conditions of surface temperature and salinity, in which the cells were found, were 28°C and 35 psu, respectively, and the culture conditions were similar. It was concluded that the species seemed to prefer warm waters and fully marine environments [249]. Vulcanodinium spp. is now also quite common in Greek waters [250], especially during the warm periods of the year. In a recent study of a V. rugosum strain isolated from the French Mediterranean lagoon Ingril in 2010, where the effect of temperature and salinity on growth, PnTX-G production and chlorophyll a levels of this dinoflagellate was tested, it was concluded that the optimal combination for growth was at a temperature of 25 °C and a salinity of 40 psu. These results indicate that V. rugosum is euryhaline and thermophile, which could explain why this dinoflagellate develops in situ only from June to September. Additionally, V. rugosum growth rate and PnTX-G production were highest at temperatures ranging between 25 and 30 °C, suggesting that in the future this dinoflagellate could give rise to extensive blooms caused by the climate change-related increases in temperature expected in the Mediterranean coasts [251]. Additionally, the fact that, to date, this genus has been detected in Australia, New Zealand, Japan [252], Mexico [253] and coastal Mediterranean areas (France, Greece) is also indicative of the species preference to warmer temperatures, thus the temperature increase resulting from climate change could indeed create risks by providing more optimal conditions for their growth. Currently, the situation remains that there are no regulations on CIs in shellfish in Europe or in other regions of the world, as no acute human intoxication by consumption of shellfish contaminated by PnTXs has been reported [201]. Given the novelty of the research area of CIs, it is crucial to better describe their mechanisms of action, as well as to widen the toxicological and pharmacological data in order to determine if they pose a public health risk. This is particularly important considering the widespread distribution of CIs in seafood and their potent binding to nicotinic acetylcholine receptors in the central and peripheral nervous systems and this might have long-term effects on human health. It would also be important to develop suitable methodologies for the detection of all CIs and implement preventive measures in monitoring programs [253].

8.3.1.4 Ciguatoxins CTX-group toxins are marine biotoxins that occur in fish as a result of biotransformation of the CTX intermediates (52-epi-54-deoxy-CTX-1B and 54-deoxy-CTX-1B)

278

Panagiota Katikou

and gambiertoxins produced by several toxic benthic dinoflagellates belonging to the genera Gambierdiscus and Fukuyoa [254–258]. These are mainly found in Pacific, Caribbean and Indian Ocean regions and they are classified as Pacific (P), Caribbean (C) and Indian Ocean (I) CTX-group toxins. Recently, however, Gambierdiscus sp. has been also detected in the Mediterranean Sea, whereas CTXgroup toxins were identified for the first time in fish caught in European territories. The first report in Europe on the occurrence of Gambierdiscus sp., potential producer of CTX, was in Crete Island, Greece, in 2003 [259]. Later, the intoxication of fishermen that have eaten some fish caught off the Madeira island archipelago revealed the presence of CTXs [260]. So far, no regulatory limits exist for CTXs in fish in Europe, but the legislation requires that no fish products containing CTXs are placed on the market [108]. A Guidance Level of 40 ng CTX kg–1 fish had been recommended by the working group of the EU-RLMB back in 2005 [202], whereas the relevant EFSA opinion on CTXs, based on case reports on human intoxications, concluded that a concentration of 0.01 μg P-CTX-1 eq. kg–1 fish is expected not to exert effects in sensitive individuals when consuming a single fish meal [108]. Similarly, the United States Food and Drug Administration (US FDA) has proposed guidance levels of ≤0.1 μg kg–1 of CCTX-1 equivalents and ≤0.01 μg kg–1 of P-CTX-1 (initially named CTX-1B, both names currently in use) equivalents [261]. According to an epidemiological risk assessment based on certain CFP events in Guadeloupe (French West Indies) a lowest observable adverse effects level (LOAEL) was established at 4.2 ng P-CTX-1 eq. per individual corresponding to 48.4 pg P-CTX-1 eq. kg–1 bw. Although based on limited data, this calculated LOAEL was consistent with both the EFSA proposed level and the US FDA guidance levels, as the lowest concentration of CTXs in fish responsible for intoxication in that study was found at 0.0220 μg P-CTX-1 eq. kg–1 of fish [262]. Moreover, since the unique currently existing certified standard is for P-CTX-1, toxicity equivalence factors (TEFs) for CTX congeners have been established by acute toxicity in mice (LD50) as follows: P-CTX-1 = 1, P-CTX-2 = 0.3, P-CTX-3 = 0.3, P-CTX-3C = 0.2, 2,3dihydroxy P-CTX-3C = 0.1, 51-hydroxy P-CTX-3C = 1, P-CTX-4A = 0.1, P-CTX-4B = 0.05, C-CTX-1 = 0.1 and C-CTX-2 = 0.3. These TEFs should be applied to express individual analogues identified with quantitative detection methods as P-CTX-1 equivalents [108, 196]. CTXs can enter into the food webs through the consumption of the toxic dinoflagellates by herbivorous fish and their subsequent consumption by carnivorous fish [263]. Humans would be intoxicated by consumption of herbivorous and carnivorous fish containing CTXs. CTX precursors accumulate in fish tissue (mainly in viscera, but also in the muscle or other parts [264]) and may later be metabolized into different CTX forms, which are responsible for human intoxication. A total of at least 40 analogues from the Pacific Ocean (P-CTXs) [254–256, 265–269], from the

8 Human impact in Mediterranean coastal ecosystems and climate change

279

Caribbean Sea (C-CTXs) [269, 270] and from the Indian Ocean (I-CTXs) [269, 271] have been identified. CFP is the foodborne illness and is responsible for the highest reported incidence of human poisoning from seafood consumption worldwide [81, 256]. At the time the EFSA opinion was issued (2010), an estimation indicated that between 10,000 and 50,000 people were suffering around the world from this disease annually [108]. As a rule, CFP is endemically found in Indo-Pacific and Caribbean areas. However, in recent years CTXs are appearing in countries not expected for their latitude, such as waters close to European and African continents, for example, in the Canary Islands (Spain) [272–274] and Madeira (Portugal) [260, 273–276]. The presence of CTX-like substances in edible fish of the eastern Mediterranean of Israel was also suspected in the case of an assumed ciguatera poisoning incident in this atypical site after consumption of the fish Siganus rivulatus [277]. This northernmore expansion has been attributed, for instance, to changes on the distribution of toxinproducing microalgae to northern latitudes [278] or to migration of fish containing the toxins [279]. In support of the former, the presence of Gambierdiscus sp. has been recorded in the eastern Mediterranean Sea, from 2003 in Crete Island [259]. Prior to this recording, the northernmost area where it had been found was the Canary Islands – approximately 28°N [259, 280, 281]. The aforementioned finding in Crete and later the detection of Gambierdiscus in Cyprus increased the latitude to which this genus is distributed from 28° N to just above 35°N. However, the northern geographical boundaries of Gambierdiscus distribution were further expanded by the detection of Gambierdiscus on February 2009 in Saronicos Gulf, Salamina Island, at a latitude of about 38°N [78]. A few years later, Gambierdiscus sp. was detected in the Western Mediterranean Sea (Balearic Islands), with the species G. australes considered as well established at different locations around the coasts of both Majorca and Minorca, at a latitude of about 39°N [282]. This latter record currently constitutes the northernmost point of Gambierdiscus distribution worldwide. One species of Fukuyoa (F. paulensis), a genus that includes species previously included in the genus Gambierdiscus, was reported in the Balearic Islands in 2015 [283]. Additionally, strains of the genus Fukuyoa have also been detected in Cyprus expanding the spatial area of the genus as it was previously restricted to the island of Formentera (Western Mediterranean), while the number of the detected Gambierdiscus species in Crete has been now increased to four (G. cf. toxicus, G. silvae, G. cf. belizeanus, G. carolinianus) [284]. Recent toxicity studies by the N2a cell assay indicate the existence of CTX-like toxicity in Gambierdiscus strains derived from the Balearic Islands, Crete and Cyprus. In the case of fish sampled in the Mediterranean Sea the N2a assay no CTX positives have been identified, but there were five dubious cases concerning fish samples originating from Cyprus [274]. All these findings, thus, suggest a possible future concern about CTX-group toxins in fish and seafood originating from Europe.

280

Panagiota Katikou

The occurrence of representatives of what was once considered a tropical or subtropical genus, like Gambierdiscus, is in accordance with the suggestions of several researchers regarding climate change impact on the geographical expansion of tropical microalgae, and the “tropicalization” of the Mediterranean Sea in recent years. This fact, together with the Ostreopsis species range expansion in the Mediterranean during the last decade, and their toxicity, could possibly constitute a serious threat to human health by both ciguatera and PLTX intoxications [113, 281].

8.3.1.5 Brevetoxin BTX or PbTx group toxins cause Neurotoxic Shellfish Poisoning (NSP), a syndrome characterized by mainly neurological and gastrointestinal effects. Symptoms and signs of NSP include, for example, nausea, vomiting, diarrhea, parasthesia, cramps, bronchoconstriction, paralysis, seizures and coma. They typically occur within 30 min to 3 h of consuming contaminated shellfish and last for a few days. Persistent symptoms and fatalities have not been reported. Dermal or inhalation exposure can result in irritant effects. BTX-group toxins are metabolized in shellfish and fish and several metabolites of BTX-group toxins have been characterized. Consumers of contaminated shellfish and fish are thus primarily exposed to BTX-group toxin metabolites rather than parent algal BTX-group toxins [110]. BTXs are marine biotoxins that can accumulate in shellfish and fish and are primarily produced by the dinoflagellate Karenia brevis (formerly called Gymnodinium breve and Ptychodiscus brevis). So far, NSP appears to be limited to the Gulf of Mexico, the east coast of the United States and the New Zealand Hauraki Gulf region. To date BTX-group toxins have not been reported in shellfish or fish from Europe [110, 285]. However, the discovery of new BTX-group toxin producing algae such as Chattonella antiqua, Chattonella marina, Fibrocapsa japonica, Heterosigma akashiwo, which have been reported to produce BTX-like toxins [286] and the apparent trend toward expansion of algal bloom distribution suggest that BTX-group toxins are emerging in other regions in the world [110, 197]. Karenia brevis (Gymnodinium breve) populations are found in warm temperate to tropical waters, most regularly from the Gulf of Mexico, off the west coast of Florida. However, K. brevis and K. brevis-like species have also been reported from the West Atlantic, Spain, Greece, Japan and New Zealand [16, 287–289]. In Greece specifically, relatively high abundances of K. brevis (2.7 × 105 cells L–1) were recorded in the urban Thessaloniki Bay, Thermaikos Gulf in June 2017, right in the middle of a mucilage aggregate phenomenon, whereas an extremely high bloom (>1.05 × 107 cells L–1) was observed at the same sampling point one year later [16]. It is known that blooms initiate offshore requiring high salinities (>30‰) and high temperatures [289–292]. K. brevis grows a temperature range between 4 and 33 °C;

8 Human impact in Mediterranean coastal ecosystems and climate change

281

however, its optimal growth range is 22–28 °C. In addition, this organism can live in a salinity of between 25 and 45 psu, with an optimum of 30–34 psu [293]. Similarly, ocean conditions that could arise from climate change, such as lowered pH and/or increased temperature, were shown to promote significant increases in growth rates of Heterosigma akashiwo compared to controls [294–296]. Since the summer of 1994, the Raphidophyceae Fibrocapsa japonica has become the most recurrent blooming species in the Western Adriatic coastal area causing heavy water discoloration, with cells concentrations exceeding 106 cells L–1. During these blooms, other Raphidophyceae (Heterosigma akashiwo, Chattonella spp.) were usually present in very low concentrations; however, these species can exceptionally form nearly monospecific blooms as occurred in July 2011 for H. akashiwo and C. globosa [297]. H. akashiwo has been also reported in other sites of the Mediterranean Sea, for example, in Turkey (Izmir Bay) [298, 299], Egypt [300] and Syria (Lattakia port) [301]. It is thus possible that climate change could trigger blooms/and or toxin production from these species in the future. Currently there are no regulations on BTX-group toxins in shellfish or fish in Europe, as due to the lack of occurrence data on shellfish or fish in Europe, the limited data on acute toxicity and the lack of data on chronic toxicity, the EFSA CONTAM Panel could not comment on the risk associated with the BTX-group toxins in shellfish and fish that could reach the European market [110]. However, some countries in other regions of the world have set action levels or maximum levels for BTX-group toxins in shellfish. In the USA, the action level is 20 mouse units (MUs) per 100 g (0.8 mg BTX-2 eq. kg–1 shellfish) [302]. In New Zealand and Australia the maximum level for BTX-group toxins is also 0.8 mg BTX-2 eq. kg–1 [110, 303, 304]. Moreover, FAO estimated that 2–3 μg BTX eq. kg–1 b.w. is toxic in humans, but the available data on human intoxications do not allow the establishment of an oral ARfD for BTX-group toxins [197, 305].

8.4 Conclusion Climate change has become one of the most critical issues for the future of our planet. It involves significant changes in the variability or average state of the atmosphere, such as related to temperature, precipitation and/or wind patterns, over durations ranging from decades to millions of years. Climate change effects will have implications for food production, food security and food safety. In particular, the safety of feed and food products arising from marine production systems is expected to be affected by climate change, specifically by increased occurrence of phycotoxins (marine biotoxins). The apparent increase in the occurrence of HABs and the recognition that climate changes may be creating a marine environment particularly suited to HAB-

282

Panagiota Katikou

forming species of algae highlight the need for governments to ensure that existing risk management measures are sufficient and are in accordance with international recommendations. Countries are encouraged to implement integrated shellfish and microalgal monitoring programs especially for emerging toxins, as part of Marine Biotoxin Management Plans to strengthen risk management capability and to enhance consumer protection.

References [1]

[2] [3]

[4] [5]

[6] [7] [8]

[9]

[10]

[11]

[12]

[13]

Anderson DM. Toxic algal blooms and red tides: a global perspective. In: Okaichi T, Anderson DM, Nemoto T [ed.] Red Tides: Biology, Environmental Science and Toxicology. New York; Elsevier: 1989. 11–16. Hallegraeff GM. A review of harmful algal blooms and their apparent global increase. Phycologia 1993;32:79–99. Smayda TJ. Novel and nuisance phytoplankton blooms in the sea: evidence for a global epidemic. In: Graneli E, Sundstrom B, Edler L, Anderson DM [ed.] Toxic Marine Phytoplankton. New York; Elsevier: 1990. 29–40. Dale B, Edwards M, Reid PC. Climate Change and Harmful Algae Blooms. In: Granéli E, Turner JT, Heidelberg [ed.] Ecol Stud Volume 189. Berlin; Springer-Verlag: 2006. 367–378. Griffith AW, Gobler CJ. Harmful algal blooms: A climate change co-stressor in marine and freshwater ecosystems. Harmful Algae 2020;91:101590. Doi: https://doi.org/10.1016/j.hal. 2019.03.008. Basterretxea G, Garcés E, Jordi A, Angles S, Masó M. Modulation of nearshore harmful algal blooms by in situ growth rate and water renewal. Mar Ecol Prog Ser 2007;352:53–65. Sellner KG, Doucette GJ, Kirkpatrick GJ. Harmful algal blooms: Causes, impacts and detection. J Ind Microbiol Biotechnol 2003;30:383–406. Moncheva S, Gotsis-Skretas O, Pagou K, Krastev A. Phytoplankton blooms in Black Sea and Mediterranean coastal ecosystems subjected to anthropogenic eutrophication: similarities and differences. Estuar Coast Shelf Sci 2001;53:281–295. Ferrante M, Conti GO, Ledda C, Zuccarello M, Bella F, Fagone E, Copat C, Fiore M, Fallico R. First Results about an Ostreopsis Ovata Monitoring along the Catania Coast (Sicily– Italy). Epidemiology 2009;20:S159. Ferrante M, Sciacca S, Fallico R, Fiore M, Conti GO, Maria F, Venerando R, Caterina L. Harmful Algal Blooms in the Mediterranean Sea: Effects on Human Health. 2013;587(2):587. Doi: 10.4172/scientificreports. Lee KH, Jeong HJ, Lee K, Franks PJS, Seong KA, Lee SY, Lee MJ, Jang SH, Potvin E, Lim AS, Yoon EY, Yoo YD, Kang NS, Kim KY. Effects of warming and eutrophication on coastal phytoplankton production. Harmful Algae 2019;81:106–118. Doi: https://doi.org/10.1016/j. hal.2018.11.017. Nwankwegu AS, Li Y, Huang Y, Wei J, Norgbey E, Sarpong L, Lai Q, Wang K. Harmful algal blooms under changing climate and constantly increasing anthropogenic actions: the review of management implications. 3 Biotech 2019;9:449. Doi: https://doi.org/10.1007/s13205019-1976-1. Wells ML, Karlson B, Wulff A, Kudela R, Trick C, Asnaghi V, Berdalet E, Cochlan W, Davidson K, De Rijcke M, Dutkiewicz S, Hallegraeff G, Flynn KJ, Legrand C, Paerl H, Silke J, Suikkanen S,

8 Human impact in Mediterranean coastal ecosystems and climate change

[14]

[15]

[16]

[17] [18]

[19]

[20]

[21]

[22]

[23]

[24]

[25]

[26]

283

Thompson P, Trainer VL. Future HAB science: Directions and challenges in a changing climate. Harmful Algae 2020;91:101632. Doi: https://doi.org/10.1016/j.hal.2019.101632. Nastasi A. Algal and Jellyfish Blooms in the Mediterranean and Black Sea: a brief review. Istanbul, Turkey; GFCM Workshop on Algal and Jellyfish Blooms in the Mediterranean and BlackSea: 2010. Ferrante M, Ledda C, Cunsolo MA, Fiore M, Fallico R, Sciacca S, Oliveri Conti GM. Harmful algal blooms in Italy and their health effects in the population. Ig Sanita Pubbl 2010;66: 649–658. Genitsaris S, Stefanidou N, Sommer U, Moustaka-Gouni M. Phytoplankton Blooms, Red Tides and Mucilaginous Aggregates in the Urban Thessaloniki Bay, Eastern Mediterranean. Diversity 2019;11:136. Doi: http://dx.doi.org/10.3390/d11080136. Masó M, Garcés E. Harmful microalgae blooms (HAB); problematic and conditions that induce them. Mar Pollut Bull 2006;53:620–630. Bel Hassen M, Drira Z, Hamza A, Ayadi H, Akrout F, Messaoudi S, Issaoui H, Aleya L, Bouaïn A. Phytoplankton dynamics related to water mass properties in the Gulf of Gabes: Ecological implications. J Mar Sys 2009;75:216–226. Silkin VA, Pautova LA, Giordano M, Chasovnikov VK, Vostokov SV, Podymov OI, Pakhomova SV, Moskalenko LV. Drivers of phytoplankton blooms in the northeastern Black Sea. Mar Pollut Bull 2019;138:274–284. Doi: https://doi.org/10.1016/j. marpolbul.2018.11.042. Lionello P, Abrantes F, Congedi L, Dulac F, Gacic M, Gomis D, Goodess C, Hoff H, Kutiel H, Luterbacher J, Planton S, Realea M, Schröder K, Struglia MV, Toretin A, Tsimplis M, Ulbrich U, Xoplaki E. Introduction: Mediterranean Climate: Background Information. In: Lionello P [ed.] The Climate of the Mediterranean Region. From the Past to the Future. Amsterdam; Elsevier: 2012. 502. Caddy JF. Contrast between recent fishery trends and evidence for nutrient enrichment in two large marine ecosystems: the Mediterranean and the Black Seas. In: Sherman K, Alexander LM, Gold BD [ed.] Large marine ecosystems. Stress, mitigation and sustainability. Washington DC; American Association for the Advancement of Science: 1993. Caddy JF, Refk R, Do-Chi T. Productivity estimates for the Mediterranean: evidence of accelerating ecological change. In: FAO, [ed.] Effects of riverine inputs on coastal ecosystems and fisheries resources. Rome; FAO Fisheries Technical Paper 349: 1995. Coll M, Piroddi C, Steenbeek J, Kaschner K, Rais Lasram FB, Aguzzi J, Ballesteros E, Nike Bianchi C, Corbera J, Dailianis T, Danovaro R, Estrada M, Froglia C, Galil BS, Gasol JM, Gertwagen R, Gil J, Guilhaumon F, Kesner-Reyes K, Kitsos MS, Koukouras A, Lampadariou N, Laxamana E, López-fé de la Cuadra CM, Lotze HK, Martin D, Mouillot D, Oro D, Raicevich S, Rius-Barile J, Saiz-Salinas JI, San VC, Somot S, Templado J, Turon X, Vafidis D, Villanueva R, Voultsiadou E. The biodiversity of the Mediterranean Sea: estimates, patterns, and threats. PLoS One 2010;5:11842–11843. Barausse A, Palmeri L. A Comparative Analysis of Trophic Structure and Functioning in LargeScale Mediterranean Marine Ecosystems. In: Goffredo S, Dubinsky Z [ed.] The Mediterranean Sea Its history and present challenges. New York London; New York London Springer: 2014. 421–434. Fabbri E, Dinelli E. Physiological Responses of Marine Animals Towards Adaptation to Climate Changes. In: Goffredo S, Dubinsky Z [ed.] The Mediterranean Sea – Its history and present challenges. New York London: 2014. 401–420. Álvarez-Berastegui D, Hidalgo M, Pilar TM, Reglero P, Aparicio-González A, Ciannelli L, et al. Pelagic seascape ecology for operational fisheries oceanography: modelling and predicting

284

[27] [28]

[29]

[30]

[31] [32]

[33] [34] [35] [36] [37] [38]

[39]

[40] [41] [42] [43]

Panagiota Katikou

spawning distribution of Atlantic bluefin tuna in Western Mediterranean. ICES J Mar Sci 2016;73:1851–1862. Doi: 10.1093/icesjms/fsw041. Schneider A, Tanhua T, Roether W, Steinfeldt R. Changes in ventilation of the Mediterranean Sea during the past 25 year. Ocean Sci 2018;10:1–16. Doi: 10.5194/os-10-1-2014. Tintoré J, Pinardi N, Álvarez-Fanjul E, Aguiar E, Álvarez-Berastegui D, Bajo M, Balbin R, Bozzano R, Nardelli BB, Cardin V, et al. Challenges for Sustained Observing and Forecasting Systems in the Mediterranean Sea. Frontiers Marine Sci 2019;6:568. Doi: 10.3389/fmars.2019.00568. Bianchi CN, Azzola A, Bertolino M, Betti F, Bo M, Cattaneo-Vietti R, Cocito S, Montefalcone M, Morri C, Oprandi A, Peirano A, Bavestrello G. Consequences of the marine climate and ecosystem shift of the 1980–90s on the Ligurian Sea biodiversity (NW Mediterranean). Eur Zool J 2019;86:458–487. Doi: 10.1080/24750263.2019.1687765. Cramer W, Guiot J, Marini K Risks associated to climate and environmental changes in the Mediterranean region: A preliminary assessment by the MedECC Network Science-policy interface – 2019. MedECC – Mediterranean Experts on Climate and Environmental Change. Available online at: https://ufmsecretariat.org/wp-content/uploads/2019/10/MedECCBooklet_EN_WEB.pdf (accessed on 07 December 2019). UNEP/FAO/WHO. Assessment of the state of eutrophication in the Mediterranean Sea, MAP technical series nº 106. UNEP 1996, Athens. Rosa R, Marques A, Nunes ML. Mediterranean Aquaculture in a Changing Climate. In: Goffredo S, Dubinsky Z [ed.]. New York London; The Mediterranean Sea – Its history and present challenges: 2014. 605–616. Sala E, Knowlton N. Global marine biodiversity trends. Annu Rev Environ Resour 2006;31: 93–122. Turley CM. The changing Mediterranean Sea: a sensitive ecosystem? Prog Oceanogr 1999;44: 387–400. Templado J. Future Trends of Mediterranean Biodiversity. In: Goffredo S, Dubinsky Z [ed.] The Mediterranean Sea Its history and present challenges. New York London: 2014. 479–498. Goren M, Aronov A. First record of the Indo-Pacific parrot fish Scarus ghobban in the Eastern Mediterranean. Cybium 2002;26:239–240. Galil BS. The Marine Caravan – The Suez Canal and the Erythrean Invasion. In: Gollasch S, Galil B, Cohen A [ed.]. Springer Netherlands; Bridging Divides, Book 83: 2006. 207–300. Hoffman R, Dubinsky Z. Invasive and Alien Rhodophyta in the Mediterranean and along the Israeli shores. In: Seckbach J, Chapman DJ [ed.] Red algae in genome age. Cellular origin, life in extreme habitats and astrobiology (13th ed). Dordrecht; Springer Publishers: 2010. 45–60. Zenetos A, Gofas S, Verlaque M, Cinar M, Garcia Raso J, Bianchi C, Morri C, Azzurro E, Bilecenoglu M, Froglia C, Siokou I, Violanti D, Sfriso A, San Martin G, Giangrande A, Katagan T, Ballesteros E, Ramos-Esplà A, Mastrototaro F, Ocana O, Zingone A, Gambi M, Streftaris N. Alien species in the Mediterranean Sea. A contribution to the application of European Union’s Marine Strategy Framework Directive (MSFD). Part I. Spatial distribution. Mediterr Mar Sci 2010;11:381–493. Papaconstantinou C. The spreading of Lessepsian fish migrants into the Aegean Sea (Greece). Scientia Marina 1990;54:313–316. Por FD. Lessepsian Migration. The Influx of Red Sea Biota into the Mediterranean by Way of the Suez Canal. Berlin, New York; Springer-Verlag Publishing: 1978. 88–194. Field J. Puffer fish poisoning. J Accid Emerg Med 1998;15:334–336. Bilecenoglu M, Kaya M, Akalin S. Range expansion of silverstripe blaasop, Lagocephalus sceleratus (Gmelin, 1789), to the northern Aegean Sea. Aq Inv 2006;1:289–291.

8 Human impact in Mediterranean coastal ecosystems and climate change

285

[44] Katikou P, Georgantelis D, Sinouris N, Petsi A, Fotaras T. First report on toxicity assessment of the Lessepsian migrant pufferfish Lagocephalus sceleratus (Gmelin, 1789) from European waters (Aegean Sea, Greece). Toxicon 2009;54:50–55. [45] United Nations Environmental Programme/ Mediterranean Action Plan Barcelona Convention (UNEP/MAP) 2017. 2017 Mediterranean Quality Status Report. Available online at: https://www.medqsr.org/sites/default/files/inline-files/2017MedQSR_Online_0.pdf (accessed on 01 December 2019). [46] Lotze HK, Lenihan HS, Bourque BJ, Bradbury RH, Cooke RG, Kay MC, Kidwell SM, Kirby MX, Peterson CH, Jackson JBC. Depletion, degradation, and recovery potential of estuaries and coastal seas. Science 2006;312:1806–1809. [47] Noyes PD, McElwee MK, Miller HD, Clark BW, Van Tiema LA, Walcott KC, Erwin KN, Levin ED. The toxicology of climate change: environmental contaminants in a warming world. Environ Int 2009;35:971–986. [48] Harley CDG, Hughes AR, Hultgren KM, Miner BG, Sorte CJ, Thornber CS, Rodriguez LF, Tomanek L, Williams SL. The impacts of climate change in coastal marine systems. Ecol Lett 2006;9:228–241. [49] Michaelidis B, Pörtner HO, Sokolova I, Tomanek L. Advances in Predicting the Impacts of Global Warming on the Mussels Mytilus galloprovincialis in the Mediterranean Sea. In: Goffredo S, Dubinsky Z [ed.] The Mediterranean Sea Its history and present challenges. New York London: 2014. 319–340. [50] Anderson DM, Cembella AD, Hallegraeff GM. Progress in Understanding Harmful Algal Blooms: Paradigm Shifts and New Technologies for Research, Monitoring, and Management. Annu Rev Mar Sci 2012;4,143–76. [51] Hoagland P, Anderson DM, Kaoru Y, White A. The economic effects of harmful algal blooms in the United States: estimates assessment issues, and information needs. Estuaries 2002;25: 819–837. [52] Anderson DM, Kaoru Y, White AW Economic Impacts from Harmful Algal Blooms (HABs) in the United States. Technical report WHOI-2000–11, 2000, Woods Hole Oceanographic Institute, Massachusetts. Available online at: https://www.whoi.edu/cms/files/Economics_report_ 18564_23050.pdf (accessed on 01 December 2019). [53] EEA. Priority issues in the Mediterranean environment. Report no 4/2006European Enviroment Agency.ISBN 92–9167–812-0. ISSN 1725–9177. Copenhagen 2006. [54] Hoagland P, Scatasta S. The economic effects of harmful algal blooms. In: Graneli E, Tyrner J [ed.] Ecology of Harmful Algae. New York; Springer-Verlag: 2006. 391–402. [55] Davidson K, Gowen RJ, Harrison PJ, Fleming LE. Hoagland P and Moschonas G. Anthropogenic nutrients and harmful algae in coastal waters. J Environ Manage 2014;146:206–216. [56] Sanseverino I, Conduto D, Pozzoli L, Dobricic S, Lettieri T Algal bloom and its economic impact, 2016. JRC Technical Reports, Publications Office of the European Union, EUR 27905 EN; DOI: 10.2788/660478. [57] Wilcove DS, Rothstein D, Dubow J, Phillips A, Losos E. Quantifying threats to imperiled species in the United States. Bioscience 1998;48:607–615. [58] Dukes JS, Mooney HA. Does global change increase the success of biological invaders? Trends Ecol Evol 1999;14:135–139. [59] Pimental D, Lach L, Zuniga R, Morrison D. Environmental and economic costs of nonindigenous species in the United States. Bioscience 2000;50:53–65. [60] Davis MA, Thompson K, Grime JP, Charles S. Elton and the dissociation of invasion ecology from the rest of ecology. Divers Distrib 2001;7:97–102.

286

[61]

[62]

[63] [64]

[65]

[66] [67]

[68]

[69] [70] [71]

[72] [73] [74]

[75] [76]

[77]

Panagiota Katikou

Streftaris N, Zenetos A, Papathanassiou E. Globalisation in marine ecosystems: the story of non-indigenous marine species across European seas. Oceanogr Mar Biol Annu Rev 2005;43: 419–453. EPA – United States Environmental Protection Agency, Aquatic nuisance species. Annual Report 2000–2001, The Western Regional Panel on Aquatic Nuisance Species, 2001, Available online at: http://web.archive.org/web/20170207062911/http://www.fws.gov/answest/annual00_01. pdf (accessed on 14/ 12/2019). Streftaris N, Zenetos A. Alien marine species in the Mediterranean – the 100 ‘worst invasives’ and their impact. Mediterr Mar Sci 2006;7:87–118. IPCC. Climate change 2007: the physical science basis: contribution of working group I to the fourth assessment report of the intergovernmental panel on climate change. Solomon S, Qin D, Manning M, Chen Z, Marquis M, Averyt KB, Tignor M, Miller HL [ed.]. Cambridge, United Kingdom and New York, NY, USA; Cambridge University Press: 2007. 996. Giannakopoulos C, Bindi M, Moriondo M, LeSager P, Tin T. Climate change impacts in the Mediterranean resulting from a 2°C global temperature rise. WWF Report for a Living Planet. Gland; The Global Conservation Organization: 2005. Tourre Y, Van Grunderbeeck P, Allal H et al Climate change and energy in the Mediterranean. Plan Bleu Regional Activity Center 2008, Sophia Antipolis. Cramer W, Guiot J, Fader M, Garrabou J, Gattuso J-P, Iglesias A, Lange MA, Lionello P, Llasat MC, Paz S, Peñuelas J, Snoussi M, Toreti A, Tsimplis MN, Xoplaki E. Climate change and interconnected risks to sustainable development in the Mediterranean. Nature Clim Change 2018;8:972–980. Doi: 10.1038/s41558-018-0299-2. Boxall ABA, Hardy A, Beulke S, Boucard T, Burgin L, Falloon PD, Haygarth PM, Hutchinson T, Kovats RS, Leonardi G, Levy LS, Nichols G, Parsons SA, Potts L, Stone D, Topp E, Turley DB, Walsh K, Wellington EMH, Williams RJ. Impacts of climate change on indirect human exposure to pathogens and chemicals from agriculture. Environ Health Persp 2009;117: 508–514. Marques A, Nunes ML, Moore SK, Strom MS. Climate change and seafood safety: Human health implications. Food Res Int 2010;43:1766–1779. Falkowski PG, Barber RT, Smetacek V. Biogeochemical controls and feedbacks on ocean primary production. Science 1998;281:200–207. Calvo E, Simó R, Coma R, Ribes M, Pascual J, Sabatés A, Gili JM, Pelejero C. Effects of climate change on Mediterranean marine ecosystems: the case of the Catalan Sea. Clim Res 2011;50: 1–29. Helmuth B. From cells to coastlines: how can we use physiology to forecast the impacts of climate change? J Exp Biol 2009;212:753–760. Monaco CJ, Helmuth B. Tipping points, thresholds and the keystone role of physiology in marine climate change research. Adv Mar Biol 2011;60:123–160. Vargas-Yáñez M, Moya F, García-Martínez MC, Zunino P, Plaza F, Salat J, Pascual J, LópezJurado JL, Serra M. Climate change in the western Mediterranean Sea 1900–2008. J Mar Syst 2010;82,171–6. Bianchi CN. Biodiversity issues for the forthcoming tropical Mediterranean Sea. Hydrobiologia 2007;580:7–21. Sabrah MM, El-Ganainy AA, Zaky MA. Biology and toxicity of the pufferfish Lagocephalus sceleratus (Gmelin, 1789) from the gulf of Suez. Egyptian Journal of Aquatic Research 2006;32:283–297. Garcés E, Masó M, Vila M, Camp J. HABs events in the Mediterranean Sea: are they increasing? A case study the last decade in the NW Mediterranean and the genus Alexandrium. Harmful Algal News 2000;20:1–11.

8 Human impact in Mediterranean coastal ecosystems and climate change

287

[78] Aligizaki K, Katikou P, Nikolaidis G. Toxic benthic dinoflagellates and potential risk in the mediterranean Sea. Nantes; International Conference on Molluscan Shellfish Safety: 2009. [79] Faimali M, Giussani V, Piazza V, Garaventa F, Corrà C, Asnaghi V, Privitera D, Gallus L, Cattaneo-Vietti R, Mangialajo L, Chiantore M. Toxic effects of harmful benthic dinoflegellate Ostreopsis ovata on invertebrate and vertebrate marine organisms. Mar Environ Res 2012;76: 97–107. [80] Hallegraeff GM. Harmful Algal Blooms, Coastal Eutrophication and climate change. Biol Mar Mediterr 2008;15:6–15. [81] Caillaud A, de la Iglesia P, Darius HT, Pauillac S, Aligizaki K, Fraga S, Chinain M, Diogene J. Update on methodologies available for ciguatoxin determination: Perspectives to confront the onset of ciguatera fish poisoning in Europe. Mar Drugs 2010;8:1838–1907. [82] Nezan E, Chomerat N. Vulcanodinium rugosum gen. nov., sp. nov. (Dinophyceae): A new marine dinoflagellate from the French Mediterranean coast. Cryptogam Algol 2011;32:3–18. [83] Hess P, Abadie E, Hervé F, Berteaux T, Séchet V, Aráoz R, Molgó J, Zakarian A, Sibat M, Rundberget T, Miles CO, Amzil Z. Pinnatoxin G is responsible for atypical toxicity in mussels (Mytilus galloprovincialis) and clams (Venerupis decussata) from Ingril, a French Mediterranean lagoon. Toxicon 2013;75:16–26. [84] García-Altares M, Casanova A, Bane V, Diogène J, Furey A, de la Iglesia P. Confirmation of Pinnatoxins and Spirolides in Shellfish and Passive Samplers from Catalonia (Spain) by Liquid Chromatography Coupled with Triple Quadrupole and High-Resolution Hybrid Tandem Mass Spectrometry. Mar Drugs 2014;12:3706–3732. [85] Rambla-Alegre M, Miles CO, de la Iglesia P, Fernandez-Tejedor M, Jacobs S, Sioen I, Verbeke W, Samdal IA, Sandvik M, Barbosa V, Tediosi A, Madorran E, Granby K, Kotterman M, Calis T, Diogene J. Occurrence of cyclic imines in European commercial seafood and consumers risk assessment. Environ Res 2019;161:392–398. Doi: https://doi.org/10.1016/j. envres.2017.11.028. [86] Goffart A, Hecq JH, Legendre L. Changes in the development of the winterspring phytoplankton bloom in the Bay of Calvi (NW Mediterranean) over the last two decades: a response to changing climate? Mar Ecol Progr Ser 2002;236:45–60. [87] Stachowich JJ, Terwin JR, Whitlatch RB, Osman RW. Linking climate change and biological invasions: ocean warming facilitates non-indigenous species invasion. Proc Nat Acad Sc USA 2002;99:15497–15500. [88] Morton SL, Norris DR, Bomber JW. Effect of temperature, salinity and light-intensity on the growth and seasonality of toxic dinoflagellates associated with ciguatera. J Exp Mar Biol Ecol 1992;157:79–90. [89] Hales S, Weinstein P, Woodward A. Ciguatera (fish poisoning), El Niño, and Pacific sea surface temperatures. Ecosyst Health 1999;5:20–25. [90] Chateau-Degat ML, Chinain M, Cerf N, Gingrasa S, Hubertc B, Dewailly É. Seawater temperature, Gambierdiscus spp. Variability and incidence of ciguatera poisoning in French Polynesia. Harmful Algae 2005;4:1053–1062. [91] Rhodes L. World-wide occurrence of the toxic dinoflagellate genus Ostreopsis Schmidt. Toxicon 2011;57:400–407. [92] Shears NT, Ross PM. Blooms of benthic dinoflagellates of the genus Ostreopsis; an increasing and ecologically important phenomenon on temperate reefs in New Zealand and worldwide. Harmful Algae 2009;8:916–925. [93] Granéli E, Vidyarathna NK, Funari EPRT. Cumaranatunga, Raffaeli Scenati. Can increases in temperature stimulate blooms of the toxic benthic dinoflagellate Ostreopsis ovata? Harmful Algae 2011;10:165–172.

288

Panagiota Katikou

[94] Vlamis and Katikou. Ecobiology and Geographical Distribution of Potentially Toxic Marine Dinoflagellates. In: LM B [ed.] Seafood and Freshwater Toxins: Pharmacology, Physiology, and Detection (3rd edition). CRC Press, Taylor and Francis group: 2014. 569–625. [95] Accoroni S, Tartaglione L, Dello Iacovo E, Pichierri S, Marini M, Campanelli A, Dell’Aversano C, Totti C. Influence of environmental factors on the toxin production of Ostreopsis cf. ovata during bloom events. Mar Pollut Bull 2017;123:261–268. Doi: https://doi.org/10.1016/j.marpolbul.2017.08.049. [96] Peyton J, Martinou AF, Pescott OL, et al. Horizon scanning for invasive alien species with the potential to threaten biodiversity and human health on a Mediterranean island. Biol Invasions 2019;21:2107–2125. Doi: https://doi.org/10.1007/s10530-019-01961-7. [97] Zenetos A, Gratsia E, Cardoso A, Tsiamis K. Time lags in reporting of biological invasions: the case of Mediterranean Sea. Mediterr Mar Sci 2019;20:469–475. Doi: http://dx.doi.org/ 10.12681/mms.20716. [98] EU Directive 91/492/EEC. Council Directive 91/492/EEC of 15 July 1991, laying down the conditions for the production and the placing of the market of live bivalve molluscs. Official J Europ Communities 261:1–14. [99] European Food Safety Authority. Scientific Opinion: Marine Biotoxins in Shellfish – Okadaic acid and analogues. EFSA J 2008;589:1–62. [100] European Food Safety Authority. Scientific Opinion: Marine Biotoxins in Shellfish – Yessotoxin group. EFSA J 2008;907:1–62. [101] European Food Safety Authority. Scientific Opinion: Marine Biotoxins in Shellfish – Azaspiracid group. EFSA J 2008;723:1–52. [102] European Food Safety Authority. Scientific Opinion: Marine Biotoxins in Shellfish – Saxitoxin group. EFSA J 2009;1019:1–76. [103] European Food Safety Authority. Scientific Opinion: Marine Biotoxins in Shellfish – Pectenotoxin group. EFSA J 2009;1109:1–47. [104] European Food Safety Authority. Scientific Opinion: Marine Biotoxins in Shellfish – Domoic acid. EFSA J 2009;1181:1–61. [105] Van Dolah FM. Diversity of Marine and Freshwater Algal Toxins. In: Botana LM [ed] Seafood and Freshwater Toxins: Pharmacology, Physiology and Detection. Boca Raton, FL, USA; CRC Press (Taylor and Francis Group): 2000. 19–43. [106] Yasumoto T. Historic considerations regarding seafood safety. In: LM B [ed.] Seafood and Freshwater Toxins: Pharmacology, Physiology and Detection. Boca Raton, FL, USA; CRC Press (Taylor and Francis Group): 2000. 1–17. [107] European Food Safety Authority. Scientific Opinion: Marine Biotoxins in Shellfish – Palytoxin group. EFSA J 2009:1393. [108] European Food Safety Authority. Scientific Opinion: Marine Biotoxins in Shellfish – Emerging toxins: Ciguatoxin group. EFSA J 2010:1627. [109] European Food Safety Authority. Scientific Opinion: Marine Biotoxins in Shellfish – Cyclic imines. EFSA J 2010:1628. [110] European Food Safety Authority. Scientific Opinion: Marine Biotoxins in Shellfish – Emerging toxin: Brevetoxin group. EFSA J 2010:1677. [111] European Food Safety Authority. Scientific opinion on the risks for public health related to the presence of tetrodotoxin (TTX) and TTX analogues in marine bivalves and gastropods. EFSA J 2017;15:4752. Doi: 10.2903/j.efsa.2017.4752. [112] Campbell K, Vilariño N, Botana LM, Elliott CT. A European perspective on progress in moving away from the mouse bioassay for marine-toxin analysis. Trends Anal Chem 2011;30: 239–253.

8 Human impact in Mediterranean coastal ecosystems and climate change

289

[113] Estevez P, Castro D, Pequeño-Valtierra A, Giraldez J, Gago-Martinez A. Emerging Marine Biotoxins in Seafood from European Coasts: Incidence and Analytical Challenges. Foods 2019;8:149. Doi: 10.3390/foods8050149. [114] Yotsu-Yamashita M, Sugimoto A, Takai A, Yasumoto T. Effects of specific modifications of several hydroxyls of tetrodotoxin on its affinity to rat brain membrane. J Pharmacol Exp Ther 1999;289:1688–1696. [115] Bane V, Lehane M, Dikshit M, O’Riordan A, Furey A. Tetrodotoxin: Chemistry, toxicity, source, distribution and detection. Toxins 2014;6:693–755. [116] Food and Agriculture Organization of the United Nations/World Health Organization, (FAO/ WHO): Technical paper on Toxicity Equivalency Factors for Marine Biotoxins Associated with Bivalve Molluscs. Rome 2016. pp. 1–133. Available online at: http://www.fao.org/3/a-i5970e. pdf (accessed on 8 December 2019). [117] Pratheepa V, Vasconcelos V. Microbial diversity associated with tetrodotoxin production in marine organisms. Environ Toxicol Pharmacol 2013;36:1046–1054. [118] Vlamis A, Katikou P, Rodriguez I, Rey V, Alfonso A, Papazachariou A, Zacharaki T, Botana AM, Botana LM. First detection of tetrodotoxin in Greek shellfish by UPLC-MS/MS potentially linked to the presence of the dinoflagellate Prorocentrum minimum. Toxins 2015;7: 1779–1807. Doi: https://doi.org/10.3390/toxins7051779. [119] Rodriguez I, Alfonso A, Alonso E, Rubiolo JA, Roel M, Vlamis A, Katikou P, Jackson SA, Menon ML, Dobson A, Botana LM. The association of bacterial C9-based TTX-like compounds with Prorocentrum minimum opens new uncertainties about shellfish seafood safety. Sci Rep 2017;7(40880). Doi: https://doi.org/10.1038/srep40880. [120] de Sousa ML. Occurrence of Tetrodotoxin Producing Bacteria on Marine Gastropods of the Northern Coast of Portugal. Portugal; University of Porto: 2011. 80. [121] Rodríguez P, Alfonso A, Otero P, Katikou P, Georgantelis D, Botana LM. Liquid chromatography mass spectrometry method to detect Tetrodotoxin and Its analogues in the puffer fish Lagocephalus sceleratus (Gmelin, 1789) from European waters. Food Chem 2012;132:1109–1117. [122] Nzoughet JK, Campbell K, Barnes P, Cooper KM, Chevallier OP, Elliott CT. Comparison of sample preparation methods, validation of an UPLC–MS/MS procedure for the quantification of tetrodotoxin present in marine gastropods and analysis of pufferfish. Food Chem 2013;136:1584–1589. [123] Kosker AR, Özogul F, Durmus M, Ucar Y, Ayas D, Regenstein JM, Özogul Y. Tetrodotoxin levels in pufferfish (Lagocephalus sceleratus) caught in the Northeastern Mediterranean Sea. Food Chem 2016;210:332–337. Doi: https://doi.org/10.1016/j.foodchem.2016.04.122. [124] Kosker AR, Özogul F, Durmus M, Ucar Y, Ayas D, Šimat V, Özogul Y. First report on TTX levels of the yellow spotted pufferfish (Torquigener flavimaculosus) in the Mediterranean Sea. Toxicon 2018;148:101–106. Doi: https://doi.org/10.1016/j.toxicon.2018.04.018. [125] Kosker AR, Özogul F, Ayas D, Durmus M, Ucar Y, Regenstein JM, Özogul Y. Tetrodotoxin levels of three pufferfish species (Lagocephalus sp.) caught in the North-Eastern Mediterranean sea. Chemosphere 2019;219:95–99. Doi: https://doi.org/10.1016/j. chemosphere.2018.12.010. [126] Rambla-Alegre M, Leonardo S, Barguil Y, Flores C, Caixach J, Campbell K, Elliott CT, Maillaud C, Boundy MJ, Harwood DT, Campàs M, Diogène J. Rapid screening and multi-toxin profile confirmation of tetrodotoxins and analogues in human body fluids derived from a puffer fish poisoning incident in New Caledonia. Food Chem Toxicol 2018;112:188–193. Doi: https://doi.org/10.1016/j.fct.2017.12.039. [127] Cassiday L. First report of TTX in a European trumpet shell. Anal Chem 2008;80:5675–5675.

290

Panagiota Katikou

[128] Rodriguez P, Alfonso A, Vale C, Alfonso C, Vale P, Tellez A, Botana LM. First toxicity report of tetrodotoxin and 5,6,11-trideoxyTTX in the trumpet shell Charonia lampas lampas in Europe. Anal Chem 2008;80:5622–5629. [129] Pigozzi S, Ceredi A, Pompei M, Bordin P, Bille L, Dell’Aversano C, Tartaglione L, Sidari L, Zanolin B, Cacitti A, Pelagatti L, Palei M, Ricci M, Milandri A First detection of tetrodotoxin in shellfish from Northern Italy, Book of abstracts of 6th International Symposium Marine and Freshwater Toxins Analysis, Baiona, Spain, October 22-25, 2017; Abstract no. P8, pp. 60–61 and relevant poster. [130] EURL-NRL. Italy NRL for Marine Biotoxins, Summary of activities 2018. In: Minutes of 21st Meeting of EU National Reference Laboratories (EU-NRL) on Marine Biotoxins. 2018, Vigo, Spain, 18–19 October 2018. [131] Dell’Aversano C, Tartaglione L, Polito G, Dean K, Giacobbe M, Casabianca S, Capellacci S, Penna A, Turner AD. First detection of tetrodotoxin and high levels of paralytic shellfish poisoning toxins in shellfish from Sicily (Italy) by three different analytical methods. Chemosphere 2019;215:881–892. Doi: https://doi.org/10.1016/j.chemosphere.2018.12.010. [132] Turner AD, Powell A, Schofield A, Lees DN, Baker-Austin C. Detection of the pufferfish toxin tetrodotoxin in European bivalves, England, 2013 to 2014. Euro Surveill 2015;20:2. Doi: https://doi.org/10.2807/1560-7917.ES2015.20.2.21009. [133] Turner AD, Dhanji-Rapkova M, Coates L, Bickerstaff L, Milligan S, O’Neill A, Faulkner D, McEneny H, Baker-Austin C, Lees DN, Algoet M. Detection of tetrodotoxin shellfish poisoning (TSP) toxins and causative factors in bivalve molluscs from the UK. Mar Drugs 2017;15:277. Doi: https://doi.org/10.3390/md15090277. [134] Gerssen A, Bovee T, Klijnstra M, Poelman M, Portier L, Hoogenboom R. First report on the occurrence of tetrodotoxins in bivalve mollusks in the Netherlands. Toxins 2018;10:450. Doi: https://doi.org/10.3390/toxins10110450. [135] Leao JM, Lozano-Leon A, Giraldez J, Vilarino O, Gago-Martinez A. Preliminary results on the evaluation of the occurrence of tetrodotoxin associated to marine Vibrio spp. in bivalves from the Galician rias (northwest of Spain). Mar Drugs 2018;16:81. Doi: https://doi.org/10.3390/ md16030081. [136] Gago-Martinez A Riesgo emergente de la TTX en moluscos bivalvos de la UE. II Reunión de la Red Nacional de Riesgos Emergentes, Agencia Española de Consumo, Seguridad Alimentaria y Nutrición (AECOSAN), 2018, Madrid, Spain, 25 September 2018. Available online at: http://www.aecosan.msssi.gob.es/AECOSAN/docs/documentos/noticias/2018/2_TTX_molus cos_bivalvos_UE.pdf (accessed on 8 December 2019). [137] Rodrigues SM, Pinto EP, Oliveira P, Pedro S, Costa PR. Evaluation of the Occurrence of Tetrodotoxin in Bivalve Mollusks from the Portuguese Coast. J Mar Sci Eng 2019;7:232. [138] Silva M, Rodríguez I, Barreiro A, Kaufmann M, Neto AI, Hassouani M, Sabour B, Alfonso A, Botana LM, Vasconcelos V. Tetrodotoxins Occurrence in Non-Traditional Vectors of the North Atlantic Waters (Portuguese Maritime Territory, and Morocco Coast). Toxins 2019;11:306. [139] Katikou P. Public Health Risks Associated with Tetrodotoxin and Its Analogues in European Waters: Recent Advances after The EFSA Scientific Opinion. Toxins 2019;11:240. [140] Lasram FB, Mouillot D. Increasing southern invasion enhances congruence between endemic and exotic Mediterranean fish fauna. Biol Invasions 2009;11:697–711. [141] Akyol OV, Unal CT, Bilecenoglu M. First confirmed report of Lagocephalus sceleratus (Gmelin, 1789) in the Mediterranean sea. J Fish Biol 2005;66:1183–1186. [142] Bentur Y, Ashkar J, Lurie Y, Levy Y, Azzam ZS, Litmanovich M, Golik M, Gurevych B, Golani D, Eisenman A. Lessepsian migration and tetrodotoxin poisoning due to Lagocephalus sceleratus in the eastern Mediterranean. Toxicon 2008;52:964–968.

8 Human impact in Mediterranean coastal ecosystems and climate change

291

[143] Mouneimne N. Liste des poissons de la cote du Liban (Mediterranee orientale). Cybium (Nouv Ser ) 1977;1:37–66. [144] Saoudi M, Abdelmouleh A, El Feki A. Tetrodotoxin: A potent marine toxin. Toxin Rev 2010;29: 60–70. [145] Abd Rabou AN. On the occurrence and health risks of the Silver-cheeked Toadfish (Lagocephalus sceleratus Gmelin, 1789) in the marine ecosystem of the Gaza Strip, Palestine. Biodiversitas 2019;20:2618–2625. Doi: 10.13057/biodiv/d200926. [146] Kleitou P, Kalogirou S, Marmara D, Giovos I. Coryphaena hippurus: A potential predator of Lagocephalus sceleratus in the Mediterranean Sea. Int J Fish Aquat Stud 2018;6:93–95. [147] Matsumoto T, Nagashima Y, Kusuhara H, Sugiyama Y, Ishizaki S, Shimakura K, Shiomi K. Involvement of carrier-mediated transport system in uptake of tetrodotoxin into liver tissue slices of puffer fish Takifugu rubripes. Toxicon 2007;50:173–179. [148] Lin SJ, Hwang DF, Shao KT, Jeng SS. Toxicity of Taiwanese gobies. Fisheries Sci 2000;66(3): 547–552. [149] Brillantes S, Samosorn W, FaknoiI S, Oshima Y. Toxicity of puffers landed and marketed in Thailand. Fisheries Sci 2003;69:1224–1230. [150] El-Sayed M, Yacout GA, El-Samra M, Ali A, Kotb SM. Toxicity of the Red Sea pufferfish Pleuranacanthus sceleratus “El-Karad”. Ecotoxicol Environ Saf 2003;56:367–372. [151] Peristeraki P, Lazarakis G, Tserpes G. First results on the maturity of the lessepsian migrant Lagocephalus sceleratus (GMELIN 1789) in the eastern Mediterranean Sea. Rapp Comm int Mer Médit (Rapport Commission International pour l’exploration scientifique de la Mer Mediterranee) 2010;39:628. [152] Sugita H, Iwata J, Miyajima C, Kubo T, Noguchi T, Hashimoto K, Deguchi Y. Changes in microflora of a puffer fish Fugu niphobles with different water temperatures. Mar Biol 1989;101:299–304. [153] Danovar R, Fonda Umani S, Pusceddu A. Climate change and the potential spreading of marine mucilage and microbial pathogens in the Mediterranean sea. PLoS ONE 2009; 4:1–8. [154] Awada A, Chalhoub V, Awada L, Yasbeck P. Coma profond aréactif réversible après intoxication par des abats d’un poison méditerranéen. Rev Neurol 2010;166:337–340. [155] Petrou A, Isaias E, Chrysanthou K, Scarcella G Fishing yields of Lagocephalus sceleratus in Cyprus. Proceedings of the Tenth International Conference on the Mediterranean Coastal Environment 2011, MEDCOAST 11, Özhan E (ed.), 25–29 October 2011, Rhodes, Greece, pp. 445–452. [156] Beköz AB, Beköz S, Yimaz E, Tüzün S, Beköz U. Consequences of the increasing prevalence of the poisonous Lagocephalus sceleratus in Southern Turkey. Emerg Med J 2013;30: 954–955. [157] Jribi I, Bradai N. First record of the Lessepsian migrant species Lagocephalus sceleratus (Gmelin, 1789) (Actinopterygii: Tetraodontidae) in the Central Mediterranean. BioInvasions Records 2012;1:49–52. [158] Milazzo M, Azzurro E, Badalamenti F. On the occurrence of the silverstripe blaasop Lagocephalus sceleratus (Gmelin, 1789) along the Libyan coast. BioInvasion Records 2012;1: 125–127. [159] Azzurro E, Castriota L, Falautano M, Giardina F, Andaloro F. The silver-cheeked toadfish Lagocephalus sceleratus (Gmelin, 1789) reaches Italian waters. J Appl Ichthyol 2014:1–3. [160] Azzurro E, Castriota L, Falautano M, Bariche M, Broglio E, Andaloro F. New records of the silver-cheeked toadfish Lagocephalus sceleratus (Gmelin, 1789) in the Tyrrhenian and Ionian

292

[161]

[162] [163]

[164]

[165] [166] [167] [168]

[169]

[170]

[171]

[172] [173] [174]

[175]

[176]

Panagiota Katikou

Seas: early detection and participatory monitoring in practice. Bioinvasions Rec 2016;5: 295–299. Doi: http://dx.doi.org/10.3391/bir.2016.5.4.16. Deidun A, Fenech-Farrugia A, Castriota L, Falautano M, Azzurro E, Andaloro F. First record of the silver-cheeked toadfish Lagocephalus sceleratus (Gmelin, 1789) from Malta. Bioinvasions Rec 4:139–142. Doi: http://dx.doi.org/10.3391/bir.2015.4.2.11. Sulić-Šprem J, Dobroslavić T, Kožul V, Kuzman A, Dulčić J. First record of Lagocephalus sceleratus in the Adriatic Sea (Croatian coast), a Lessepsian migrant. Cybium 2014;38:147–148. Kara MH, Ben Lamine E, Francour P. Range expansion of an invasive pufferfish, Lagocephalus sceleratus (Actinopterygii: Tetraodontiformes: Tetraodontidae), to the southwestern Mediterranean. Acta Ichthyol Piscat 2015;45:103–108. Izquierdo-Muñoz A, Izquierdo-Gomez D First record of Lagocephalus sceleratus (Gmelin, 1789) (Actinopterygii, Tetraodontidae) on the Mediterranean Spanish coast, pp. 686–687, In: New Mediterranean Biodiversity Records (October, 2014), Katsanevakis S et al. [Eds.]. Mediterr. Mar. Sci. 2014, 15, 675–695. Schmitt C, Haro LD. Clinical Marine Toxicology: A European Perspective for Clinical Toxicologists and Poison Centers. Toxins (Basel) 2013;5:1343–1352. Botana LM. Toxicological Perspective on Climate Change: Aquatic Toxins. Chem Res Toxicol 2016;29:619–625. Doi: 10.1021/acs.chemrestox.6b00020. Botana LM, Alfonso A, Rodriguez I, Botana AM, Louzao MC, Vieytes MR. How safe is safe for marine toxins monitoring? Toxins 2016;8:208. Doi: https://doi.org/10.3390/toxins8070208. European Union. Corrigendum to Regulation (EC) 853/2004 of the European Parliament and of the Council of 29 April 2004 laying down specific hygiene rules for food of animal origin. Off J Eur Union 2004;226:22–82. European Union. Corrigendum to Regulation (EC) 854/2004 of the European Parliament and of the Council of 29 April 2004 laying down specific rules for the organisation of official controls on products of animal origin intended for human consumption. Off J Eur Union 2004;226:83–127. European Union. Commission Implementing Regulation (EU) 2019/627 of 15 March 2019 laying down uniform practical arrangements for the performance of official controls on products of animal origin intended for human consumption in accordance with Regulation (EU) 2017/625 of the European Parliament and of the Council and amending Commission Regulation (EC) No 2074/2005 as regards official controls. Off J Eur Union 2019;131: 51–100. EU Commission. Notification Detail: Policy Guideline on tetrodotoxin in live bivalve molluscs. Notification Number: 2016/175/NL (Netherlands). TRIS Database 2016, Available online at: http://ec.europa.eu/growth/tools-databases/tris/en/search/?trisaction=search.detail& year=2016&num=175 (accessed on 11 December 2019). Moore RE, Bartolini G. Structure of palytoxin. J Am Chem Soc 1981;103:2491–2494. Deeds JR, Schwartz MD. Human risk associated with palytoxin exposure. Toxicon 2010;56: 150–162. Aligizaki K, Katikou P, Nikolaidis G, Panou A. First episode of shellfish contamination by palytoxin-like compounds from Ostreopsis species (Aegean Sea, Greece). Toxicon 2008;51: 418–427. Ciminiello P, Dell’Aversano C, Fattorusso E, Forino M, Tartaglione L, Grillo C, Melchiorre N. Putative palytoxin and its new analogue, ovatoxin-a, in Ostreopsis ovate collected along the Ligurian coasts during the 2006 toxic outbreak. J Am Soc Mass Spectrom 2008;19: 111–120. Tartaglione L, Mazzeo A, Dell’Aversano C, Forino M, Giussani V, Capellacci S, Penna A, Asnaghi V, Faimali M, Chiantore M, Yasumoto T, Ciminiello P. Chemical, molecular, and eco-

8 Human impact in Mediterranean coastal ecosystems and climate change

[177]

[178]

[179]

[180]

[181]

[182]

[183]

[184]

[185]

[186] [187]

[188]

293

toxicological investigation of Ostreopsis sp. from Cyprus Island: structural insights into four new ovatoxins by LC-HRMS/MS. Anal Bioanal Chem 2016;408:915–932. Doi: https://doi.org/ 10.1007/s00216-015-9183-3. Ninčević Gladan Ž, Arapov J, Casabianca S, Penna A, Honsell G, Brovedani V, Pelin M, Tartaglione L, Sosa S, Dell’Aversano C, Tubaro A, Žuljević A, Grbec B, Čavar M, Bužančić M, Bakrač A, Skejić S. Massive Occurrence of the Harmful Benthic Dinoflagellate Ostreopsis cf. ovata in the Eastern Adriatic Sea. Toxins 2019;11:300. Doi: https://doi.org/10.3390/ toxins11050300. Accoroni S, Romagnoli T, Penna A, Capellacci S, Ciminiello P, Dell’Aversano C, Tartaglione L, Abboud–Abi Saab M, Giussani V, Asnaghi V, Chiantore M, Totti C. Ostreopsis fattorussoi sp. nov. (Dinophyceae), a new benthic toxic Ostreopsis species from the eastern Mediterranean Sea. J Phycol 2016;52:1064–1084. Doi: 10.1111/jpy.12464. García-Altares M, Tartaglione L, Dell’Aversano C, Carnicer O, de la Iglesia P, Forino M, Diogène J, Ciminiello P. The novel ovatoxin-g and isobaric palytoxin (so far referred to as putative palytoxin) from Ostreopsis cf. ovata (NW Mediterranean Sea): structural insights by LC-high resolution MSn. Anal Bioanal Chem 2015;407:1191–1204. Doi: https://doi.org/ 10.1007/s00216-014-8338-y. Ciminiello P, Dell’Aversano C, Dello Iacovo E, Fattorusso E, Forino M, Grauso L, Tartaglione L, Guerrini F, Pistocchi R. Complex palytoxin-like profile of Ostreopsis ovata. Identification of four new ovatoxins by high-resolution liquid chromatography/mass spectrometry. Rapid Commun Mass Spectrom 2010;24:2735–2744. Ciminiello P, Dell’Aversano C, Dello Iacovo E, Fattorusso E, Forino M, Tartaglione L, Battocchi C, Crinelli R, Carloni E, Magnani M, Penna A. Unique toxin profile of a Mediterranean Ostreopsis cf. ovata Strain: HR LC-MSn characterization of ovatoxin-f, a new palytoxin congener. Chem Res Toxicol 2012;25:1243–1252. Brissard C, Hervé F, Sibat M, Séchet V, Hess P, Amzil Z, Herrenknecht C. Characterization of ovatoxin-h, a new ovatoxin analog, and evaluation of chromatographic columns for ovatoxin analysis and purification. J Chromatogr A 2015;1388:87–101. Doi: https://doi.org/10.1016/j. chroma.2015.02.015. Tartaglione L, Dello Iacovo E, Mazzeo A, Casabianca S, Ciminiello P, Penna A, Dell’Aversano C. Variability in Toxin Profiles of the Mediterranean Ostreopsis cf. ovata and in Structural Features of the Produced Ovatoxins. Environ Sci Technol 2017;51:13920–13928. Doi: https://doi.org/10.1021/acs.est.7b03827. David H, Moita MT, Aitor L-M, Silva A, Marcos M, de Pablo H, Orive E. First bloom of Ostreopsis cf. ovata in the continental Portuguese coast. Harmful Algae News 2012;45: 12–13. Santos M, Oliveira PB, Moita MT, David H, Caeiro MF, Zingone A, Amorim A, Silva A. Ocurrence of Ostreopsis in two temperate coastal bays (SW iberia): Insights from the plankton. Harmful Algae 2019;86:20–36. Doi: https://doi.org/10.1016/j.hal.2019.03.003. Vasconcelos V. Emergent Marine Toxins in Europe: is there a new Invasion? J Marine Sci Res Dev 2013;3:3. Ashton M, Tosteson T, Tosteson C. The effect of elevated temperature on the toxicity of the laboratory cultured dinoflagellate Ostreopsis lenticularis (Dinophyceae). Rev Biol 2003;51: 1–6. Pezzolesi L, Guerrini F, Ciminiello P, Dell’Aversano C, Dello Iacovo E, Fattorusso E, Forino M, Tartaglione L, Pistocchi R. Influence of temperature and salinity on Ostreopsis cf. ovata growth and evaluation of toxin content through HR LC-MS and biological assays. Water Res 2012;1(46):82–92.

294

Panagiota Katikou

[189] Yamaguchi H, Yoshimatsu T, Tanimoto Y, Sato S, Nishimura T, Uehara K, Adachi M. Effects of temperature, salinity and their interaction on growth of the benthic dinoflagellate Ostreopsis cf. ovata (Dinophyceae) from Japanese coastal waters. Phycol Res 2012;60:297–304. [190] Munday R. Palytoxin toxicology: Animal studies. Toxicon 2011;57:470–477. [191] Biré R, Trotereau S, Lemée R, Delpont C, Chabot B, Aumond Y, Krys S. Occurrence of palytoxins in marine organisms from different trophic levels of the French Mediterranean coast harvested in 2009. Harmful Algae 2013;28:10–22. Doi: https://doi.org/10.1016/j. hal.2013.04.007. [192] Brissard C, Herrenknecht C, Séchet V, Hervé F, Pisapia F, Harcouet J, Lémée R, Chomérat N, Hess P, Amzil Z. Complex Toxin Profile of French Mediterranean Ostreopsis cf. ovata Strains, Seafood Accumulation and Ovatoxins Prepurification. Mar Drugs 2014;12:2851–2876. Doi: 10.3390/md12052851. [193] Biré R, Trotereau S, Lemée R, Oregioni D, Delpont C, Krys S, Guérin T. Hunt for Palytoxins in a Wide Variety of Marine Organisms Harvested in 2010 on the French Mediterranean Coast. Mar Drugs 2015;13:5425–5446. Doi: 10.3390/md13085425. [194] Katikou and Vlamis. Palytoxin and Analogs: Ecobiology and Origin, Chemistry, and Chemical Analysis. In: Botana L [ed.] Seafood and Freshwater Toxins: Pharmacology, Physiology, and Detection (3rd edition). CRC Press, Taylor and Francis group: 2014. 695–740. [195] Klijnstra MD, Gerssen A. A Sensitive LC-MS/MS Method for Palytoxin Using Lithium Cationization. Toxins 2018;10:537. Doi: 10.3390/toxins10120537. [196] Reverté L, Soliño L, Carnicer O, Diogène J, Campàs M. Alternative Methods for the Detection of Emerging Marine Toxins: Biosensors, Biochemical Assays and Cell-Based Assays. Mar Drugs 2014;12:5719–5763. [197] FAO/IOC/WHO (Food and Agriculture Organization of the United Nations/ Intergovernmental Oceanographic Commission of UNESCO/World Health Organization). Report of the Joint FAO/ IOC/WHO ad hoc Expert Consultation on Biotoxins in Bivalve Molluscs, Oslo, Norway, 2004, 8. [198] Cembella A, Krock B. Cyclic imine toxins: chemistry, biogeography, biosynthesis and pharmacology. In: Botana LM [ed.] Seafood and Freshwater toxins: Pharmacology, Physiology and Detection (2nd edition). Boca Raton, FL; CRC Press (Taylor and Francys Group): 2008. 561–580. [199] Rhodes L, Smith K, Selwood A, McNabb P, van Ginkel R, Holland P, Munday R. Production of pinnatoxins by a peridinoid dinoflagellate isolated from Northland, New Zealand. Harmful Algae 2010;9:384–389. [200] Selwood AI, Miles CO, Wilkins AL, van Ginkel R, Munday R, Rise F, McNabb P. Isolation, structural determination and acute toxicity of pinnatoxins E, F and G. J Agric Food Chem 2010;58:6532–6542. [201] Otero P, Miguéns N, Rodríguez I, Botana LM. LC–MS/MS Analysis of the Emerging Toxin Pinnatoxin-G and High Levels of Esterified OA Group Toxins in Galician Commercial Mussels. Toxins 2019;11:394. Doi: 10.3390/toxins11070394. [202] CRLMB (Community Reference Laboratory for Marine Biotoxins). Report on toxicology working group meeting. CRLMB: Cesenatico, Italy, 24–25 October 2005, 24–25. [203] Pigozzi S, Bianchi L, Boschetti L, Cangini M, Ceredi A, Magnani F, Milandri A, Montanari S, Pompei M, Riccardi E First evidence of spirolide accumulation in northwestern Adriatic shellfish. In: Proceedings of the 12th ICHA, 4–8 September 2006, ISSHA and IOC of UNESCO: Copenhagen, Denmark 2008, 319–322. [204] Agence nationale de sécurité sanitaire de l’alimentation, de l’environnement et du travail (ANSES). Risques liés aux pinnatoxines dans les coquillages: Avis de l’Anses -Rapport d’expertise collective, Mars 2019 – Édition scientifique. Available online at: https://archimer.ifremer.fr/doc/00504/61546/65428.pdf.

8 Human impact in Mediterranean coastal ecosystems and climate change

295

[205] Cembella AD. Lewis NI and Quilliam MA. The marine dinoflagellate Alexandrium ostenfeldii (Dinophyceae) as the causative organism of spirolide shellfish toxins. Phycologia 2000;39: 67–74. [206] Touzet N, Franco JM, Raine R. Morphogenetic diversity and biotoxin composition of Alexandrium (Dinophyceae) in Irish coastal waters. Harmful Algae 2008;7:782–797. [207] Seki T, Satake M, Mackenzie L, Kaspar HF, Yasumoto T. Gymnodimine, a new marine toxin of unprecedented structure isolated from New-Zeland oysters and the dinofalgellate Gymnnodinium sp. Tetrahedron Lett 1995;36:7093–7096. [208] Salgado P, Riobó P, Rodríguez F, Franco JM, Bravo I. Differences in the toxin profiles of Alexandrium ostenfeldii (Dinophyceae) strains isolated from different geographic origins: Evidence of paralytic toxin, spirolide, and gymnodimine. Toxicon 2015;103:85–98. Doi: https://doi.org/10.1016/j.toxicon.2015.06.015. [209] Martens H, Tillmann U, Harju K, Dell’Aversano C, Tartaglione L, Krock B. Toxin Variability Estimations of 68 Alexandrium ostenfeldii (Dinophyceae) Strains from The Netherlands Reveal a Novel Abundant Gymnodimine. Microorganisms 2017;5:29. Doi: 10.3390/ microorganisms5020029. [210] Torigoe K, Murata M, Yasumoto T, Iwashita T. Prorocentrolide, a toxic nitrogenous macrocycle from a marine dinoflagellate, Prorocentrum lima. J Am Chem Soc 1988;110: 7876–7877. [211] Lu CK, Lee GH, Huang R, Chou HN. Spiro-prorocentrimine, a novel macrocyclic lactone from a benthic Prorocentrum sp. of Taiwan. Tetrahedron Lett 2001;42:1713–1716. [212] John U, Cembella AD, Hummert C, Elbrächter M, Groben R, Medlin L. Discrimination of the toxigenic dinoflagellates Alexandrium tamarense and A. ostenfeldii in co-occurring natural populations from Scottish coastal waters. Eur J Phycol 2003;38:25–40. [213] Ciminiello P, Dell’Aversano C, Fattorusso E, Magno S, Tartaglione L, Cangini M, Pompei M, Guerrini F, Boni L, Pistocchi R. Toxin profile of Alexandrium ostenfeldii (Dinophyceae) from the Northern Adriatic Sea revealed by liquid chromatography–mass spectrometry. Toxicon 2006;47:597–604. [214] MacKinnon SL, Walter JA, Quilliam MA, Cembella AD, Leblanc P, Burton IW, Hardstaff WR, Lewis NI. Spirolides isolated from Danish strains of the toxigenic dinoflagellate Alexandrium ostenfeldii. J, Nat, Prod 2006;69:983–987. [215] Aasen J, MacKinnon SL, LeBlanc P, Walter JA, Hovgaard P, Aune T, Quilliam MA. Detection and identification of spirolides in Norwegian shellfish and plankton. Chem Res Toxicol 2005;18: 509–515. [216] Villar González A, Rodríguez-Velasco ML, Ben-Gigirey B, Botana LM. First evidence of spirolides in Spanish shellfish. Toxicon 2006;48:1068–1074. [217] Nincevic Gladan Z, Ujevic I, Milandri A, Marasovic I, Ceredi A, Pigozzi S, Arapov J, Skejic S. Lipophilic Toxin Profile in Mytilus galloprovincialis during Episodes of Diarrhetic Shellfish Poisoning (DSP) in the N.E. Adriatic Sea in 2006. Molecules 2011;16:888–899. Doi: https://doi.org/10.3390/molecules16010888. [218] Ujević I, Nazlić N, Ninčević-Gladan Ž, Marasović I Gymnodimine and spirolide in shellfish during DSP toxicity in central and southern Adriatic Sea; Proceedings of the ICHA; Wellington, New Zealand: 2014. Oct, pp. 27–31. [219] Ujević I, Roje-Busatto R, Ezgeta-Balić D. Comparison of amnesic, paralytic and lipophilic toxins profiles in cockle (Acanthocardia tuberculata) and smooth clam (Callista chione) from the central Adriatic Sea (Croatia). Toxicon 2019;159:32–37. Doi: https://doi.org/10.1016/j. toxicon.2018.12.008. [220] Katikou P, Aligizaki K, Zacharaki T, Iossifidis D, Nikolaidis G First report on the presence of spirolides in Greek shellfish associated with the detection of the causative Alexandrium

296

[221] [222]

[223] [224]

[225]

[226] [227]

[228]

[229] [230]

[231] [232] [233]

[234] [235] [236]

[237]

Panagiota Katikou

species. In: Proceedings of the 14th International Conference on Harmful Algal Blooms, 1–5 November 2010, Crete, Greece, ed. Pagou KA, Hallegraeff GM International Society for the Study of Harmful Algae and Intergovernmental Oceanographic Commission of UNESCO 2012, 220–222. Stirling DJ. Survey of historical New Zealand shellfish samples for accumulation of gymnodimine. N Z J Mar Freshw Res 2001;35:851–857. MacKenzie L, Holland P, McNabb P, Beuzenberg V, Selwood A, Suzuki T. Complex toxin profi les in phytoplankton and Greenshell mussels (Perna canaliculus), revealed by LC-MS/MS analysis. Toxicon 2002;40:1321–1330. Biré R, Krys S, Frémy JM, Dragacci S, Stirling D, Kharrat R. First evidence on occurrence of gymnodimine in clams from Tunisia. J Nat Toxins 2002;11:269–275. Marrouchi R, Dziri F, Belayouni N, Hamza A, Benoit E, Molgó J, Kharrat R. Quantitative Determination of Gymnodimine-A by High Performance Liquid Chromatography in Contaminated Clams from Tunisia Coastline. Mar Biotech 2010;12:579–585. Ben Naila I, Hamza A, Gdoura R, Diogène J, de la Iglesia P. Prevalence and persistence of gymnodimines in clams from the Gulf of Gabes (Tunisia) studied by mouse bioassay and LC–MS/MS. Harmful Algae 2012;18:56–64. Munday R, Towers NR, Mackenzie L, Beuzenberg V, Holland PT, Miles CO. Acute toxicity of gymnodimine to mice. Toxicon 2004;44:173–178. Alonso E, Vale C, Vieytes MR, Laferla FM, Giménez-Llort L, Botana LM. The cholinergic antagonist Gymnodimine improves Aβ and Tau neuropathology in an in vitro model of Alzheimer disease. Cell Physiol Biochem 2011;27:783–794. Marques A, Rosa R, Nunes ML. Seafood Safety and Human Health Implications. In: Goffredo S, Dubinsky Z [ed.] The Mediterranean Sea – Its history and present challenges. New York London: 2014. 589–603. Miles CO, Rundberget T, Sandvik M, Aasen J, Selwood AI. The Presence of Pinnatoxins in Norwegian Mussels. Oslo, Norway; National Veterinary Institute: 2010. Balech E, Tangen K. Morphology and taxonomy of toxic species in the tamarensis group (Dinophyceae): Alexandrium excavatum (Braarud) comb. nov. and Alexandrium ostenfeldii (Paulsen) comb. nov. Sarsia 1985;70:333–343. Okolodkov YB, Dodge JD. Biodiversity and biogeography of planktonic dinoflagellates in the Arctic Ocean. J Exp Mar Bio Ecol 1996;202:19–27. Konovalova GV. The morphology of Alexandrium ostenfeldii (Dinophyta) from littoral waters of eastern Kamchatka. Botanichyeskii Zhurnal (Leningrad) 1991;76:79–94. Fraga S, Sanchez FJ. Toxic and potentially toxic dinoflagellates found in Galician Rias (NW Spain). In: Anderson DM, White AW, Baden DG [ed.] Toxic Dinoflagellates. North Holland, New York; Elsevier: 1985. 51–54. Balech E. The Genus Alexandrium Halim (Dinoflagellata). Cork, Ireland; Sherkin Island Marine Station, Sherkin Island Co: 1995. Mackenzie L, White D, Oshima Y, Kapa J. The resting cysts and toxicity of Alexandrium ostenfeldii (Dinophyceae) in New Zealand. Phycologia 1996;35:148–155. Sánchez S, Villanueva P, Carbajo L Distribution and concentration of Alexandrium peruvianum (Balech and de Mendiola) in the Peruvian coast (038240–188200 LS) between 1982–2004. In: Abstracts, XI International Conference on Harmful Algal Blooms, Cape Town, South Africa 2004, 15–19, 227. Nagai S, Baba B, Miyazono A, Tahvanainen P, Kremp A, Godhe A, MacKenzie L, Anderson DM. Polymorphisms of the nuclear ribosomal RNA genes found in the different geographic origins in the toxic dinoflagellate Alexandrium ostenfeldii and the species detection from a single cell by LAMP. DNA Polymorph 2010;18:122–126.

8 Human impact in Mediterranean coastal ecosystems and climate change

297

[238] Moestrup Ø, Hansen PJ. On the occurrence of the potentially toxic dinoflagellates Alexandrium tamarense (=Gonyaulax excavata) and A. ostenfeldii in Danish and Faroese waters. Ophelia 1988;28:195–213. [239] Kremp A, Lindholm T, Dreßler N, Erler K, Gerds G, Eirtovaara S, Leskinen E. Bloom forming Alexandrium ostenfeldii (Dinophyceae) in shallow waters of the Aland Archipelago, Northern Baltic Sea. Harmful Algae 2009;8:318–328. [240] Tomas CR, Van Wagoner RM, Tatters AO, White KD, Hall S, Wright JLC. Alexandrium peruvianum (Balech and Mediola) Balech and Tangen a new toxic species for coastal North Carolina. Harmful Algae 2012;17:54–63. [241] Burson A, Matthijs HCP, de Bruijne W, Talens R, Hoogenboom R, Gerssen A, Visser PM, Stomp M, Steur K, van Scheppingen Y, Huisman J. Termination of a toxic Alexandrium bloom with hydrogen peroxide. Harmful Algae 2014;31:125–135. [242] Brandenburg KM, Domis LNDS, Wohlrab S, Krock B, John U, van Scheppingen Y, van Donk E. van de Waal DB. Combined physical, chemical and biological factors shape Alexandrium ostenfeldii blooms in The Netherlands. Harmful Algae 2017;63:146–153. Doi: https://doi.org/10.1016/j.hal.2017.02.004. [243] Hakanen P, Suikkanen S, Franzén J, Franzén H, Kankaanpää H, Kremp A. Bloom and toxin dynamics of Alexandrium ostenfeldii in a shallow embayment at the SW coast of Finland, northern Baltic Sea. Harmful Algae 2012;15:91–99. [244] Kremp A, Godhe A, Egardt J, Dupont S, Suikkanen S, Casabianca S, Penna A. Intraspecific variability in the response of bloom forming marine microalgae to changing climatic conditions. Ecol Evol 2012;2:1195–1207. [245] Hansen PJ, Cembella AD, Moestrup Ø. The marine dinoflagellate Alexandrium ostenfeldii: paralytic shellfish toxin concentration, composition, and toxicity to a tintinnid ciliate. J Phycol 1992;28:597–603. [246] Van Wagoner RM, Misner I, Tomas CR, Wright JLC. Occurrence of 12– methylgymnodimine in a spirolide-producing dinoflagellate Alexandrium peruvianum and the biogenetic implications. Tetrahedron Lett 2011;52:4243–4246. [247] Tillmann U, Kremp A, Tahvanainen P, Krock B. Characterization of spirolide producing Alexandrium ostenfeldii (Dinophyceae) from the western Arctic. Harmful Algae 2014;39:259–270. [248] Medhioub A, Medhioub W, Amzil Z, Sibat M, Bardouil M, Ben Neila I, Mezghani S, Hamza A, Lassus P. Influence of environmental parameters on Karenia selliformis toxin content in culture. Cah Biol Mar 2009;50:333–342. [249] Hernández-Becerril DU, Rodríguez-Palacio MC, Lozano-Ramírez C. Morphology and life stages of the potentially pinnatoxin-producing thecate dinoflagellate Vulcanodinium rugosum from the tropical Mexican Pacific. Botanica Marina 2013;56:535–540. [250] Aligizaki K, Moschandreou K, Arsenakis M Monitoring Programme of Potentially Toxic Marine Microalgae: a review. In: Gkelis S, Karousou R, Kokkini S, Panteris E (eds.). 2013. Program and Abstracts. 13th Panhellenic Scientific Conference, Hellenic Botanical Society, Thessaloniki 3–6 October 2013, p. 29. Available online at: https://www.hbs.gr/images/files/ conferences/13/EBE13_BOOK_OF_ABSTRACTS.pdf (accessed on 14 December 2019). [251] Abadie E, Muguet A, Berteaux T, Chomérat N, Hess P, Roque D’OrbCastel E, Masseret E, Laabir M. Toxin and Growth Responses of the Neurotoxic Dinoflagellate Vulcanodinium rugosum to Varying Temperature and Salinity. Toxins 2016;8:136. Doi: https://doi.org/ 10.3390/toxins8050136. [252] Rhodes L, Smith K, Selwood A, McNabb P, Munday R, Suda S, Molenaar S, Hallegraeff G. Dinoflagellate Vulcanodinium rugosum identified as the causative organism of pinnatoxins in Australia, New Zealand and Japan. Phycologia: November 2011;50:624–628.

298

Panagiota Katikou

[253] Stivala CE, Benoit E, Araoz R, Servent D, Novikov A, Molgo J, Zakarian A. Synthesis and biology of cyclic imine toxins, an emerging class of potent, globally distributed marine toxins. Nat Prod Rep 2014. Doi: 10.1039/C4NP00089G. [254] Murata M, Legrand AM, Ishibashi Y, Yasumoto T. Structures of ciguatoxin and its congener. J Am Chem Soc 1989;111:8929–8931. [255] Murata M, Legrand AN, Ishibashi Y, Fukui M, Yasumoto T. Conformations of ciguatoxin and related polyethers. Abstr Pap Am Chem Soc 1990;200:54-AGFD. [256] Lehane L, Lewis RJ. Ciguatera: Recent advances but the risk remains. Int J Food Microbiol 2000;61:91–125. [257] Lehane L. Ciguatera update. Med J Australia 2000;172:176–179. [258] Litaker RW, Holland WC, Hardison DR, Pisapia F, Hess P, Kibler SR, Tester PA. Ciguatoxicity of Gambierdiscus and Fukuyoa species from the Caribbean and Gulf of Mexico. PLOS ONE 2017;12:e0185776. Doi: https://doi.org/10.1371/journal.pone.0185776. [259] Aligizaki K, Nikolaidis G. Morphological identification of two tropical dinoflagellates of the genera Gambierdiscus and Sinophysis in the Mediterranean Sea. J Biol Res -Thessalon 2008;9:75–82. [260] Otero P, Pérez S, Alfonso A, Vale C, Rodríguez P, Gouveia NN, Gouveia N, Delgado J, Vale P, Hirama M, Ishihara Y, Molgó J, Botana LM. First toxin profile of ciguateric fish in Madeira Arquipelago (Europe). Anal Chem 2010;82:6032–6039. [261] CDC (Centers for Disease Control and Prevention). Cluster of Ciguatera Fish Poisoning, 2007. Morbidity and Mortality Weekly Report (MMWR). North Carolina, NC, USA; CDC: 2009. 283–285. [262] Hossen V, Soliño L, Leroy P, David E, Velge P, Dragacci S, Krys S, Quintana HF, Diogène J. Contribution to the risk characterization of ciguatoxins: LOAEL estimated from eight ciguatera fish poisoning events in Guadeloupe (French West Indies). Environ Res 2015;143: 100–108. Doi: https://doi.org/10.1016/j.envres.2015.09.014. [263] Mills AR. Poisonous fish in the South Pacific. J Trop Med Hyg 1956;59:99–103. [264] Vernoux JP, Lahlou N, Elandaloussi SA, Riyeche N, Magras LP. A study of the distribution of ciguatoxin in indiviual Caribbean fish. Acta Trop 1985;42:225–233. [265] Lewis RJ, Sellin M, Poli MA, Norton RS, MacLeod JK, Sheil MM. Purification and characterization of ciguatoxins from moray eel (Lycodontis javanicus, Muraenidae). Toxicon 1991;29:1115–1127. [266] Satake M, Ishimaru T, Legrand AM, Yasumoto T. Isolation of a ciguatoxin analog from cultures of Gambierdiscus toxicus. In: Samyda TJ, Shimizu Y [ed.] Toxic Pyhtoplankton Blooms in the Sea. New York, NY, USA; Elsevier: 1993, Vol. 3, 575–579. [267] Satake M, Murata M, Yasumoto T. The structure of CTX3C, a ciguatoxin congener isolated from cultured Gambierdiscus toxicus. Tetrahedron Lett 1993;34:1975–1978. [268] Satake M, Fukui M, Legrand AM, Cruchet P, Yasumoto T. Isolation and structures of new ciguatoxin analogs, 2,3-dihydroxyCTX3C and 51-hydroxyCTX3C, accumulated in tropical reef fish. Tetrahedron Lett 1998;39:1197–1198. [269] Soliño L, Costa PR. Differential toxin profiles of ciguatoxins in marine organisms: Chemistry, fate and global distribution. Toxicon 2018;150:124–143. Doi: https://doi.org/10.1016/j. toxicon.2018.05.005. [270] Lewis RJ, Vernoux JP, Brereton IM. Structure of Caribbean ciguatoxin isolated from Caranx latus. J Am Chem Soc 1998;120:5914–5920. [271] Hamilton B, Hurbungs M, Vernoux JP, Jones A, Lewis RJ. Isolation and characterisation of Indian Ocean ciguatoxin. Toxicon 2002;40:685–693.

8 Human impact in Mediterranean coastal ecosystems and climate change

299

[272] Perez-Arellano JL, Luzardo OP, Brito AP, Cabrera MH, Zumbado M, Carranza C, Angel-Moreno A, Dickey RW, Boada LD. Ciguatera fish poisoning, Canary Islands. Emerg Infect Dis 2005;11: 1981–1982. [273] Estevez P, Castro D, Leao JM, Yasumoto T, Dickey R, Gago-Martinez A. Implementation of liquid chromatography tandem mass spectrometry for the analysis of ciguatera fish poisoning in contaminated fish samples from Atlantic coasts. Food Chem 2019;280:8–14. Doi: https://doi.org/10.1016/j.foodchem.2018.12.038. [274] Canals-Caballero A Risk characterization of ciguatera food poisoning in Europe: Update on the EuroCigua Project. Presentation in the 71st Advisory Forum Meeting 03–04 April 2019, Bucharest, Romania. Available online at: https://www.efsa.europa.eu/sites/default/files/ event/AF190403-p16a_ES.pdf (accessed on 15 December 2019). [275] Gouveia N, Delgado J, Vale P. Primeiro registo da ocorrência de episódios do tipo ciguatérico no arquipélago da Madeira. In: Abstract Book of X Reuniao Oberica, Fitoplancton Toxico e Biotoxinas. Lisbon, Portugal; IPIMAR: 2009. [276] Estevez P, Castro D, Pequeño-Valtierra A, Leao JM, Vilariño O, Diogène J, Gago-Martínez A. An Attempt to Characterize the Ciguatoxin Profile in Seriola fasciata Causing Ciguatera Fish Poisoning in Macaronesia. Toxins 2019;11:221. Doi: https://doi.org/10.3390/toxins11040221. [277] Bentur Y, Spanier E. Ciguatoxin-like substances in edible fish on the eastern Mediterranean. Clin Toxicol (Phila) 2007;45:695–700. [278] Fraga S. Global climate change and harmful algal blooms (HABs). In: Abstract Book of 4th European Phycological Congress. Oviedo, Spain; Elsevier Science: 2007. 41. [279] Stebbing ARD, Turk SMT, Wheeler A, Clarke KR. Immigration of southern fish species to south-west England linked to warming of the North Atlantic (1960–2001). J Mar Biol Assoc UK 2002;82:177–180. [280] Fraga S, Riobό P, Diogène J, Paz B, Franco JM Toxic and potentially toxic benthic dinoflagellates observed in Macaronesia (NE Atlantic Archipelagos). In: Programme and Abstracts. XI Int Conf Harmful Algae, 14–19 November, Capetown 2004, 115. [281] Aligizaki K, Nikolaidis G, Fraga S. Is Gambierdiscus expanding to new areas? Harmful Algae News 2008;36:6–7. [282] Tudó A, Toldrá A, Andree K, Rey M, Fernández-Tejedor M, Campás M, Diogène J. First report of Gambierdiscus in the Western Mediterranean Sea (Balearic Islands). Harmful Algae News 2018;59:22–23. [283] Laza‐Martínez A, David H, Riobó P, Miguel I, Orive E. Characterization of a Strain of Fukuyoa paulensis (Dinophyceae) from the Western Mediterranean Sea. J Eukaryot Microbiol 2016;63: 481–497. Doi: 10.1111/jeu.12292. [284] À T, Toldrà A, Andree KB, Rey M, Fernández-Tejedor M, Aligizaki K, Iliadou M, Arsenakis M, Campàs M, Diogène J, Ricciardi W, Marcheggiani S, Puccinelli C, Carere M, Sofia T, Giuliano F, Dogliotti E, Mancini M (ed.). First Scientific Symposium Health and Climate Change. Istituto Superiore di Sanità. Rome, December 3-5, 2018. Abstract book. Roma: Istituto Superiore di Sanità, 2018 (ISTISAN Congressi 18/C5), p. 156. [285] Turner AD, Higgins C, Davidson K, Veszelovszki A, Payne D, Hungerford J, Higman W. Potential Threats Posed by New or Emerging Marine Biotoxins in UK Waters and Examination of Detection Methodology Used in Their Control: Brevetoxins. Mar Drugs 2015;13:1224–1254. Doi: https://doi.org/10.3390/md13031224. [286] FAO (Food and Agriculture Organization of the United Nations). Marine biotoxins. FAO Food and nutrition paper 2004, 80, 1–287. [287] Fukuyo Y, Takano H, Chihara M, Matsuoka K. Red Tide Organisms in Japan. An Illustrated Taxonomic Guide. Tokyo; Uchida Rokakuho, Co., Ltd.: 1990. 407.

300

Panagiota Katikou

[288] Taylor FJR, Fukuyo Y, Larsen J. Taxonomy of harmful dinoflagellates. In: Hallegraeff GM, Anderson DM, Cembella AD [ed.] Manual on Harmful Marine Microalgae, IOC Manuals and Guides No. 33. France; UNESCO: 1995. 283–317. [289] Steidinger KA, Tangen K. Dinoflagellates. In: CR T [ed.] Identifying Marine Diatoms and Dinoflagellates. New York; Academic Press: 1996. 387–598. [290] Steidinger KA. Implications of dinoflagellate life cycles on initiation of Gymnodinium breve red tides. Environ Lett 1975;9:129–139. [291] Steidinger KA, Truby EW, Dawes CJ. Ultrastructure of the red tide dinoflagellate Gymnodinium breve. I General description J Phycol 1978;14:72–79. [292] Faust MA, Gulledge RA. Identifying Harmful Marine Dinoflagellates. Smithsonian Institution, Contributions from the United States National Herbarium: 2002, Vol. 42, 1–144. [293] Vargo GA. A brief summary of the physiology and ecology of Karenia brevis Davis (G. Hansen and Moestrup comb. nov.) red tides on the West Florida Shelf and of hypotheses posed for their initiation, growth, maintenance, and termination. Harmful Algae Volume 2009;8: 573–584. [294] Fu FX, Zhang Y, Warner ME, Feng Y, Sun J, Hutchins DA. A comparison of future increased CO2 and temperature effects on sympatric Heterosigma akashiwo and Prorocentrum minimum. Harmful Algae 2008;7:76–90. [295] Sun J, Hutchins DA, Feng Y, Seubert EL, D A C, Fu FX. Effects of changing pCO2 and phosphate availability on domoic acid production and physiology of the marine harmful bloom diatom Pseudo-nitzschia multiseries. Limnol Oceanogr 2011;56:829–840. [296] Matheson JR The Effects of Ocean Acidification and Eutrophication on the Growth, Lipid Composition and Toxicity of the Marine Raphidophyte Heterosigma Akashiwo. University of Western Ontario – Electronic Thesis and Dissertation Repository 2014, Paper 1983. [297] Pistocchi R, Guerrini F, Pezzolesi L, Riccardi M, Vanucci S, Ciminiello P, Dell’Aversano C, Forino M, Fattorusso E, Tartaglione L, Milandri A, Pompei M, Cangini M, Pigozzi S, Riccardi E. Toxin Levels and Profiles in Microalgae from the North-Western Adriatic Sea—15 Years of Studies on Cultured Species. Mar Drugs 2012;10:140–162. [298] Koray T. The occurrence of red tides and causative organisms in Izmir Bay. Ege Universitesi Fen Fakültesi Dergisi, Seri B 1984;1:75–83. [299] Bizsel N, Bizsel C. New records of toxic algae Heterosigma cf. akashiwo and Gymnodinium cf. mikimotoi in the hypereutrophic Izmir Bay (Aegean Sea): Coupling between organisms and water quality parameters. Isr J Plant Sci 2002;50:33–44. [300] Labib W Red tide occurrence in Alexandria (Egypt). A review. GFCM Workshop on Algal and Jellyfish Blooms in the Mediterranean and Black Sea. 6th /8th October 2010, Istanbul, Turkey. [301] Durgham H, Ikhtiyar S. First records of alien toxic algae Heterosigma akashiwo (Raphidophyceae) from the Mediterranean Coast of Syria. Arab Gulf J Sci Res 2012;30(1): 58–60. [302] U.S. FDA (United States Food and Drug Administration), 2001. Fish and Fisheries Products Hazards and Controls Guidance, 3rd edition. Appendix 5 – FDA & EPA Safety Levels in Regulations and Guidance. June 2001. Available online at: http://www.fda.gov/downloads/ Food/GuidanceRegulation/UCM252448.pdf (accessed on 15/ 12/2019). [303] NZG (New Zealand Government), 2018. Animal Products Notice (Regulated Control Scheme – Bivalve Molluscan Shellfish for Human Consumption, 2 August 2018. Available online at: https://www.mpi.govt.nz/dmsdocument/30282-animal-products-noticeregulated-control-scheme-bivalve-molluscan-shellfish-for-human-consumption-2018 (accessed on 15/ 12/2019).

8 Human impact in Mediterranean coastal ecosystems and climate change

301

[304] FSANZ (Food Standards Australia New Zealand), 2010. Food Standard Code, Incorporating amendments up to and including Amendment 116, Standard 4.1.1, Primary Production and Processing Standards, Preliminary provisions, Standard 1.4.1, Contaminants and Natural toxicants, Issue 111. Available online at: http://www.comlaw.gov.au/Details/F2012C00285/ Download (accessed on 15/ 12/2019). [305] Vilariño N, Louzao MC, Abal P, Cagide E, Carrera C, Vieytes MR, Botana LM. Human Poisoning from Marine Toxins: Unknowns for Optimal Consumer Protection. Toxins 2018;10:324. Doi: https://doi.org/10.3390/toxins10080324.

Hazel Farrell, Gurjeet S. Kohli and Shauna A. Murray

9 Gambierdiscus, the cause of ciguatera fish poisoning: an increased human health threat influenced by climate change 9.1 The genus Gambierdiscus Recent advances in population and species genetics for phytoplankton have revealed immense biodiversity at different taxonomic levels [1]. Vast numbers of species remain to be documented, aided by rapidly developing molecular methods [2]. Approximately 200 benthic (sand dwelling and epiphytic) dinoflagellates are now known [3, 4]. The first report by Yasumoto et al. [5] of the involvement of a benthic dinoflagellate in ciguatera fish poisoning (CFP) brought increased attention to this group. The implicated species were described as Gambierdiscus based on the type species Gambierdiscus toxicus, from samples collected in the Gambier Islands, French Polynesia [6]. Species within the genus Gambierdiscus are recognized as the main producers of ciguatoxins (CTXs) and maitotoxins (MTXs), as well as other compounds such as 44-methylgambierone (previously known as MTX3) [7–14]. More recently, molecular taxonomy resulted in the new genus Fukuyoa being described for some of the species previously considered within the genus Gambierdiscus [15]. It has been reported that some Fukuyoa species may produce MTX1 and compounds that have shown some CTX-like activities in a bioassay [16]. CFP is the most common nonbacterial illnesses associated with fish consumption [17], affecting between 50,000 and 500,000 people per year [18]. The ingestion of herbivorous and carnivorous fish that have orally accumulated effective levels of CTXs, and possible MTXs, can cause CFP in humans [19–21]. Recent reviews have illustrated the global increase in the frequency and intensity of harmful algal events [22, 23]. Despite being significantly underreported, CFP occurrence worldwide is increasing, with reports of a 60% increase in CFP in the Pacific Islands over the past decade [24]. New species of Gambieriscus are continually being described, with evidence showing that each species might have its own characteristic toxin profile [10, 12, 13]. As in the case of other dinoflagellate genera such as Alexandrium or Karenia, the production or not of certain toxin groups appears to generally vary at the species level, rather than being consistent within the genus [16]. For this reason, species of harmful algal bloom (HAB)-forming taxa are monitored, acting as early-warning systems for shellfish and seafood safety. This review highlights the significant advances in the Hazel Farrell, NSW Department of Primary Industries, NSW Food Authority, NSW, Australia Gurjeet S. Kohli, University of Technology Sydney, NSW, Australia Shauna A. Murray, School of Life Sciences, University of Technology Sydney, NSW, Australia https://doi.org/10.1515/9783110625738-009

304

Hazel Farrell, Gurjeet S. Kohli and Shauna A. Murray

study of Gambierdiscus and Fukuyoa. We provide a summary of the morphology and phylogenetics of species of Gambierdiscus and Fukuyoa, their toxicology, distribution, chemistry and methods for the detection of CTXs and MTXs in seafood. While, in recent years, advances have been made in our understanding of these genera, this review further outlines the major gaps in our current understanding of Gambierdiscus and Fukuyoa and outlines goals for future research in this field.

9.2 Morphology and phylogenetics Although originally considered as a monotypic taxon [6], variability in the morphology, differences in ribosomal RNA (rRNA) genes, toxicity and physiological characteristics [6, 7, 25–31] led to the description of new species of Gambierdiscus. Currently, 18 species of Gambierdiscus have been described, based on their distinct morphological and molecular genetic characteristics (Table 9.1). The new genus Fukuyoa has three described species to date. The original species descriptions for Gambierdiscus consisted of a comprehensive account of their characteristics. Gambierdiscus cells are generally large (60–100 μm), armored, having a distinct thecal plate pattern and fishhook-shaped apical pore. The descriptions differentiated species as either being anterioposteriorly compressed (lenticular) or slightly laterally compressed (globular). Originally, the two globular species (G. yasumotoi and G. ruetzleri) were distinguished from each other by cell size, size and shape of the 2ʹ apical and 2ʹʹʹʹ antapical plate and depth to width ratio, described in detail in Litaker et al. [32]. These two species have since been redescribed within the genus Fukuyoa as Fukuyoa yasumotoi and Fukuyoa ruetzleri, with the highly similar species Fukuyoa paulensis also included in this grouping [15]. The morphological differences between the three Fukuyoa species are described in detail in [15] (Figure 9.1). Within the genus Gambierdiscus, eighteen currently described species (Figure 9.2) are anterioposteriorly compressed and broadly classified by either a narrow (G. australes, G. balechii, G. belizeanus, G. cheloniae, G. excentricus, G. honu, G. lapillus, G. lewisii, G. pacificus and G. scabrosus) or broad (G. caribaeus, G. carolinianus, G. carpenteri, G. holmesii, G. jejuensis, G. polynesiensis, G. silvae and G. toxicus) 1p posterior intercalary plate or equivalent 2ʹʹʹʹ plate. Among the species with a narrow 1p posterior intercalary plate, further distinguishing characteristics are either having a smooth cell surface (G. australes, G. excentricus, G. honu and G. pacificus) or not (G. balechii, G. belizeanus, G. cheloniae, G. lapillus, G. lewisii and G. scabrosus). Species with a smooth cell surface can be distinguished by either having a hatchet-shaped 2ʹ apical plate (G. honu and G. pacificus) or a more conventional rectangular-shaped 2ʹ apical plate (G. australes and G. excentricus). G. excentricus is at least 1.5 times wider and deeper than G. australes; further specifics distinguishing the two are described in detail in the original descriptions of the species

Smaller cell size Plate ʹ narrow pentagonal

Fukuyoa ruetzleri (. ± .) × (Faust, Litaker, Vandersea, (. ± .) × Kibler, Holland et Tester) (. ± .) Gómez, Qiu, Lopes et Lin

Gambierdiscus australes Chinian et Faust

(. ± .) × (. ± .) × (. ± .)

Narrow p plate Smooth cell surface ʹ Rectangular shaped Smaller than G. excentricus

Larger cell size Plate ʹ nearly rectangular

(. ± .) × (. ± .) × (. ± .)

Fukuyoa yasumotoi (Holmes) Gómez, Qiu, Lopes et Lin

Gambierdiscus spp. Narrow p plate (smooth cell surface)

Broad ʹ plate Thecal pores are round and numerous. Intermediate in size between F. yasumotoi and F. ruetzleri

Morphological characteristics (plate formula)

( ± ) × ( ± ) × ( ± )

Cell size (μm) (depth × width × length)

Fukuyoa paulensis Gómez, Qiu, Lopes et Lin Type species

Fukuyoa spp.

Species

Table 9.1: Taxonomic and genetic identifications of different species of Gambierdiscus and Fukuyoa.

SSU: EF- D-D LSU: EF- D-D LSU: EF-

SSU: EF- D-D LSU: EF- D-D LSU: EF-

SSU: EF- D-D LSU: EF- D-D LSU: EF-

D-D LSU: LN., D-D LSU LN

Genetics

(continued )

[, ]

[]

[, ]

[]

References

9 Gambierdiscus, the cause of ciguatera fish poisoning

305

Cell size (μm) (depth × width × length)

Gambierdiscus balechii Fraga, Rodríguez et Bravo Gambierdiscus sp. type 

(. ± .) × (. ± .) × ( ± .) length/width ratio = .

Narrow ʹʹʹʹ plate (p plate equivalent) Ornamented cell surface ʹ Hatchet shaped ʹ’ Asymmetrical shape

Narrow p plate Smooth cell surface ʹ Hatchet shaped

(. ± .) × (. ± .) × (. ± .)

Gambierdiscus pacificus Chinain et Faust

Narrow p plate (ornamented cell surface)

p plate is long relative to cell size Smooth cell surface ʹ Hatchet shaped ʹ Pentagonal shaped Similar cell size to G. cheloniae (rugose cell surface)

(. ± .) × (. ± .) × (. ± .)

Narrow ʹʹʹʹ plate (p plate equivalent) Smooth cell surface ʹ Rectangular shaped Cell size bigger than G. australes (. times wider and deeper)

Morphological characteristics (plate formula)

Gambierdiscus honu Rhodes, Smith et Murray

Gambierdiscus excentricus (. ± ) × Fraga ( ± ) × ( ± )

Species

Table 9.1 (continued )

LSU: KY; SSU: KY

SSU: EF- D-D LSU: EF- D-D LSU: EF-, EF-

SSU: KU; D-D LSU KU.; D-D LSU KU;

D-D LSU: HQ, JF, JF- D-D LSU: JF-

Genetics

[, ]

[, ]

[]

[]

References

306 Hazel Farrell, Gurjeet S. Kohli and Shauna A. Murray

Narrow ″″ plate (p plate equivalent) SSU: KU.; LSU D-D Strong reticulate-foveate ornamentation on cell KU.; surface Small cell size compared to other Gambierdiscus spp. Similar to G. belizeanus, but can be distinguished by the asymmetry of the ″ plate Narrow, short ʹʹʹʹ plate (p plate equivalent) ʹ Hatchet shaped Reticulate-foveate cell surface – two different types

(depth: . range . –. SD:.) × (width:  range –. SD:.) Depth to width ratio .

(depth: . range . –. SD: .) × (width: . range .–. SD: .) x (length: . range ..-. SD: .) Depth to width ratio . SD: .

Gambierdiscus lapillus Kretzschmar, Hoppenrath et Murray

Gambierdiscus lewisii Larsson, Kretzschmar, Hoppenrath, Doblin et Murray

SSU: MH; LSU D-D: MH, LSU D-: MH

SSU: KU.; LSU D-D KU; LSU D-D KU

Narrow p plate with pointed dorsal end Rugose cell surface ʹ Hatchet shaped Apical pore plate varies from those of G. belizeanus and G. pacificus, which are shorter and narrower, and from G. toxicus, which is larger

(. ± .) × (. ± .) × (. ± .)

Gambierdiscus cheloniae Smith, Rhodes et Murray

SSU: EF- D-D LSU: EF- D-D LSU: EF-

Narrow p plate Heavily aerolated cell surface Different ʹ plate symmetry and size

(. ± .) × (. ± .) × (. ± .)

Gambierdiscus belizeanus Faust

(continued )

[]

[]

[]

[, ]

9 Gambierdiscus, the cause of ciguatera fish poisoning

307

Narrow ʹʹʹʹ plate (p plate equivalent) Aerolated cell surface ʹ Rectangle shaped Resembles G. belizeanus

(. ± .) × (. ± .) × (. ± .)

Gambierdiscus scabrosus Nishimura, Shinya Sato et Adachi

Broad p plate ʹ Rectangular shaped Symmetric ʹ’ Broad p plate, dorsal end oblique ʹ Hatchet shaped Larger cell size than G. polynesiensis Broad p plate ʹ’ Rectangular shaped Asymmetric ʹ’ Broad p plate, pentagonal shape ʹ Hatchet shaped, largest apical plate Smooth cell surface

(. ± .) × (. ± .) × ( ± .)

(. ± .) × (. ± .) × (. ± .)

(. ± .) × (. ± .) × (. ± .)

(–) × (–) × (–)

Gambierdiscus caribaeus Vandersea, Litaker, Faust, Kibler, Holland et Tester

Gambierdiscus carolinianus Litaker, Vandersea, Faust, Kibler, Holland et Tester

Gambierdiscus carpenteri Kibler, Litaker, Faust, Holland, Vandersea et Tester Gambierdiscus jejuensis Jang et Jeong Gambierdiscus sp. type 

Broad p plate

Morphological characteristics (plate formula)

Cell size (μm) (depth × width × length)

Species

Table 9.1 (continued )

SSU: EF- D-D LSU: EF-, EF D-D LSU: EF- SSU: MH; LSU: MH

SSU: EF-EF D-D LSU: EF- D-D LSU: EF-

SSU: EF- D-D LSU: EF-, EF, EF D-D LSU: EF-

SSU: AB, LSU D-D: AB.;

Genetics

[–]

[]

[]

[]

[]

References

308 Hazel Farrell, Gurjeet S. Kohli and Shauna A. Murray

Broad p plate, dorsal end oblique ʹ Hatchet shaped Smaller cell size than G. carolinianus Broad ʹʹʹʹ plate (p plate equivalent), largest plate of hypotheca ʹ Hatchet shaped Thick theca with many scattered pores Broad p plate, dorsal end pointed ʹ Hatchet shaped

(. ± ) × (. ± .) × (. ± .)

( ± ) × ( ± ) × ( ± )

 ± .) × ( ± .) × ( ± .)

Gambierdiscus polynesiensis Chinain et Faust

Gambierdiscus silvae Fraga et Rodríguez Gambierdiscus ribotype 

G. toxicus (Adachi et Fukuyo) Chinain, Faust, Holmes, Litaker et Tester Type species

Not described

Not described

Gambierdiscus ribotype 

Gambierdiscus sp. type 

Genetically described phylotypes

Large ʹʹʹʹ plate (p plate equivalent) ʹ Hatchet shaped, wide Smooth to foveate cell surface

(depth: . range . –. SD: .) × (width: . range .–. SD: .) Depth to width ratio . SD: .

Gambierdiscus holmesii Kretzschmar, Larsson, Hoppenrath, Doblin et Murray

SSU: AB-, AB, AB- LSU D-D: AB-

D-D LSU: GU-, GU, GU, GU-

SSU: EF- D-D LSU: EF- D-D LSU: EF-

LSU D-D: KJ.; LSU D-D: KJ

SSU: EF- D-D LSU: EF- D-D LSU: EF-

SSU: MH, LSU D-D: MH, LSU D-D: MH

(continued )

[, ]

[]

[, , , ]

[, ]

[, ]

[]

9 Gambierdiscus, the cause of ciguatera fish poisoning

309

Cell size (μm) (depth × width × length)

Not described

Not described

Not described

Species

Gambierdiscus sp. type 

Gambierdiscus sp. type 

Gambierdiscus sp. type 

Table 9.1 (continued ) Morphological characteristics (plate formula)

SSU: AB-LSU DD: AB- LSU D-D: AB-

Genetics

[]

[, ]

[]

References

310 Hazel Farrell, Gurjeet S. Kohli and Shauna A. Murray

9 Gambierdiscus, the cause of ciguatera fish poisoning

F. paulensis 4″ 5″ 3′ 2′ 3″ 6″

5″

4″

2″ 1″

7″

F. yasumotoi 4″

3′ 2′ 3″ 6″

1′ 7″

F. ruetzleri

311

5″ 3′ 2′ 3″

1′ 2″

6″ 1′

1″

7″

2″ 1″

Figure 9.1: Comparative line drawings of the epitheca for F. paulensis, F. ruetzleri and F. yasumotoi from [15]. For relative cell measurements refer to Table 9.1.

[9, 12]. Some species that have ornamented thecal plates can be further discriminated by having either a rectangular- (G. scabrosus) or hatchet- (G. balechii G. cheloniae and G. lewisii) shaped 2ʹ plate. G. belizeanus has a heavily aerolated cell surface with different 2ʹ plate size and symmetry. The cell size of G. lapillus is small compared to other species in the genus, and cells have a strong reticulatefoveate surface ornamentation. G. lapillus and G. lewisii are similar to G. belizeanus but G. lapillus and G. lewisii have a more asymmetrical 4ʹ’ plate shape in contrast to the more symmetrical 4ʹ’ plate of G. belizeanus. G. lewisii and G. lapillus are morphologically similar but they can be distinguished based on size, as G. lapillus is smaller. Species of Gambierdiscus with a broad 1p posterior intercalary plate can be further differentiated as having a rectangular-shaped (G. caribeaus and G. carpenteri) or a hatchet-shaped (G. carolinianus, G. holmseii, G. jejuensis, G. polynesiensis, G. silvae and G. toxicus) 2ʹ apical plate. G. toxicus is further discerned by a pointed dorsal end to the 1p posterior intercalary plate. Further differences between G. polynesiensis and G. carolinianus are detailed in Litaker et al. [32]. Generally, G. holmesii is significantly larger than G. silvae and distinctly smaller than G. toxicus. G. silvae is similar to G. polynesiensis and the distinct differences between them and G. silvae and G. carolinianus are outlined in [44]. G. caribeaus and G. carpenteri, both possessing a rectangular-shaped 2ʹ apical plate, are distinguished by the shape of the 4ʹ’ precingular plate, which is symmetric in G. caribaeus and asymmetric in G. carpenteri. The size and shape of the sulcal plates and various other specific morphological characteristics for Gambierdiscus, as summarized in Table 9.1, can be reviewed in the original descriptions of the species [8, 9, 12, 32–40, 44]. These features are relatively straightforward to observe using light and scanning electron microscopy; however, within some species, a considerable amount of variability in features such as the size and shape of individual plates may be present. Another technique to identify different species of Gambierdiscus is to compare sequences that are known to be characteristic at the species level, such as regions of rRNA genes. Based on phylogenetic analysis of regions of the SSU (small ribosomal subunit) rDNA, LSU (large ribosomal subunit) rDNA and ITS (internal transcribed spacer) rDNA, the genus Gambierdiscus is monophyletic [9, 12, 29, 32, 43–45].

312

Hazel Farrell, Gurjeet S. Kohli and Shauna A. Murray

Figure 9.2: Comparative line drawings of the epitheca for Gambierdiscus spp. compiled from [32–34, 37, 38, 40, 41, 44]. Scale bar = 50 µm. For relative cell measurements refer to Table 9.1.

Further, the lenticular and globular species formed two distinct clades, which has recently resulted in the three known globular species being described as the new genus Fukuyoa [15]. Phylogenetic analysis has shown that species within the genus Fukuyoa diverged relatively early in the evolution of the genus Gambierdiscus [15, 32, 43]. Also, F. ruetzleri and F. yasumotoi are the two most closely related species in the genus. Based on LSU rDNA D08-D10 sequences, the mean p distance within species is 0.002 ± 0.002, and between species is 0.121 ± 0.036 (calculated based on sequences from 10 species/phylotypes) where minimum p distance between F. ruetzleri and F. yasumotoi is 0.007 [43]. Using SSU rDNA sequences, the mean p distance within species is 0.003 ± 0.002 and between species is 0.139 ± 0.042 (calculated based on sequences from 10 species/phylotypes) where minimum p distance between F. ruetzleri and F. yasumotoi is 0.004 [43]. These statistics are indicative of putative unknown species and can be very useful in cases where morphological, physiological or other data is not yet available, or a strain is not present in culture.

9 Gambierdiscus, the cause of ciguatera fish poisoning

CH3

OH O

O

O

O

R2

O

O

HO

CH3

O O

O

O

OH O

H3C

R1

O

O

O

O

CH3

O O

O

O

OH O

Type II P-CTX-3C

O

O

O

Type III C-CTX-1 O

OH CH3 O

O OH

O

H O

H O

O

H

O H

H3C H 3C

H

O H

H

R1

CH3

O O

O

O

HO

O

O

O HO

H3C

CH3

OH O

O

H3C

HO

O

O

CH3

OH O HO

O

H 3C

Type I CTX backbone Ciguatoxin, R1;R2 P-CTX-1, OH; CH(OH)CH2OH P-CTX-2, H; CH(OH)CH2OH P-CTX-4B, H; CH=CH2

O

313

O

OH H O

H

H O

O H

O H H

H

H

OH

OH

H O

O

H O H O

HO H O OH OH

OH H HO

O O H H OH

H

H

O O

H

H

H OH

OH O

H

OH H OH

OH

H

O

H

O

H

H

H O

H

O

H

H OH

OH

O HO

OH H HO H O O H OH H OH HO H H OH HO O H O H H OH

O

H

OH

OH

Maitotoxin-1 Figure 9.3: Structure of Ciguatoxins (CTX) and Maitotoxin-1. P-CTX-1, P-CTX-2 and C-CTX-1 were derived from fish and P-CTX-3C, P-CTX-4B and Maitotoxin-1 were derived from Gambierdiscus spp.

Based on D8-D-10 LSU rDNA phylogenetic analysis, two new putative phylotypes Gambierdiscus ribotype 1 and Gambierdiscus ribotype 2 were reported [45], as the two clusters/clades separated from the others and their genetic distances equaled or exceeded those among the 11 described species [45] (Table 9.1). More

N/K

N/D [, , ]

N/K

HELA positive [] MBA positive [, , ], RBA positive []

N/K

N/D []

NCBA positive []

N/K

North Carolina, USA, Belize – Caribbean []

French Polynesia [], Japan [], Cook Islands [], Hawaii USA [], Pakistan [], Canary Islands []

Canary Islands []; Brazil [, ]; Madeira []; Florida, USA []

Cook Islands []

Fukuyoa ruetzleri

Gambierdiscus australes

Gambierdiscus excentricus

Gambierdiscus honu

N/K

NA positive []

N/K

N/D []

N/K

N/D [, ]

N/K

Singapore [], Japan [], Mexican MBA positive [] Caribbean [], Queensland, Australia [], Nha Trang – Vietnam []

MNA negative []

Fukuyoa yasumotoi

CTX

Brazil []

MTX

Fukuyoa paulensis

CTX

Various assays

Toxicity

Geographical distribution

Species

Table 9.2: Geographic distribution and toxicity of different Gambierdiscus species.

N/D []

N/K

Yes [, ]

N/K

N/D

N/K

MTX

LCMS

314 Hazel Farrell, Gurjeet S. Kohli and Shauna A. Murray

N/K

N/K

MBA positive [, ]

N/K

CA+ positive []

N/K

MBA positive [, ]

Cook Islands []

Great Barrier Reef, Australia []

Heron Island, Queensland, Australia []

Japan []

Gambierdiscus cheloniae

Gambierdiscus lapillus

Gambierdiscus lewisii

Gambierdiscus scabrosus

CA+ positive []

N/D []

HELA positive []

Belize [], Florida [], Mexican Caribbean RBA positive [] [], Malaysia [], Pakistan [], Queensland, Australia (Murray unpubl. Data), St. Barthelemy Island – Caribbean [], Red Sea []

Gambierdiscus belizeanus

N/K

N/D []

N/D []

N/D []

N/K

N/D [, ]

Manado, Indonesia Celebes Sea (SW Pacific MBA positive [], NCBA N/K Ocean) [], Marakei Island, Republic of positive (Celebes Sea), NCBA Kiribati and Rawa Island, Malaysia [] negative (Malaysia) []

MBA positive []

Gambierdiscus balechii

MBA positive []

French Polynesia [], Marshall Islands & Society Islands Micronesia [], Kota Kinabalu and Sipandan Island [], Nha Trang – Vietnam []

Gambierdiscus pacificus

N/K

(continued )

N/D for MTX-, putative MTX- (methylgamberone was detected) []

N/D []

N/D []

N/D []

N/K

N/K

9 Gambierdiscus, the cause of ciguatera fish poisoning

315

N/K

N/K

N/K

N/K

CA+ positive []

Florida, Belize – Caribbean, Tahiti, Palau, Hawaii [], Flower Gardens – Gulf of Mexico, Osho Rios – Jamaica [], Bahamas, Grand Caymam Island, Tol-truk Micronesia [], Jeju Island Korea []

North Carolina, USA, Atlantic ocean [], Bermuda, Mexico [], Puerto Rico, Flower Gardens – Gulf of Mexico, Osho Rios – Jamaica, Crete – Greece []

French Polynesia [] Belize, Guam, Fiji [], Hawaii [], Dry Tortugas – Florida, Flower Gardens – Gulf of Mexico, Osho Rios – Jamaica [], New South Wales, Australia []

Jeju Island, Korea []

Heron Island, Queensland, Australia []

Gambierdiscus caribaeus

Gambierdiscus carolinianus

Gambierdiscus carpenteri

Gambierdiscus jejuensis

Gambierdiscus holmesii

CTX

Geographical distribution

Species

Table 9.2 (continued )

CA+ positive []

N/K

HELA positive [], CA+ positive []

HELA positive []

HELA positive []

MTX

Various assays

Toxicity

N/D []

N/K

N/D []

N/K

N/K

CTX

N/D for MTX-, putative MTX- (methylgamberone was detected) [, ]

N/K

N/D []

N/K

N/K

MTX

LCMS

316 Hazel Farrell, Gurjeet S. Kohli and Shauna A. Murray

French Polynesia [], Canary Islands [], Pakistan [], Nha Trang – Vietnam [], Cook islands [, ]

Tahiti, French Polynesia [, ], Mexican Caribbean [], New Caledonia, Reunion Island, Indian Ocean [], Nha Trang – Vietnam [, ] MBA negative [], RBA positive []

N/K

MBA positive [], RBA positive []

MBA positive []

MBA negative []

Gambierdiscus Japan [] sp. type  MNA positive []

MNA positive []

Gambierdiscus Marakei, Republic of Kiribati [] sp. type 

N/K

N/K

MBA positive []

MBA positive []

HELA positive []

MBA positive []

N/K

MBA positive []

N/K

N/K

N/K

N/K

N/K

N/K

N/K

Yes [, , ]

N/K

N/K

N/K

N/K

N/K

N/K

N/K

N/K

The abbreviations are: N/K, not known, N/D, not detected, MBA, mouse bioassay, RBA, receptor-binding assay, HELA, human erythrocyte lysis assay, NCBA, neuro-2a cell-binding assay, MNA, mouse neuroblastoma assay., CA2+, Ca 2+ influx SH-SY5Y cell FLIPR bioassay.

Gambierdiscus sp. type 

Marakei, Republic of Kiribati []

Gambierdiscus sp. type 

Japan []

Gambierdiscus Belize – Caribbean, Martinique – Caribbean N/K ribotype  [], Puerto Rico []

G. toxicus

Gambierdiscus Belize – Caribbean [] silvae

Gambierdiscus polynesiensis

9 Gambierdiscus, the cause of ciguatera fish poisoning

317

318

Hazel Farrell, Gurjeet S. Kohli and Shauna A. Murray

recently, Gambierdiscus ribotype 1 was described as G. silvae [44] (Table 9.1). Similarly, three new putative species/phylotypes of Gambierdiscus (Gambierdiscus sp. type 1, type 2 and type 3) were reported from Japan based on the differences in the regions D8–D10 of the LSU and SSU rDNA [42, 43] (Table 9.1). In that case, the p distances between these two novel clades and known species of Gambierdiscus were larger than those separating F. yasumotoi from F. ruetzleri. Although the genetic data indicates that these phylotypes may be new species, their morphological circumscriptions are needed to support their status as new species. As sampling around the world becomes more intensive, it is likely that new species of Gambierdiscus will be described. For example, Gambierdiscus sp. type 2 has recently been described as G. jejuensis [41] and [46] reported three additional putative species/phylotypes (Gambierdiscus sp. type 4, Gambierdiscus sp. type 5 and Gambierdiscus sp. type 6) from the Republic of Kiribati (Marakei), of which Gambierdiscus sp. type 6 was subsequently described as G. balechii [35].

9.3 Geographic distribution and abundance Gambierdiscus is widely distributed in coastal zones at tropical and subtropical latitudes. The distribution of species of Gambierdiscus is still poorly understood, as the discrimination of different species of Gambierdiscus has only occurred recently (Table 9.2).

9.3.1 The Pacific and Indian Ocean regions Gambierdiscus is named after the Gambier Islands in French Polynesia, where it was first identified [6], and since then, G. toxicus, G. belizeanus, F. yasumotoi, G. australes, G. pacificus, G. polynesiensis, G. caribaeus and G. carpenteri have been reported from various Pacific islands, Hawaii, Australia, Southeast Asia and the Northern Indian Ocean (Table 9.2). Recently, three genetically distinct types from coastal and temperate waters of Japan were reported [43]. One of these types was recently described as G. jejuensis [41]. G. scabrosus has also been documented in Japanese coastal waters [40]. Three genetically distinct types of Gambierdiscus from the Republic of Kiribati were reported by [46], of which one was later described as G. balechii [34, 35] (Table 9.2). In addition, Gambierdiscus has been reported from the Philippines [20], Hong Kong [67], Indonesia [68] and Mauritius [69], although species diversity in these areas is not known. Gambierdiscus has also been reported from the Mexican Pacific coast [70] and regions around Madagascar [71], where cases of CFP have also been previously reported [72, 73].

9 Gambierdiscus, the cause of ciguatera fish poisoning

319

9.3.2 The Atlantic Ocean region Early accounts of Gambierdiscus “look-alike” species date back to 1948 from Cape Verde Islands [74] and 1979 from Key Largo, Florida [75]. So far, G. toxicus, G. belizeanus, F. yasumotoi, G. polynesiensis, G. caribaeus, G. carolinianus, G. carpenteri, F. ruetzleri, G. excentricus, Gambierdiscus silvae and Gambierdiscus ribotype 2 have been reported from the east coast of the USA, Caribbean and the Mediterranean regions (Table 9.2). There are many other regions where Gambierdiscus has been reported; however, the exact species are yet to be determined. These include Cyprus, Rhodes, Saronikos Gulf [76, 77], French West Indies [78], Cuba [79] and Veracruz [80]. Other confirmed reports of Gambierdiscus occurrence in Central and South America in the literature are from Costa Rica and Brazil (M. Montero pers. comm. in [81]). From Africa, there has been a direct observation of Gambierdiscus, from the coast of Angola [82]; however, CFP cases from the west coast (Canary Islands and Cameroon) [83] of Africa have been reported, and Gambierdiscus species have been reported in that region [12, 44]. Certain species of Gambierdiscus have been designated as being endemic to either the Pacific or the Atlantic regions [45, 82]. So far, G. pacificus, G. honu, G. cheloniae, G. lapillus, G. holmesii, G. lewisii have only been reported from the Pacific, and F. ruetzleri, G. excentricus and G. ribotype 2 are only reported from the Atlantic region (Table 9.2). G. belizeanus, G. caribeaus, G. carpenteri and G. carolinianus are widely distributed in the Atlantic and the Pacific Oceans [32, 45, 82]. Recently, the type species for Fukuyoa, F. paulensis, was described from Brazil [15]. F. yasumotoi is widely distributed in the Pacific; however, there is only one report of its occurrence in the Mexican Caribbean [47], which was reported before the discovery of the other globular species F. ruetzleri, which is widely distributed in the Atlantic region [32]. The distribution of G. toxicus needs to be refined due to numerous misidentifications in the literature. G. polynesiensis is widespread in the Pacific (Table 9.2) with only one confirmed report from the Canary Islands in the Atlantic region [12]. Both, Litaker et al. [45], and Berdalet et al. [82], mention that none of the Pacific-specific species have been observed in hundreds of field samples analyzed from the Atlantic regions (Caribbean/ Gulf of Mexico/West Indies/Southeast US coast from Florida to North Carolina). The absence of Atlantic-specific species in the Pacific region has not been confirmed, as the majority of the vast Pacific region remains unexplored. As under-sampling and under- reporting have occurred worldwide, but particularly in the Pacific region, much more work needs to be undertaken in order to determine whether endemism or restricted distributions exist in species of Gambierdiscus. While multiple species of Gambierdiscus can co-occur in one region, equally, there are regions from where only one species has been reported. For example, in Heron Island (Queensland, Australia) there are at least three species of Gambierdiscus that cooccur, however, further south in Merimbula, New South Wales only G. carpenteri is known to occur [65]. Localized benthic blooms of Gambierdiscus have been noted in

320

Hazel Farrell, Gurjeet S. Kohli and Shauna A. Murray

the literature from both the Pacific and Atlantic regions [84–87]. Cell densities in such blooms can range from anywhere between 10,000 to 100,000 cells g–1 wet weight algae [45]. There are no accurate estimates of cell densities at which a Gambierdiscus bloom leads to a CFP epidemic. The onset of CFP may also depend on other factors, such as the fact that different species of Gambierdiscus have varying toxicities. For example, in 2010, an unidentified species of Gambierdiscus was reported in Greece, however, no CFP outbreaks have been reported there [88]. In most habitats where species of Gambierdiscus occur, cell densities are below 1,000 cells g–1 wet weight algae [45], however, in some environments Gambierdiscus spp. are known to occur year-round at such cell densities [84]. A constant exposure of low densities of cells could also lead to a buildup of CFP-related toxins in fish. Extensive research needs to be done to understand the relationship between Gambierdiscus abundance and CFP outbreaks. This is particularly challenging, as benthic dinoflagellates inhabit areas where quantitative sampling of microbial eukaryotes is not straightforward, for example, in sediments and on the surface of dead corals. Also, Gambierdiscus cell distribution can be very patchy, even over small distances, making it hard to estimate mean Gambierdiscus cell densities over a larger area [45, 78, 89].

9.4 CTXs and MTXs CTXs are sodium channel activators, particularly affecting the voltage-sensitive channels located along the nodes of Ranvier (peripheral nerve cells) [90–92]. When the sodium channels are activated, there is a massive influx of Na + ions, resulting in cell depolarization [90–92]. This leads to the onset of spontaneous action potentials in affected cells, causing various symptoms in humans. Symptoms can include but are not limited to gastrointestinal, neurological and in cases of severe intoxication, sometimes, cardiovascular [21], and can vary depending on geographical region [92, 93]. This can be due to the structural differences of CTXs in different regions; therefore, it is very important to characterize CTXs from the Pacific, Caribbean and the Indian Oceans. Local understanding of CTX accumulation patterns in different fish species can also help prevent CFP. However, the accurate identification of exact congeners of CTXs is necessary to understand the toxicology and evaluate the local risks of CFP. Structurally, CTXs are thermostable, cyclic polyether ladders, which are liposoluble (Figure 9.3). They have been isolated from fish and different species of Gambierdiscus (Table 9.3). Based on their origin and differences in the structure of these toxins, they are divided into P-CTXs (Pacific Ocean), C-CTXs (Caribbean region) and I-CTXs (Indian Ocean). Due to their structural differences, P-CTXs are further divided into type I and type II, as suggested by Legrand et al. [94]. Type I P-CTXs have 13 rings and 60 carbon atoms [95–98]. This category consists of the

Caribbean

Pacific (type II)

Pacific (type I)

Ciguatoxins

Origin

Moray eel (Lycodontis javanicus, Muraenidae) [] Moray eel (Lycodontis javanicus, Muraenidae) [] Gambierdiscus sp. [] G. polynesiensis [] Gambierdiscus sp. [] G. polynesiensis [] Gambierdiscus sp. [] G. polynesiensis [] Gambierdiscus sp. [] G. polynesiensis [] Gambierdiscus sp. [] G. polynesiensis [] Horse-eye jack (Caranx latus)

. [] . [] . [] . [] . [] . [] . [] . [, ]

CTX- CTX- CTXA

CTXB

CTXC -epi-CTX-C M-seco-CTX-C

C-CTX-

Moray eel (Gymnothorax javanicus) [] Moray eel (Lycodontis javanicus, Muraenidae) []

Source

. [, ]

Molecular weight

CTXB [], CTX[]

Toxin name

Table 9.3: Different congeners of CTXs and MTXs.

(continued )

. μg kg– []

 μg kg– []

 μg kg– []

. μg kg– []

 μg kg– []

 μg kg– []

. μg kg– []

. μg kg– []

CTXB: . μg kg– [] CTX-: . μg kg– []

Toxicity*

9 Gambierdiscus, the cause of ciguatera fish poisoning

321

Pacific

Maitotoxins and methylgambierone

Indian *LD doses calculated via i. p. injection in mice.

Origin

Table 9.3 (continued )

Horse-eye jack (Caranx latus) Red bass (Lutjanus bohar) [] and red emperor (Lutjanus sebae) [] Bull shark (Carcharhinus leucas) []

. [, ] . [, ] . [] . [] . [] . [] . []

C-CTX-

Gambierdiscus spp. []

Gambierdiscus sp. []

 []

MTX-  [] methylgambierone

Gambierdiscus sp. []

 [, ]

MTX-

I-CTX- I-CTX- I-CTX- I-CTX- I-CTX- I-CTX-

Source

Molecular weight

Toxin name

Toxic [, ].

. μg kg– []

. μg kg– []

Toxic [, ], ]

Toxic []

Toxicity*

322 Hazel Farrell, Gurjeet S. Kohli and Shauna A. Murray

9 Gambierdiscus, the cause of ciguatera fish poisoning

323

first CTX to be fully structurally described as CTX1B [95] (or CTX-1 as described by Lewis et al. [96]) from moray eels, which is the principal toxin in the carnivorous fish from the Pacific [95, 96]. Two other type I P-CTXs, that is, CTX-2 and CTX-3, were also described from the same extracts; they have slight variations in their structures leading to different toxicities in mice [96] (Table 9.3). It has also been suggested that CTX-1, CTX-2 and CTX-3 may be derived from dinoflagellate precursors known as CTX-4A and CTX4B (also named as GTX-4B in [95]) [97, 98]. Recently, CTX-4A and CTX-4B have been isolated from G. polynesiensis culture extracts [10]. CTX3C is a type II P-CTX with 13 rings, 57 carbon atoms and was first isolated from cultures of Gambierdiscus sp [99] and later from G. polynesiensis [10]. Two more congeners of CTX3C called as 49-epi-CTX-3C (also called as CTX-3B in [10]) and M-seco-CTX-3C have also been isolated from Gambierdiscus sp [99] and G. polynesiensis [10]. Later, 2 new type II P-CTXs, that is, 2,3 dihydroxyCTX3C (also called as CTX2-A1) and 51-hydroxyCTX3C, were isolated from Moray eel [100] that might be oxygenated metabolites of CTX3C [98]. Caribbean CTXs are slightly bigger than P-CTXs and have 14 rings and 62 carbon atoms [101–104]. Many congeners of C-CTXs have been isolated from carnivorous fish including C-CTX1, C-CTX-2, C-CTX -1141, C-CTX-1127, C-CTX-1143, C-CTX-1157, C-CTX-1159 [101–104]. Unlike P-CTXs, there have been no reports of C-CTXs originating from Gambierdiscus sp. However, recently G. excentricus has been identified as a major CTX producer in the Caribbean [12], and CTXs from this strain are being characterized. Recently, 4 CTXs (I-CTX-1, I-CTX-2, I-CTX-3, I-CTX-4) have been isolated from carnivorous fish from the Indian Ocean and have higher molecular ion masses than P-CTXs and C-CTXs [88, 105, 106]. However, their structures need to be elucidated [105, 106]. I-CTX-1 is toxic to mice via intraperitoneal injection [106]. Based on mouse bioassays (MBA), different congeners of CTXs can have variable toxicities (Table 9.3), however, this needs to be further validated as well. Maitotoxins are one of the largest nonproteinous and highly toxic natural products known [109, 111]. This polyether ladder-type compound was first discovered as a water-soluble toxin in the stomach contents of herbivorous fish Acanthurids (surgeonfish) in 1976 [112]. In the 1990s, stereoscopic studies and partial synthesis were used to determine the structural elucidation and stereochemistry of the extraordinary complex and large MTX [109, 113–117]. Simultaneously, Holmes and Lewis described two large (MTX-1, MTX-2) and one small MTX (MTX-3) from different strains of Gambierdiscus sp. isolated from Queensland, Australia [110] (Table 9.3). MTX-3 has recently been described as 44-methylgambierone [14]. MTX-1 from this study may have been the MTX originally described from the stomach contents of Acanthurids; however, it is not clearly proven. When compared to other natural toxins, MTX is a highly potent calcium channel inhibitor (LD50 0.05 μg kg–1, i.p., mice), only exceeded by a handful bacterial proteinous toxins [109, 111]. Despite its high level of potency, the complete mode of action and the primary target of MTX in mammalian cells have not yet been fully elucidated. In fact, the activation of voltage-dependent

324

Hazel Farrell, Gurjeet S. Kohli and Shauna A. Murray

calcium channels induced via MTXs is a secondary effect of membrane depolarization (for review and more details see [118]). Recently, it has been reported that the biophysical mechanisms of pacific MTXs are different to Caribbean MTXs [119]. Whether this is due to a structural difference is not known, as the Caribbean MTXs have not been fully characterized. Although MTX appears to have a low tendency of accumulating in fish flesh, as compared to stomach or intestines [112], its possible role in CFP cannot be disregarded, as eating non-eviscerated fish is a common practice in many Pacific Island nations. The sulphate esters in the structures of MTXs make it amenable to detect and quantify MTX by LC_ESI_MS (liquid chromatography-electrospray ionization-mass spectrometry) (T. Harwood, pers. comm.) and solvolysis (desulfonation) reduces the toxicity of MTXs significantly, at least 100-fold [120]. However, more research is essential to understand the exact role of MTXs in CFP including its mode of action and target in mammalian cells. Cyclic polyether ladders are almost exclusively known to be produced by dinoflagellates. Other than CTXs and MTXs, this class of secondary metabolites also includes brevetoxins (BTXs), produced by Karenia spp. [121] and yessotoxins (YTXs), produced by a wide array of dinoflagellates including Lingulodinium polydrum [122], Gonyaulax spinifera [123] and Protoceratium reticulatum [124]. Based on their high structural similarities, the synthesis of these compounds likely involves common biosynthetic mechanisms [125–127]. Stable isotope labeling of precursors to elucidate the biosynthesis pathway of CTXs and MTXs has never been performed. However, precursor studies to reveal the biosynthesis pathways of BTXs and YTXs have indicated the polyketide origin of these cyclic polyether ladders [128, 129]. Several schematic pathways involving different enzymes have been suggested and are detailed in Kalaitzis et al. [130] and Kellmann et al. [129]. It is speculated that the biosynthesis involves the normal polyketide synthase (PKS) enzyme complex with a few additional enzymes, that is, epoxidases and thioesterases [131]. Essential domains present in the PKS are: acyltransferase domain (AT); β-ketosynthase domain (KS); and acyl carrier protein (ACP) [132]. In addition, PKS can include βketoacyl reductase (KR), enoyl reductase (ER) and dehydrogenase (DH) domains [132]. In the past 10 years, a few genes that encode the essential domains of the PKSs, particularly KS domains in dinoflagellates, have been identified for the first time. However, with the availability of next-generation sequencing tools, a few candidate genes encoding KS and KR domains in Karenia brevis have been associated with biosynthesis of BTXs [133]. A recent study published a comprehensive transcriptome library of Lingulodinium polydrum for which genes encoding KS domains were reported, however, no link between these genes and YTX production has been established [134]. In the past, a few studies have identified genes encoding KS domains in Amphidinium sp [135], which produces numerous macrolides (cyclized linear polyethers) such as amphidinolides. No studies have been done to identify genes involved in CTX and MTX biosynthesis. However, an extensive marine microbial eukaryote transcriptome project, undertaken by the Moore Foundation, is

9 Gambierdiscus, the cause of ciguatera fish poisoning

325

in the process of sequencing 652 trancriptomes, which includes two strains of Gambierdiscus species. Analysis of data obtained for such diverse arrays of dinoflagellate species may shed light on the genes involved in secondary metabolite synthesis in dinoflagellates.

9.5 Toxicity of different species of Gambierdiscus There is clear evidence that Gambierdiscus species produce CTXs and/or MTXs [28, 95, 99, 110, 136]. However, many wild and cultured strains of Gambierdiscus have not been found to produce detectable amounts of CTXs [27, 137]. The CTX analogue that has shown the most potent toxicity in fish in the Pacific region is P-CTX-1B [138, 139] while other CTX analogues found in ciguatoxic fish in the Pacific are 52-epi-54-deoxyCTX1B, 54-deoxyCTX1B, 2-hydroxyCTX3C, and 2,3-dihydroxyCTX3. These analogues originate from the oxidization by fish liver enzymes during digestion of the Gambierdiscusproduced CTX analogues P-CTX4A, 4B and 3C [139], P-CTX4A, 4B, and 3C have been detected in multiple cultures of G. polynesiensis, in SPATT bags attached to macroalgae in a CFP endemic locations, and in herbivorous fish in areas of endemic CFP toxicity [64, 138, 140]. The species G. polynesiensis and G. excentricus produce compounds that are several orders of magnitude higher in their CTX-activity than those of other cooccurring Gambierdiscus species [10, 16, 50, 51]. It is unclear if species other than G. polynesiensis and G. excentricus are contributing to CTX toxicity in fish, or whether only a few species are responsible for this toxicity in CFP endemic regions. Unfortunately, many older studies of CTX toxicity in Gambierdiscus strains document the identity of the cultures as Gambierdiscus toxicus, since it was the only known species of Gambieriscus at that time. It is imperative to study the toxin profile of all the species and genotypes now known (Table 9.2). Table 9.2 provides the data available on the toxicity of each species of Gambierdiscus detected via cell bioassays and LC-MS (liquid chromatography-mass spectrometry)-based detection. In 2010, Chinain et al. [10] described the toxin profile of G. polynesiensis based on LC-MS analysis and receptor-binding assay (RBA). This species produces both Type 1 (CTX-4A, CTX-4B) and Type 2 P-CTXs (CTX-3C, M-seco-CTX-3C, 49-epiCTX-3C), however, P-CTX-3C was the major toxin produced by this species. Two different strains of G. polynesiensis were tested and found to produce same suite of toxins, in different proportions [10]. Similar results were found in a study of 56 strains of six different species (G. belizeanus, G. caribaeus, G. carolinianus, G. carpenteri, Gambierdiscus ribotype 2) over a period of 2 years, using the human erythrocyte lysis assay (HELA) [13]. The intraspecific toxicity varied slightly among different strains of same species; however, the level of toxicity of each strain remained unchanged over the period of the study [13]. HELA assay toxicity is indicative of MTX production by the species. The water-soluble fraction of the extracts of G. polynesiensis has been found to be toxic via MBA [9], indicating the

326

Hazel Farrell, Gurjeet S. Kohli and Shauna A. Murray

presence of MTXs. However, the toxins that produced this effect have not been characterized from this strain. Another species from the Caribbean, G. excentricus, may produce CTXs and MTXs (as determined via Neuro-2a cell-based assay) [12], however, the exact toxin profile needs to be verified via LC-MS analysis. The toxicities of the liposoluble and water-soluble fractions of G. australes extracts, isolated from the Cook Islands, were found to be toxic via MBA. This indicated the presence of CTXs and MTXs, however, no CTXs were detected via LC-MS analysis [11]. Another strain of G. australes from French Polynesia tested positive for CTXs via the RBA; however, the level of toxicity was low when compared to G. polynesiensis [10]. These results are intriguing and require further analysis. While bioassays are important to determine toxicity, only LC-MS-based analysis techniques can determine the exact toxin profile of different species of Gambierdiscus. As we only know the partial toxin profiles of two species of Gambierdiscus via LC-MS-based techniques, this area of research needs urgent attention.

9.6 Detection of CTXs and MTXs in seafood Originally, CFP was derived from the word “cigua,” used by native Cubans to describe a turban-shelled snail, implicated in an outbreak of the sickness in Spanish explorers to Cuba in the 1500s [141]. The occurrence of CTX in the turban snail Turbo argyrostoma has since been confirmed [142], however, most CFP cases have followed consumption of large reef fish (e.g., [143–146]). This circumstance has been a critical factor in the diagnosis of the disease, as in many cases there has been no fish sample retained for chemical verification or the appropriate test facilities have not been available. Although hundreds of cases of CFP have been documented worldwide, it is estimated that less than 20% of actual cases have been reported [147]. There is a high likelihood of misdiagnosis for CFP. The number of documented symptoms, which are in excess of 175 [21], may vary depending on portion size [148], individual susceptibility or accumulation of toxin with age [19, 149] and could also be associated with other illnesses (e.g., decompression sickness [150], chronic fatigue syndrome, multiple sclerosis [151, 152] and brain tumors [151]). The number of fish species implicated in ciguatera outbreaks is suggested to be of the order of several hundred [153, 154]. With the abovementioned limitations and the absence of a reliable, commercially available test kit, it is difficult to express an exact figure. While carnivorous fish are the main source of CTXs in reported cases of CFP, herbivorous fish (e.g., surgeonfish and parrotfish), a key component of the toxic food chain [155, 156], have also been linked to CFP outbreaks. Table 9.4 provides a summary of over 180 fish species and other marine fauna that have tested positive for CTXs, from ciguatera-prone regions and following reported outbreaks.

C-CTX- []

Canary Islands [, , ], Hawaii [, ], St. Thomas, Caribbean Sea []

Tenerife, Spain []

Seriola rivoliana (Almaco jack,-Kahala)

Seriola spp.

Sphyraena barracuda (Great barracuda)

Barracuda

The Bahamas [], West Africa [], Florida C-CTX- [, ] Keys, USA [], French West Indies [], St. Barthelemy, Caribbean Sea [, ], Guadeloupe [], French Polynesia [], Meridia, Mexico []

C-CTX- []

C-CTX- [], CTX-B [], CTX-C and CTX analogues from Caribbean or Indian waters) [], C-CTX- and three CCTX- metabolites []

Selvagens Islands (Madeira, Portugal) [, ], West Africa []

Seriola fasciata (Lesser amberjack)

C

C-CTX- [], CTX-B [], CTX-C and CTX analogues from Caribbean or Indian waters []

CTX (if detected)

Canary Islands [–], Selvagens Islands (Madeira Portugal) [], Hawaii [–], Haiti [], St. Barthelemy, Caribbean Sea [], St. Thomas, Caribbean Sea []

Source

Seriola dumerili (Greater amberjack, Kahala)

AmberjackC

Latin name (Common name)

Table 9.4: Different congeners of CTXs detected by various assays in seafood and other animals.

(continued )

Cat BA [], Chick BA [], MQBA [], MBA [, ], NA []

LCMS/MS []

LCMS/MS [], BSBA [], MGBA [], ELISA [, ], NA [, , , , ]

LCMS/MS [], ], UPLC/MS) [], NA []

UPLC/MS [], HPLC/MS [], TLC [], BSBA [], MGBA [, ], MBA [, , ], S-EIA [], SPIA [], RIA [], ELISA [, ], NA [–, ], RBA []

Method of detection

9 Gambierdiscus, the cause of ciguatera fish poisoning

327

P-CTX- []

Implicated in CFP investigations in Queensland, Australia (specific source location unconfirmed) []

Sphyraena putnamae (Sawtooth barracuda)

Butterflyfish and AngelfishO, C

Priacanthus hamrur (Lunar tailed bigeye)

BigeyeC

Lates calcarifer (Asian seabass, Barramundi)

French Polynesia []

CTX – positive []

P-CTX- P-CTX- P-CTX- []

MBA, NA []

CTX – positive []

South Taiwan []

Sphyraena spp. (Barracuda fish eggs)

Imported to Hong Kong (source unconfirmed) []

S-EIA [], SPIA [], NA [], LCMS/MS []

C-CTX- []

California [] imported to New York City (source unconfirmed) []

Sphyraena sp. (Barracuda)

BarramundiC

TLC, MBA []

CTX – positive []

Sphyraena jello Hervey Bay, Queensland, Australia [] (Pickhandle barracuda)

RBA []

MBA, LCMS/MS []

HPLC-MS/MS []

Method of detection

Source

CTX (if detected)

Latin name (Common name)

Table 9.4 (continued )

328 Hazel Farrell, Gurjeet S. Kohli and Shauna A. Murray

CTX positive []

P-CTX- []

P-CTX-, P-CTX-, P-CTX- []

Southern coast of China []

Rep. Kiribati []

Rep. Kiribati []

Chaetodontoplus septentrionalis (Blue-striped angelfish)

Forcipiger longirostris (Longnose butterflyfish)

Pomacanthus imperator (Emperor angelfish)

Abudefduf sexfasciatus Southern coast of China [] (Stripetailed damselfish)

Cigua-check®, MBA []

LCMS/MS []

LCMS/MS []

Cigua-check®, MBA []

Cigua-check®, MBA []

(continued )

CTX positive []

Southern coast of China []

Chaetodon strigangulus (Chevron butterflyfish)

CTX positive []

P-CTX-, P-CTX-, P-CTX- []

Rep. Kiribati []

Chaetodon meyeri (Meyer’s butterflyfish)

Damselfish and ClownfishO

Cigua-check®, MBA []

CTX – positive []

Southern coast of China []

Chaetodon auripes (Oriental butterflyfish) LCMS/MS []

Cigua-check®, MBA [], LCMS/MS []

P-CTX-, P-CTX-, P-CTX- []

Southern coast of China [], Rep. Kiribati []

Chaetodon auriga (Threadfin butterflyfish)

9 Gambierdiscus, the cause of ciguatera fish poisoning

329

CTX positive [] CTX positive [], ]

French Polynesia []

French Polynesia [, ]

Lethrinus olivaceus (Longface emperor)

Southern coast of China []

Lethrinus atkinsoni (Yellowtail emperor)

Emperor breamC

Myrichthys maculosus (Spotted snake eel)

CTX positive []

RBA [, ]

RBA []

Cigua-check®, MBA []

HPLC/MS [, ], HPLC/HNMR [, , ], TLC [], DLBA [], MBA [, , ], LCMS/MS []

CTX- [, ], CTX-B [, ], CTX-[], CTX- [], P-CTX- [, ], P-CTX- [, ], P-CTX- [, ], analogues of CTX C: ,-dihydroxyCTXC and -hydroxyCTXC []

Gymnothorax javanicus Tuamotu Archipelago and Tahiti (French (Moray eel) Polynesia) [, , ], Rep. Kiribati [, ], Hawaii []

TLC, MBA []

CTX positive []

St. Barthelemy, Caribbean Sea []

Gymnothorax funebris (Green moray)

LCMS/MS []

Method of detection

P-CTX-, P-CTX-, P-CTX- []

CTX (if detected)

Rep. Kiribati []

Source

Gymnothorax flavimarginatus (Yellowmargin morray)

EelC

Latin name (Common name)

Table 9.4 (continued )

330 Hazel Farrell, Gurjeet S. Kohli and Shauna A. Murray

P-CTX- []

French Polynesia [], Rep. Kiribati []

Monotaxis grandoculis (Bigeye bream)

CTX positive []

CTX positive []

CTX positive []

P-CTX- []

CTX positive []

Hawaii []

St. Barthelemy, Caribbean Sea []

French Polynesia []

Rep. Kiribati []

French Polynesia []

Mulloidichthys auriflamma (Goldstriped goatfish)

Mulloidichthys martinicus (Yellow goatfish)

Parupeneus barberinus (Dash and dot goatfish)

Parupeneus bifasciatus (Doublebar goatfish)

Parupeneus ciliatus (Diamondscale goatfish)

GoatfishC

P-CTX-B [], P-CTX- []

French Polynesia [], Capel Bank Seamount [], implicated in CFP investigations in Queensland, Australia (specific source location unconfirmed) []

Lethrinus miniatus (Trumpet emperor, Redthroat emperor, Sweetlip emperor)

RBA []

LCMS/MS []

RBA []

TLC, MBA []

S-EIA, SPIA []

(continued )

Cat BA [], MQBA [], MBA [], LCMS/MS []

Cat BA [], MQBA [], MBA [], LCMS/MS [], HPLC-MS/MS []

9 Gambierdiscus, the cause of ciguatera fish poisoning

331

CTX positive []

CTX positive []

Nuku Hiva (Marquesas) []

Southern coast of China []

Parupeneus insularis (Twosaddle goatfish)

Upeneus bensai (accepted name = U. japonicus) (Japanese goatfish)

CTX positive []

CTX positive []

French Polynesia []

New Caledonia, French Polynesia []

Rep. of Vanuatu []

Tridacna maxima (Giant clam)

Tridacna sp. (Giant clam)

Hippopus hippopus (Giant Clam)

CTX positive []

P-CTX-B [], Darius et al., ], P-CTX -C, P-CTX-A, P-CTX-B [Darius et al., ]

French Polynesia [], Darius et al., ]

Tectus niloticus (Trochus, top shell)

Giant clamH

CTX positive []

Hawaii []

Conus spp. (Cone snails)

GastropodC

CTX (if detected)

Source

Latin name (Common name)

Table 9.4 (continued )

NA, RBA []

MBA, NA, RBA []

NA, RBA []

NA, LCMS/MS [], Darius et al., ]

Ciguatect® []

Cigua-check®, MBA []

RBA []

Method of detection

332 Hazel Farrell, Gurjeet S. Kohli and Shauna A. Murray

P-CTX- []

CTX positive []

P-CTX-, P-CTX-, P-CTX- []

Rep. Kiribati []

Southern coast of China []

Rep. Kiribati []

Cephalopholis sexmaculata (Six blotch hind)

Cephalopholis sonnerati (Red coral cod)

Epinephelus coeruleopunctatus (White-spotted grouper) CTX positive []

P-CTX- [, , ]

Fiji [, ], Arafura Sea, Australia []

Cephalopholis miniata (Coral rockcod)

Epinephelus corallicola Southern coast of China [] (Coral grouper)

P-CTX-, P-CTX-, P-CTX- []

P-CTX- []

Hawaii [], French Polynesia [, , , ], Rep. Kiribati []

Implicated in CFP investigations in Queensland, Australia (specific source location unconfirmed) []

Cephalopholis argus (Blue-spotted grouper, Roi)

GrouperC

Scomberoides commersonnianus (Giant Queenfish)

Giant queenfishC

Cigua-check®, MBA []

LCMS/MS []

Cigua-check®, MBA []

LCMS/MS []

(continued )

HPLC/MS [], MBA [], NA [, ], LCMS/MS []

Cat BA [], MQBA [], MBA [], ELISA [], NA [, ], RBA [, , ]

HPLC-MS/MS []

9 Gambierdiscus, the cause of ciguatera fish poisoning

333

Source

CTX-B [], P-CTX-, P-CTX-, P-CTX- [], P-CTX- []

Japan [], Rep. Kiribati [], implicated in CFP investigations in Queensland, Australia (specific source location unconfirmed) []

Epinephelus fuscoguttatus (Brown-marbled grouper)

Epinephelus lanceolatus (Giant grouper)

Imported to Hong Kong (source unconfirmed) [, ]

P-CTX-, P-CTX-, P-CTX- []

C-CTX- []

C-CTX- []

Tenerife, Spain [], French Polynesia []

Epinephelus fasciatus (Dusky grouper)

Epinephelus guttatus US Virgin Islands [] (Red hind, Koon, Lucky grouper)

CTX positive []

Imported to Hong Kong (source unconfirmed) []

P-CTX-B []

CTX (if detected)

Epinephelus coioides (Orange-spotted grouper)

Epinephelus Capel Bank Seamount [] cyanopodus (Purple rockcod, Speckled blue grouper)

Latin name (Common name)

Table 9.4 (continued )

MBA [, ], LCMS/MS []

NA, LCMS/MS []

LC/MS [, ], MBA [], HPLCMS/MS []

LCMS/MS [], NA []

MBA []

LCMS/MS []

Method of detection

334 Hazel Farrell, Gurjeet S. Kohli and Shauna A. Murray

RBA [], LCMS/MS []

P-CTX- []

CTX positive []

P-CTX-, P-CTX-, P-CTX- []

CTX positive []

CTX positive [] P-CTX-, P-CTX-, P-CTX- [, ]

French Polynesia [], Rep. Kiribati []

French Polynesia []

Rep. Kiribati []

St. Thomas, Caribbean Sea []

St. Barthelemy, Caribbean Sea []

French Polynesia [, ], imported to Hong Kong (source unconfirmed) [], Rep. Kiribati []

Epinephelus merra (Honeycomb grouper or Dwarf-spotted rockcod)

Epinephelus microdon (Marble grouper)

Epinephelus multinotatus (White blotched grouper)

Epinephelus mystacinus (Misty grouper)

Epinephelus morio (Red grouper)

Epinephelus polyphekadion (Camouflage grouper)

(continued )

RBA [, ], MBA, LCMS/MS [, ], NA []

TLC, MBA []

BSBA, MGBA []

LCMS/MS []

Cat BA, MQBA, MBA []

RBA [], NA [], LCMS/MS []

French Polynesia [], Canary Islands C-CTX- [] [], Selvagens Islands (Madeira, Portugal) []

Epinephelus marginatus (Honeycomb grouper or Dwarf-spotted rockcod)

9 Gambierdiscus, the cause of ciguatera fish poisoning

335

P-CTX-, P-CTX-, P-CTX- []

P-CTX-B [], C-CTX- [], P-CTX- []

P-CTX-, P-CTX-, P-CTX- []

C-CTX- [], C-CTX- []

C-CTX- []

CTX- []

Rep. Kiribati []

Baja California, Mexico [], off Queensland, Australia (between Cooktown and Lizard Island) [], imported to New York City (source unconfirmed) [], implicated in CFP investigations in Queensland, Australia (specific source location unconfirmed) []

Southern coast of China [], Rep. Kiribati []

Key Largo, Florida, USA []

Selvagens Islands (Madeira, Portugal) []

Baja California, Mexico []

Baja California, Mexico []

Epinephelus spilotoceps (Foursaddle grouper) Epinephelus spp.

Epinephelus tauvina (Greasy rockcod)

Mycteroperca bonaci (Black grouper)

Mycteroperca fusca (Island grouper, Comb grouper)

Mycteroperca prionura (Sawtail grouper)

Mycteroperca sp.

CTX positive []

CTX (if detected)

Source

Latin name (Common name)

Table 9.4 (continued )

MBA []

HPLC/MS, MBA []

LCMS/MS, []

LCMS/MS, NA []

Cigua-check®, MBA [], LCMS/MS []

MBA [], LCMS/MS [], LCMS/ MS [], HPLC-MS/MS []

LCMS/MS []

Method of detection

336 Hazel Farrell, Gurjeet S. Kohli and Shauna A. Murray

LC/MS [, , ], MBA [, , ], HPLC-MS/MS []

P-CTX-, P-CTX-, P-CTX- [] CTX-B [, ], -epi-Deoxy CTXB, -Deoxy CTXB [], P-CTX-, P-CTX- [], P-CTX- []

Imported to Hong Kong (source unconfirmed) [, ] Japan [, , ], Rep. Kiribati [], implicated in CFP investigations in Queensland, Australia (specific source location unconfirmed) []

Variola albimarginata (Lyretail) Variola louti (Yellow-edged lyretail, yellow-edged coronation trout)

MBA []

(continued )

MBA [, ], LCMS/MS []

C-CTX- [], C-CTX- and isomers [], CTX congeners, other compounds []

French West Indies []

Serranidae (Grouper)

HPLC/MS, MBA []

CTX- [], CTX- [], CTX- []

Great Barrier Reef, Australia []

Plectropomus spp. (Coral trout)

Cat BA [], MQBA [], MBA [, , , ], RBA [], LCMS/MS []

P-CTX-, P-CTX-, P-CTX- []

MBA [, ], RBA [], HPLC-MS /MS [, ]

P-CTX- [, ], P-CTX-, P-CTX- []

Imported to Hong Kong (source unconfirmed) [, ], French Polynesia [], implicated in CFP investigations in Queensland, Australia (specific source location unconfirmed) []

Plectropomus laevis (Blacksaddled coral grouper)

Plectropomus Imported to Hong Kong (source leopardus (Coral trout/ unconfirmed) [, ], Tahiti [], leopard coral grouper) French Polynesia [, ]

MBA []

CTX positive []

Imported to Hong Kong (source unconfirmed) []

Plectropomus areolatus (Squaretail coral grouper)

Chick BA [], MBA [], NA []

CTX positive [, ]

St. Barthelemy, Caribbean Sea [], French West Indies [, ]

Mycteroperca venenosa (Yellowfin grouper)

9 Gambierdiscus, the cause of ciguatera fish poisoning

337

CTX positive []

CTX positive []

Hawaii []

St. Barthelemy, Caribbean Sea []

Bodianus bilunulatus (Tarry hogfish (a’awa))

Bodianus rufus (Spanish hogfish)

HogfishC

Paracirrhites hemistictus (Whitespot hawkfish)

P-CTX-, P-CTX- []

CTX positive []

French Polynesia []

Plectorhinchus picus (Painted sweetlip)

Rep. Kiribati []

CTX- [], CTX- [], CTX- [], P-CTX- []

Platypus Bay, Queensland, Australia [], implicated in CFP investigations in Queensland, Australia (specific source location unconfirmed) []

Pomadasys maculatus (Blotched javelin, Saddle grunt)

HawkfishC

C-CTX- []

CTX (if detected)

US Virgin Islands []

Source

Haemulon plumierii (White grunt)

GruntC

Latin name (Common name)

Table 9.4 (continued )

TLC, MBA []

MBA, S-EIA []

LCMS/MS []

RBA []

HPLC/MS [], MBA [], HPLC-MS /MS []

NA, LCMS/MS []

Method of detection

338 Hazel Farrell, Gurjeet S. Kohli and Shauna A. Murray

 CTXs (incl. C-CTX-, C-CTX-a, C-CTX-) [], C-CTX- [, ] and C-CTX- [, ] C-CTX- and isomers, CTX congeners [] CTX positive [, ]

CTX positive [] CTX positive [, , ] CTX positive [, ]

French West Indies [, ], St. Barthelemy, Caribbean Sea [, ], The Bahamas [], St. Thomas, Carribean Sea []

French West Indies []

French Polynesia [, ]

French Polynesia, Tubuai (Australes) []

Hawaii [, ], French West Indies []

French Polynesia [, ]

Caranx latus (Horse-eye jack)

Caranx lugubris (Black jack)

Caranx melampygus (Bluefin trevally)

Caranx papuensis (Brassy trevally)

Caranx sp. (Trevally (ulua, papio)) Pseudocaranx dentex (White trevally)

JellyfishO

MBA [], S-EIA [], RBA []

CTX positive [, ]

French Polynesia, Tubuai (Australes) [], St. Barthelemy, Caribbean Sea []

Caranx ignobilis (Giant trevally (ulua))

(continued )

MBA [, ], S-EIA [, ] SPIA [, ], NA [] RBA [], NA []

RBA []

Cat BA [], MQBA [], MBA [], RBA []

HPLC/MS, MBA []

HPLC/MS [–], BSBA [], Cat BA [], MGBA [], MBA [, , ], NA []

MBA []

CTX positive []

SPIA []

LCMS/MS []

French West Indies []

CTX positive []

C-CTX- []

Caranx bartholomaei (Yellow jack)

Jacks and Scads

Hawaii []

Bodianus sp.

C

Selvagens Islands (Madeira, Portugal) []

Bodianus scrofa. (Barred hogfish)

9 Gambierdiscus, the cause of ciguatera fish poisoning

339

American Samoa []

Cnidaria sp.

Scomberomorus cavalla (King mackerel “Coronado” (Kingfish))

MackerelO

Panulirus penicillatus (Green spiny lobster)

LobsterO

Pterois miles/volitans (Lionfish)

LionfishC

Florida, USA [], St. Barthelemy, Caribbean Sea [, ], Guadeloupe []

Rep. Kiribati []

St. Croix and St Thomas/St. John Caribbean Sea [] Caribbean Sea and Gulf of Mexico [], Saint Barthelmy Islands (Caribbean) []

KnifejawO Oplegnathus punctatus Miyazaki, Japan [] (Spotted knifejaw)

Source

Latin name (Common name)

Table 9.4 (continued )

C-CTX- [], C-CTX- []

P-CTX- []

C-CTX- [–, ] C-CTX- []

CTX-C []

CTX positive []

CTX (if detected)

LCMS/MS [], TLC [], Chick BA [], MBA [], NA []

LCMS/MS []

LCMS/MS [–, ], RBA [], NA [–, ]

HPLC/MS []

SPIA []

Method of detection

340 Hazel Farrell, Gurjeet S. Kohli and Shauna A. Murray

Parrotfish

H

Octopodidae

Rep. Kiribati []

P-CTX- []

CTX positive []

Nuku Hiva (Marquesas) []

Liza vaigiensis (Thinlip grey mullet)

OctopusC

CTX positive [, ]

French Polynesia [, ]

CTX positive []

P-CTX- []

Crenimugil crenilabis (Fringelip mullet)

MulletO

Cheilodactylus plessisi French Polynesia [] (Plessis’ morwong)

MorwongC

Zanclus cornutus

Rep. Kiribati []

P-CTX- []

Implicated in CFP investigations in Queensland, Australia (specific source location unconfirmed) []

Scomberomorus queenslandicus (School mackerel)

Moorish IdolC

CTX- [], CTX- [], CTX- [], P-CTX- [], P-CTX-B [, , ]

Hervey Bay, Queensland, Australia [, ], NSW, Australia [, ], eastern Australia [], implicated in CFP investigations in Queensland, Australia (specific source location unconfirmed) []

Scomberomorus commerson (Spanish mackerel)

LCMS/MS []

RBA []

(continued )

MQBA [], MBA [], RBA []

RBA []

LCMS/MS []

HPLC-MS/MS []

HPLC/MS [], TLC [], MBA [, , ], LCMS/MS [, , ], HPLC-MS/MS []

9 Gambierdiscus, the cause of ciguatera fish poisoning

341

CTX positive [, ] P-CTX-, P-CTX- []

CTX positive [] CTX positive [, , ]

CTX positive []

P-CTX-, P-CTX-, P-CTX- []

French Polynesia [, ]

French Polynesia [], southern coast of China [], Rep. Kiribati []

French Polynesia []

French Polynesia [, , ]

French Polynesia []

French Polynesia, Tubuai (Australes) [], southern coast of China [], Rep. Kiribati []

Chlorurus microrhinos (Steephead parrotfish) Hipposcarus longiceps (Pacific longnose parrotfish)

Leptoscarus vaigiensis (Marbled parrotfish)

Scarus altipinnis (Filament-finned parrotfish)

Scarus forsteni (Whitespot parrotfish)

Scarus ghobban (Blue-barred parrotfish)

Scarus jonesi

French Polynesia []

CTX positive []

CTX-A []

CTX positive [, ]

French Polynesia [, ]

Chlorurus frontalis (Pacific slopehead parrotfish)

Scarus gibbus French Polynesia [], Tahiti [], (Heavy beak parrotfish) French Polynesia []

CTX (if detected)

Source

Latin name (Common name)

Table 9.4 (continued )

Cat BA [], MQBA [], MBA []

HPLC/HNMR [], MQBA, MBA [, , ]

RBA [], MBA [], Cigua-check® [], LCMS/MS []

RBA []

RBA [, , ], NA []

RBA []

RBA [], MBA [], Cigua-check® [], LCMS/MS []

RBA [, ]

RBA [, ]

Method of detection

342 Hazel Farrell, Gurjeet S. Kohli and Shauna A. Murray

CTX positive [, ] P-CTX-, P-CTX-, P-CTX- []

CTX positive []

French Polynesia [, ]

Rep. Kiribati []

French Polynesia []

Scarus rubroviolaceus (Ember parrotfish)

Scarus russelii (Eclipse parrotfish)

Scarus schlegeli (Schlegel’s parrotfish)

P-CTX- []

Rep. Kiribati []

Siganus argenteus (Forktail rabbitfish)

Rep. Kiribati []

P-CTX- []

CTX positive []

Southern coast of China []

Fugu obscurus (Takifugu obscurus)

RabbitfishH

P-CTX-, P-CTX-, P-CTX- []

Rep. Kiribati []

Arothron nigropunctatus (Black-spotted puffer)

Puffer fishO

Diodon liturosus (Black blotched porcupinefish)

P-CTX- []

Diodon hystrix Rep. Kiribati [] (Spotfin porcupinefish)

PorcupinefishC

CTX positive []

French Polynesia []

Scarus psittacus (Palenose parrotfish)

LCMS/MS []

Cigua-check®, MBA []

LCMS/MS []

LCMS/MS []

LCMS/MS []

RBA []

LCMS/MS []

RBA [, ]

RBA []

(continued )

9 Gambierdiscus, the cause of ciguatera fish poisoning

343

CTX positive []

Eastern Mediterranean []

Siganus rivulatus (Marbled spinefoot)

SharkC

Monachus schauinslandi (Hawaiian monk seal)

SealC

Holothuria spp.

Sea cucumberH

Kyphosus cinerascens (Blue sea chub)

Hawaii []

Hawaii [, ]

French Polynesia [, , ]

P-CTX-C []

CTX positive [, ]

CTX positive [, , ]

CTX positive []

Southern coast of China []

Scatophagus argus (Spotted scat)

Sea chubO

CTX positive [, ]

Chile [, ]

Farmed salmon ScatO

SalmonO

CTX (if detected)

Source

Latin name (Common name)

Table 9.4 (continued )

LCMS/MS, NA []

Ciguatect® [, ]

RBA [, ], NA []

Cigua-check®, MBA []

SPIA [, ]

Cigua-check® []

Method of detection

344 Hazel Farrell, Gurjeet S. Kohli and Shauna A. Murray

CTX positive []

CTX positive []

Hawaii []

Hawaii []

Carcharhinus menisorrah (Grey sand shark)

Triaenodon obesus (White tip shark)

CTX positive []

CTX positive [], P-CTX-B []

C-CTX- []

P-CTX-, P-CTX-, P-CTX- []

Hawaii []

French Polynesia [], Capel Bank Seamount []

United States Virgin Islands waters []

Iimported to Hong Kong (source unconfirmed) [, ]

Aphareus furca (Black forktail snapper (wahanui))

Aprion virescens (Green jobfish, Blue green snapper)

Lutjanus apodus (Schoolmaster snapper) and Ocyurus chrysurus (Yellowtail snapper) hybrid

Lutjanus argentimaculatus (Mangrove red snapper)

SnapperC

I-CTX- & , I-CTX- & , I-CTX-, I-CTX-, gambieric acid D [Diogène et al., ]

Madagascar [Diogène et al., ]

Carcharhinus leucas (Bull shark)

(continued )

MBA [, ], LCMS/MS []

LCMS/MS, NA []

Cat BA [], MQBA [], MBA [], LCMS/MS []

S-EIA, SPIA []

MGBA []

MGBA []

MBA, NA, LC-ESI-HRMS [Diogène et al., ]

9 Gambierdiscus, the cause of ciguatera fish poisoning

345

TLC, MBA [, ], NA []

CTX positive [, ] C-CTX- []

CTX positive []

P-CTX-, P-CTX- []

St. Croix, US Virgin Islands [], French West Indies []

Fuerteventura, Spain []

Southern coast of China []

French Polynesia [, ], Rep. Kiribati []

Lutjanus buccanella (Blackfin snapper)

Lutjanus cyanopterus (Cubera snapper)

Lutjanus fulviflamma (Dory snapper)

Lutjanus fulvus (Blacktail snapper)

RBA [, ] NA [], LCMS/MS []

Cigua-check®, MBA []

LCMS/MS []

HPLC/MS [, ], Cat BA [], MGBA [, ], MQBA [], MBA [, –], Rajeish et al., , ], RBA [, ], LC/MS [] LCMS/MS [, , ], NA [], HPLC-MS /MS []

I-CTX- [, ], CTX-B [, ], deoxyCTX-B [], -epi--deoxyCTX -B [], P-CTX-B [], P-CTX-, P-CTX -, P-CTX- [, ], -hydroxy CTXC [], P-CTX- []

Republic of Mauritius [, ], Minamitorishima (Marcus) Island, Japan [, ], Hawaii [], French Polynesia [, , ], Vietnam [, ], Sri Lankan fishing grounds [], India [Rajeish et al., , , ], Taiwan [–], Rep. Kiribati [], implicated in CFP investigations in Queensland, Australia (specific source location unconfirmed) [], imported to Hong Kong (source unconfirmed) []

Lutjanus bohar (Two spot red snapper (Red bass, Red snapper))

Method of detection

CTX (if detected)

Source

Latin name (Common name)

Table 9.4 (continued )

346 Hazel Farrell, Gurjeet S. Kohli and Shauna A. Murray

RBA [, ], MBA [, , ], LC/MS [] LCMS/MS [] HPLC/MS [, ], HPLC/MS/RLB [, ], MGBA [, ], MBA [, ]

CTX positive []

CTX positive []

P-CTX-, P-CTX-, P-CTX- []

CTX-B [, ], -epi-deoxy CTXB, -Deoxy CTXB [] I-CTX [, ], I-CTX- [, ], I-CTX [, ], I-CTX- [, ] CTX-B [], C-CTX- []

P-CTX-, P-CTX-, P-CTX- []

French West Indies []

Hawaii []

Imported to Hong Kong (source unconfirmed) []

French Polynesia, [, ], Japan [, , ]

Republic of Mauritius (Nazareth, Saya de Malha, Soudan) [, ]

Antigua [], Okinawa, Japan [], West Africa [], Baja California, Mexico [], St. Thomas, Caribbean Sea [], French West Indies [], Indonesia [], India []

Imported to Hong Kong (source unconfirmed) []

Lutjanus jocu (Dog snapper)

Lutjanus kasmira (Bluestripe snapper (taape))

Lutjanus malbaricus (Malabar blood snapper)

Lutjanus monostigma (One-spot snapper)

Lutjanus sebae (Red emperor)

Lutjanus spp. (Snapper)

Lutjanus stellatus (Star snapper)

(continued )

MBA [, ], LCMS/MS []

HPLC/MS [], BSBA [], MGBA [], MBA [, , ], S-EIA [], SPIA [], NA [, , ], LCMS/MS [, ]

MBA, LCMS/MS []

MBA, S-EIA, SPIA []

MBA, NA []

HPLC/MS, MBA [, ]

C-CTX- and isomers [], CTX congeners []

French West Indies [, ]

Lutjanus griseus (Grey snapper)

MQBA [], MBA [], RBA []

CTX positive [, ]

French Polynesia [, ]

Lutjanus gibbus (Humpback red snapper)

9 Gambierdiscus, the cause of ciguatera fish poisoning

347

CTX positive [] P-CTX- []

CTX positive []

Rep. Kiribati []

Hawaii []

Myripristis berndti (Bigscale soldierfish

Myripristis kuntee (Epaulette Soldierfish (squirrelfish))

P-CTX-B, -OH-P-CTX-C, P-CTX-C, P-CTX-A, P-CTX-B. []

Southern coast of China []

French Polyneisa [, ]

Dispinus ruber

Squirrelfish and SoldierfishC

Tripneustes gratilla (Collector urchin)

Sea UrchinH

Pagrus pagrus (Red porgy)

C-CTX- []

C-CTX- []

US Virgin Islands []

Selvagen Islands (Maderia, Portugal) []

P-CTX- []

Rep. Kiribati []

Macolor niger (Black and white snapper) Ocyurus chrysurus (Yellowtail snapper)

Seabream (Porgies)C

CTX (if detected)

Source

Latin name (Common name)

Table 9.4 (continued )

MBA, S-EIA, SPIA []

LCMS/MS []

Cigua-check®, MBA []

RBA [] NA [, ], LCMS/MS []

LCMS/MS []

NA, LCMS/MS []

LCMS/MS []

Method of detection

348 Hazel Farrell, Gurjeet S. Kohli and Shauna A. Murray

Nuku Hiva (Marquesas) [] southern coast of China []

Hawaii []

Ophiocoma spp. (Ophiuroids (Brittle stars))

P-CTX-, P-CTX-, P-CTX- []

P-CTX- []

Rep. Kiribati []

Rep. Kiribati []

Acanthurus glaucopareius (Chocolate surgeonfish)

CTX positive []

Acanthurus gahhm (Black surgeonfish)

Acanthurus dussumieri Hawaii [] (Dussumier’s surgeonfish (palani))

SurgeonfishH, O

Three CTX analogues reported as CTX-C equivalents []

Madeira and Azores archipelagos []

Ophidiaster ophidianus (Red starfish) CTX positive []

Three CTX analogues reported as CTX-C equivalents []

P-CTX- []

CTX positive [, ]

Marthasterias glacialis Madeira and Azores archipelagos (Spiny starfish) []

StarfishO

Sargocentron tiere Rep. Kiribati [] (Bluelined squirrelfish)

Sargocentron spiniferum (Sabre squirrelfish)

LCMS/MS []

LCMS/MS []

MBA, S-EIA []

Ciguatect® []

UPLC/MS []

UPLC/MS []

LCMS/MS []

(continued )

RBA [], MBA [], Cigua-check® []

9 Gambierdiscus, the cause of ciguatera fish poisoning

349

Acanthurus Nuku Hiva (Marquesas) [], Rep. Kiribati xanthopterus [] (Yellowfin surgeonfish)

Hawaii []

P-CTX-, P-CTX-, P-CTX- []

CTX positive []

CTX positive []

Hawaii []

Acanthurus sp.

CTX positive []

Hawaii []

Acanthurus nigroris (Bluelined surgeonfish (maiko)) Acanthurus olivaceus (Orangeband surgeonfish (naenae))

RBA [], LCMS/MS []

S-EIA [], SPIA []

MBA, S-EIA []

MBA, S-EIA []

LCMS/MS []

LCMS/MS []

P-CTX- []

P-CTX-, P-CTX-, P-CTX- []

LCMS/MS []

P-CTX-, P-CTX-, P-CTX- []

Acanthurus lineatus Rep. Kiribati [] (Striped surgeonfish) Acanthurus maculiceps Rep. Kiribati [] (White freckled surgeonfish)

Rep. Kiribati []

RBA []

CTX – positive []

Acanthurus French Polynesia [] leucopareius (Whitebar surgeonfish)

Acanthurus mata (White freckled surgeonfish)

Method of detection

Source

CTX (if detected)

Latin name (Common name)

Table 9.4 (continued )

350 Hazel Farrell, Gurjeet S. Kohli and Shauna A. Murray

French Polynesia [, ], Tahiti [], Rep. Kiribati []

UnicornfishH, O

Gymnosarda unicolor (Dogtooth tuna)

CTX positive []

P-CTX-, P-CTX-, P-CTX- []

Rep. Kiribati []

Balistapus undulatus (Orange stripe triggerfish)

French Polynesia [, ]

C-CTX- []

US Virgin Islands []

Balistes vetula (Queen triggerfish)

TunaC

C-CTX- []

CTX positive []

Selvagens Islands (Madeira, Portugal) []

St. Barthelemy, Caribbean Sea []

CTX positive []

P-CTX-, P-CTX-, P-CTX- []

Balistes capriscus (Queen triggerfish)

TriggerfishC

Malacanthus plumieri (Sand tilefish)

TilefishC

Nemipterus Southern coast of China [] nematophorus (Double-whip threadfin bream)

Threadfin breamC

Ctenochaetus striatus (Striped Bristletooth)

(continued )

Cat BA [], MQBA [], MBA [], RBA []

LCMS/MS []

NA, LCMS/MS []

LCMS/MS []

TLC, MBA []

Cigua-check®, MBA []

RBA [, ], LCMS/MS []

9 Gambierdiscus, the cause of ciguatera fish poisoning

351

CTX positive [, ] CTX positive [] CTX positive [, ]

French Polynesia [, ]

Nuku Hiva (Marquesas) []

Nuku Hiva (Marquesas) [, ]

Naso brevirostris (Spotted unicornfish)

Naso hexacanthus (Sleek unicornfish) Naso lituratus (Orangespine unicornfish)

P-CTX-, P-CTX-, P-CTX- []

P-CTX-, P-CTX- []

French Polynesia [], Rep. Kiribati []

Coris aygula (Clown coris)

CTX positive []

French Polynesia [], Imported to Hong Kong (source unconfirmed) []

Canary Islands []

Cheilinus undulatus (Humphead wrasse)

WrasseC

Acanthocybium solandri

WahooC

CTX positive [, ]

CTX positive []

Nuku Hiva (Marquesas) []

Naso brachycentron (Humpback unicornfish)

Naso unicornis French Polynesia [, ] (Bluespine unicornfish)

CTX (if detected)

Source

Latin name (Common name)

Table 9.4 (continued )

RBA [, ], LCMS/MS []

Cat BA [], MQBA [], MBA [, , ], LCMS/MS []

NA []

RBA [, ]

Cat BA [], MQBA [], MBA [], RBA []

RBA []

RBA [, ]

RBA []

Method of detection

352 Hazel Farrell, Gurjeet S. Kohli and Shauna A. Murray

Baja California, Mexico []

O: omnivore.

Semicossyphus sp.

Typical feeding behavior: C: carnivore, H: herbivore,

CTX positive []

P-CTX- []

MBA []

LCMS/MS []

The abbreviations are: LC/MS, liquid chromatography mass spectrometry; LCMS/MS, liquid chromatography tandem mass spectrometry; UPLC/MS, ultra-performance liquid chromatography/mass spectrometry; HPLC/MS, high-performance liquid chromatography/mass spectrometry; HPLC-MS/MS, high-performance liquid chromatography tandem mass spectrometry; HPLC/HNMR, high-performance liquid chromatography/H nuclear magnetic resonance; HPLC/MS/RLB, highperformance liquid chromatography/mass spectrometry/radio ligand binding; LC-ESI-HRMS, liquid chromatography electrospray ionization highresolution mass spectrometry; TLC, thin-layer chromatography; BSBA, brine shrimp bioassay; DLBA, diptera larvae bioassay; MGBA, mongoose bioassay; MQBA, mosquito bioassay; MBA, mouse bioassay; SEIA, stick enzyme immunoassay; SPIA, solid-phase immunoassay; RIA, radioimmunoassay; ELISA, enzyme-linked immunosorbent assay; N2A, neuroblastoma cytotoxicity assays; RBA, receptor-binding assay; MA, membrane assay; BA, bioassay.

Rep. Kiribati []

Epibulus insidiator (Slingjaw wrasse)

9 Gambierdiscus, the cause of ciguatera fish poisoning

353

354

Hazel Farrell, Gurjeet S. Kohli and Shauna A. Murray

The CTX-positive reports in Table 9.4 are predominantly concerned with the midlatitude tropical and subtropical zones. This is fitting with the distribution of Gambierdiscus as described in Table 9.2. However, CFP has also been reported in non-endemic areas because of an increase in seafood imports [57, 152]. While the majority of studies have focused on reef fish, toxin accumulation has been observed in eels, sea cucumbers, starfish, seals and jellyfish (see Table 9.4 and references therein). Sharks have also been suspected of causing CFP following outbreaks of human illness, remnant samples for testing were unavailable [108, 239]. Further studies are required to address the deficit in information for species other than fish and to identify potential toxin vectors in coastal systems. For the most part, CFP studies have focused on CTX rather than MTX. The MBA has been used previously to test for MTX, with positive results in Ctenochaetus striatus (striped bristletooth) [112]. A gap in our existing knowledge is whether the presence of MTX in small (herbivorous) fish species is transferred up the food chain to larger carnivorous species. Often, in small island nations, local fishing communities are aware of ciguateraprone zones and avoid certain fish species. Such knowledge certainly has its merits; however, a study by Darius et al. [87] in French Polynesia demonstrated the presence of CTXs in fish species that were considered safe to eat by locals. Experimentally, CTX toxin profiles and structures have been determined by chromatographic techniques (HPLC, UPLC and LC-MS), accompanied by nuclear magnetic resonance [95, 96, 136, 187] and radio ligand binding [105, 106]. However, these methods are not commonplace or practical for routine testing, as they are costly and require special expertise. Confirmation of toxin by UPLC/HPLC followed by LC-MS involves the isolation and fractionation of the various CTX compounds and their known molecular weights (see Table 9.3). Although a rapid method for sample analysis has been proposed [240], acquiring purified CTX standards is problematic due to the limited supply of natural CTX compounds [82]; though artificial synthesis of CTX is possible [241], it is highly complex. Without a consistent source of reference material, absolute quantification of CTXs and their congeners is hard to achieve. In addition, technical issues such as coeluting peaks of similar compounds and inhibiting/promoting matrix effects remain unresolved. Several biological assays have been developed for the detection of ciguateric fish. These include the use of chickens [175], cats [213], mongooses [161], diptera larva [188], brine shrimp [167] and mosquitos [176]. However, each assay has its own constraints and limitations, largely relating to toxin specificity and quantification, but also due to inefficiencies and ethical considerations (summarized in de Fouw, 2001 [147]). While the MBA by intraperitoneal injection does not provide a linear dose-response relationship with CTX toxicity [234], it remains the most widely used biological assay (see Table 9.4). Numerous biochemical assays have been proposed as alternatives to biological assays for testing seafood. The development of a radioimmunoassay [161] progressed to a cheaper alternative enzyme-linked immunosorbent assay (ELISA) with higher throughput [162]. The ELISA test has recently shown promising correlations

9 Gambierdiscus, the cause of ciguatera fish poisoning

355

with biological assays [163, 172]. Stick enzyme immunoassay (SEIA) [164] and solid phase immunoassay (SPIA) [179] tests have led to the development of commercial kits (i.e., Cigua-check® and Ciguatect®). However, these products have yielded a large number of false-positive and false-negative results [200] and the Cigua-check® test is no longer being manufactured. Other assays utilized for screening CTXs in fish are the sodium channel binding assay (N2A) [93] and RBA [87, 165]. Both of these assays have shown promising results, however, these assays cannot quantify specific congeners of CTXs and MTXs. This can only be achieved via further development and validation via LC-MS analysis, and there is an urgent need to do so. The progress has been disadvantaged by the lack of available purified standards [242]. Other challenges are the presence of more than one type of CTXs, (see Table 9.3) being present in fish specimens [101, 216].

9.7 Conclusion Major advances have been made in the study of Gambierdiscus and Fukuyoa species, and their contributions to CFP. Concurrently, new questions and challenges have also been raised. Here, we outline the major areas that require research efforts to significantly advance our understanding of the causes of the production of toxins leading to CFP: 1. It is highly likely that species of Gambierdiscus vary in their toxicity, whereas intraspecific toxin production appears to be more consistent. Exact Gambierdiscus and Fukuyoa species identifications in CFP-affected areas around the world are therefore required. New molecular and taxonomic tools to identify Gambierdiscus and Fukuyoa species accurately and simply will therefore be required. 2. Toxin profiles of Gambierdiscus and Fukuyoa strains and species are needed to identify the exact CTXs and MTXs produced. 3. The development and standardization of chromatographic techniques to accurately quantify different CTXs and MTXs are required. This involves a further characterization of already known and new congeners of CTXs and MTXs. 4. The development of commercially available CTX and MTX standards is very important, as it is one of the major hurdles that prevent further advancement of areas mentioned in the earlier two goals. 5. The elucidation of genes involved in biosynthesis of CTX and MTX in Gambierdiscus and Fukuyoa species will allow for an increased understanding of the causes and triggers of toxin production and the potential for the development of novel CFP monitoring tools.

356

Hazel Farrell, Gurjeet S. Kohli and Shauna A. Murray

References [1]

Simon N, Cras AL, Foulon E, Lem ER. Diversity and evolution of marine phytoplankton. Comptes rendus. Biologies 2009;332:159–170. [2] Murray SA, Patterson DJ, Thesson A. Transcriptomics and microbial eukaryotic diversity: a way forward. Trends Ecol Evol 2012;27:651–652. [3] Taylor F, Hoppenrath M, Saldarriaga JF. Dinoflagellate diversity and distribution. Biodivers Conserv 2008;17:407–418. [4] Hoppenrath M, Murray SA, Chomérat N, Horiguchi T. Marine Benthic DinoflagellatesUnveiling Their Worldwide Biodiversity. Stuttgart; E. Schweizerbart’sche Verlagsbuchhandlung: 2014. [5] Yasumoto T, Nakajima I, Bagnis R, Adachi R. Finding of a dinoflagellate as a likely culprit of ciguatera. Bull Jap Soc Sci Fish 1977;43:1021–1026. [6] Adachi R, Fukuyo Y. The thecal structure of a toxic marine dinoflagellate Gambierdiscus toxicus gen. et spec. nov. collected in a ciguatera-endemic area. Bull Jap Soc Sci Fish 1979;45:67–71. [7] Chinain M, Germain M, Sako Y, Pauillac S, Legrand A-M. Intraspecific variation in the dinoflagellate Gambierdiscus toxicus (Dinophyceae) isozyme analysis. J Phycol 1997;33:36–43. [8] Holmes MJ. Gambierdiscus yasumotoi sp. nov. (Dinophyceae), a toxic benthic dinoflagellate from southeastern asia. J Phycol 1998;34:661–668. [9] Chinain M, Faust MA, Pauillac S. Morphology and molecular analyses of three toxic species of Gambierdiscus (Dinophyceae): G. pacificus, sp. nov., G. australes, sp. nov., and G. polynesiensis, sp. nov. J Phycol 1999;35:1282–1296. [10] Chinain M, Darius HT, Ung A, et al. Growth and toxin production in the ciguatera-causing dinoflagellate Gambierdiscus polynesiensis (Dinophyceae) in culture. Toxicon 2010;56: 739–750. [11] Rhodes LL, Smith KF, Munday R, et al. Toxic dinoflagellates (Dinophyceae) from Rarotonga, Cook Islands. Toxicon 2010;56:751–758. [12] Fraga S, Rodriguez F, Caillaud A, Diogene J, Raho N, Zapata M. Gambierdiscus excentricus sp. nov. (Dinophyceae), a benthic toxic dinoflagellate from the Canary Islands (NE Atlantic Ocean). Harmful Algae 2011;11:10–22. [13] Holland WC, Litaker RW, Tomas CR, et al. Differences in the toxicity of six Gambierdiscus (Dinophyceae) species measured using an in vitro human erythrocyte lysis assay. Toxicon 2013;65:15–33. [14] Murray JS, Selwood AI, Harwood DT, van Ginkel R, Puddick J, Rhodes LL, . . . Wilkins AL. 44-Methylgambierone, a new gambierone analogue isolated from Gambierdiscus australes. Tetrahedron Lett 2019;60(8):621–625. [15] Gomez F, Qiu D, Lopes RM, Lin S. Fukuyoa paulensis gen. et sp. nov., a new genus for the globular species of the dinoflagellate Gambierdiscus (Dinophyceae). PLoS One 2015;10: e0119676. [16] Litaker RW, Holland WC, Hardison DR, Pisapia F, Hess P, Kibler SR, Tester PA. Ciguatoxicity of Gambierdiscus and Fukuyoa species from the Caribbean and Gulf of Mexico. PLoS One 2017;12(10):e0185776. [17] Friedman MA, Fleming LE, Fernandez M, et al. Ciguatera fish poisoning: treatment, prevention and management. Mar Drugs 2008;6:456–479. [18] Fleming LE, Baden DG, Bean JA, Weisman R, Blythe DG. Seafood toxin diseases: issues in epidemiology and community outreach. In: Reguera B, Blanco J, Fernandez ML, Wyatt T [eds.]. Harmful Algae; Xunta de Galicia and Intergovernmental Oceanographic Commission of UNESCO: 1998. 245–248.

9 Gambierdiscus, the cause of ciguatera fish poisoning

[19] [20] [21] [22] [23] [24] [25] [26] [27]

[28]

[29]

[30]

[31]

[32]

[33]

[34] [35]

[36]

[37]

357

Bagnis R, Kuberski T, Laugier S. Clinical observations on 3,009 cases of ciguatera (fish poisoning) in the South Pacific. Am J trop Med Hygine 1979;28:1067. Gillespie NC. Possible origins of ciguatera. In: Covacevich J, Davie P, Pearn J [eds.] Toxic Plants and Animals, a Guide for Australia. Brisbane; Queensland Museum: 1987. 171–179. Sims JK. A theoretical discourse on the pharmacology of toxic marine ingestions. Ann Emergency Med 1987;16:1006. Glibert PM, Anderson DM, Gentien P, Graneli E, Sellner KG. The global, complex phenomena of harmful algal blooms. Oceanography 2005;18(2):136–147. Hallegraeff GM. Ocean climate change, phytoplankton community responses, and harmful algal blooms: a formidable predictive challenge. J Phycol 2010;46:220–235. Skinner MP, Brewer TD, Johnstone R, Fleming LE, Lewis RJ. Ciguatera fish poisoning in the Pacific islands. PLoS Negl Trop Dis 2011;5:e1416. Bomber JW, Rubio MG, Norris DR. Epiphytism of dinoflagellates associated with the disease ciguatera: substrate specificity and nutrition. Phycologia 1989;28:360–368. Bomber JW, Tindall DR, Miller DM. Genetic variability in toxin potencies among seventeen clones of Gambierdiscus toxicus (Dinophyceae). J Phycol 1989;25:615–625. Holmes MJ, Lewis RJ, Gillespie NC. Toxicity of Australian and French Polynesian strains of Gambierdiscus toxicus (Dinophyceae) grown in culture: Characterization of a new type of maitotoxin. Toxicon 1990;28:1159–1172. Holmes MJ, Lewis RJ, Poli MA, Gillespie NC. Strain dependent production of ciguatoxin precursors (gambiertoxins) by Gambierdiscus toxicus (Dinophyceae) in culture. Toxicon 1991;29:761–775. Richlen ML, Morton SL, Barber PH, Lobel PS. Phylogeography, morphological variation and taxonomy of the toxic dinoflagellate Gambierdiscus toxicus (Dinophyceae). Harmful Algae 2008;7:614–629. Morton SL, Bomber JW, Tindall DR, Aikman KE. Response of Gambieridscus toxicus to light: cell physiology and toxicity. In: Smayda TJ, Shimizu Y [eds.]. Toxic Phytoplankton Blooms in the sea. New York; Elsevier: 1993. 541–546. Bomber JW, Morton SL, Babinchak JA, Norris DR, Morton JG. Epiphytic dinoflagellates of drift algae another toxigenic community in the ciguatera food chain. Bull Mar Sci 1988;43: 204–214. Litaker RW, Vandersea MW, Faust MA, et al. Taxonomy of Gambierdiscus including four new species, Gambierdiscus caribaeus, Gambierdiscus carolinianus, Gambierdiscus carpenteri and Gambierdiscus ruetzleri (Gonyaulacales, Dinophyceae). Phycologia 2009;48:344–390. Fraga S, Rodríguez F. Genus Gambierdiscus in the Canary Islands (NE Atlantic Ocean) with description of Gambierdiscus silvae sp. nov., a new potentially toxic epiphytic benthic dinoflagellate. Protist 2014;165(6):839–853. Faust MA. Observation of sand-dwelling toxic Dinoflagellates (Dinophyceae) from widely differing sites, including two new species. J Phycol 1995;31:996–1003. Nishimura T, Sato S, Tawong W, Sakanari H, Yamaguchi H, Adachi M. Morphology of Gambierdiscus scabrosus sp. nov. (Gonyaulacales): a new epiphytic toxic dinoflagellate from coastal areas of Japan. J Phycol 2014;50:506–514. Fraga S, Rodriguez F, Riobo P, Bravo I. Gambierdiscus balechii sp. nov (Dinophyceae), a new benthic toxic dinoflagellate from the Celebes Sea (SW Pacific Ocean). Harmful Algae 2016;58: 93–105. Smith KF, Rhodes L, Verma A, Curley BG, Harwood DT, Kohli GS, et al. A new Gambierdiscus species (Dinophyceae) from Rarotonga, Cook Islands: Gambierdiscus cheloniae sp nov. Harmful Algae 2016;60:45–56.

358

Hazel Farrell, Gurjeet S. Kohli and Shauna A. Murray

[38] Kretzschmar AL, Verma A, Harwood T, Hoppenrath M, Murray S. Characterization of Gambierdiscus lapillus sp. nov. (Gonyaulacales, Dinophyceae): a new toxic dinoflagellate from the Great Barrier Reef (Australia). J Phycol 2017;53:283–297. [39] Rhodes L, Smith KF, Verma A, Curley BG, Harwood DT, Murray S, et al. A new species of Gambierdiscus (Dinophyceae) from the south-west Pacific: Gambierdiscus honu sp nov. Harmful Algae 2017;65:61–70. [40] Jang SH, Jeong HJ, Yoo YD. Gambierdiscus jejuensis sp. nov., an epiphytic dinoflagellate from the waters of Jeju Island, Korea, effect of temperature on the growth, and its global distribution. Harmful Algae 2018;80:149–157. [41] Kretzschmar A, Larsson ME, Hoppenrath M, Doblin MA, Murray SA. Characterisation of Two Toxic Gambierdiscus spp. (Gonyaulacales, Dinophyceae) from the Great Barrier Reef (Australia): G. lewisii sp. nov. and G. holmesii sp. nov. Protist 2019; 170 (6): 125699. [42] Dai X, Mak YL, Lu CK, Mei HH, Wu JJ, Lee WH, et al. Taxonomic assignment of the benthic toxigenic dinoflagellate Gambierdiscus sp. type 6 as Gambierdiscus balechii (Dinophyceae), including its distribution and ciguatoxicity. Harmful Algae 2017;67:107–118. [43] Kuno S, Kamikawa R, Yoshimatsu S, Sagara T, Nishio S, Sako Y. Genetic diversity of Gambierdiscus spp. (Gonyaulacales, Dinophyceae) in Japanese coastal areas. Phycol Res 2010;58:44–52. [44] Nishimura T, Sato S, Tawong W, et al. Genetic diversity and distribution of the ciguateracausing dinoflagellate Gambierdiscus spp. (Dinophyceae) in coastal areas of Japan. PLoS ONE 2013;8:e60882. [45] Litaker RW, Vandersea MW, Faust MA, et al. Global distribution of ciguatera causing dinoflagellates in the genus Gambierdiscus. Toxicon 2010;56:711–730. [46] Xu Y, Richlen ML, Morton SL, Mak YL, Chan LL, Tekiau A, Anderson DM. Distribution, abundance and diversity of Gambierdiscus spp. from a ciguatera-endemic area in Marakei, Republic of Kiribati. Harmful Algae 2014;34:56–68. [47] Hernández-Becerril DU, Almazán Becerril A. Especies de dinoflagelados del género Gambierdiscus (Dinophyceae) del Mar Caribe mexicano. Rev Biol Trop 2004;52:77–87. [48] Murray, S., Momigliano, P., Heimann, K. and Blair, D. 2014. Molecular phylogenetics and morphology of Gambierdiscus yasumotoi from tropical eastern Australia. Harmful Algae 39: 242–52. [49] The HV. Sinh Hoc Tao Hai Roi Co Vo Song Day Vung Bien Ven Bo Viet Nam. Institute of Oceano- graphy, Vinh Nguyen, Nha Trang, Nha Trang, 2009. [50] Rhodes L, Harwood T, Smith K, Argyle P, Munday R. Production of ciguatoxin and maitotoxin by strains of Gambierdiscus australes, G. pacificus and G. polynesiensis (Dinophyceae) isolated from Rarotonga, Cook Islands. Harmful Algae 2014;39:185–190. [51] Rhodes L, Harwood T, Smith K, Argyle P, Munday R. Corrigendum to ‘Production of ciguatoxin and maitotoxin by strains of Gambierdiscus australes, G. pacificus and G. polynesiensis (Dinophyceae) isolated from Rarotonga, Cook Islands’ [Harmful Algae 39 (2014) 185–190]. Harmful Algae 2016;55:295. [52] Munir S, Siddiqui PJA, Morton SL. The occurrence of the ciguatera fish poisioning producing dinoflagellate genus Gambieridiscis in Pakistan waters. Algae 2011;26:317–325. [53] Rodríguez, F., Fraga, S., Ramilo, I., Rial, P., Figueroa, R.I., Riobó, P. and Bravo, I., 2017. Canary Islands (NE Atlantic) as a biodiversity ‘hotspot’of Gambierdiscus: Implications for future trends of ciguatera in the area. Harmful algae, 67, pp.131–143. [54] Lewis RJ, Inserra M, Vetter I, Holland WC, Hardison DR, et al. (2016) Rapid Extraction and Identification of Maitotoxin and Ciguatoxin-Like Toxins from Caribbean and Pacific Gambierdiscus Using a New Functional Bioassay. PLOS ONE 11(7): e0160006. https://doi. org/10.1371/journal.pone.0160006.

9 Gambierdiscus, the cause of ciguatera fish poisoning

[55] [56]

[57] [58]

[59]

[60]

[61]

[63]

[64]

[65]

[66] [67] [68]

[69]

[70] [71]

[72]

359

Nascimento SM, Melo G, Salgueiro F, Diniz BS, Fraga S (2019). Morphology of Gambierdiscus excentricus (Dinophyceae) with emphasis on sulcal plates. Phycologia 54: 628–639. Hoppenrath, M., Kretzschmar, A.L., Kaufmann, M.J. et al. Morphological and molecular phylogenetic identification and record verification of Gambierdiscus excentricus (Dinophyceae) from Madeira Island (NE Atlantic Ocean). Mar Biodivers Rec 12, 16 (2019) doi:10.1186/s41200-019-0175-4. Glaziou P, Legrand AM. The epidemiology of ciguatera fish poisoning. Toxicon 1994;32: 863–873. Mohammad-Noor N, Daugbjerg N, Moestrup Ø, Anton A. Marine epibenthic dinoflagellates from Malaysia – a study of live cultures and preserved samples based on light and scanning electron microscopy. Nord J Bot;24. Leaw C-P, Lim P-T, Tan T-H, et al. First report of the benthic dinoflagellate, Gambierdiscus belizeanus (Gonyaulacales: Dinophyceae) for the east coast of Sabah, Malaysian Borneo. Phycol Res 2011;59:143–146. Catania D, Richlen ML, Mak YL, Morton SL, Laban EH, Xu Y et al (2017). The prevalence of benthic dinoflagellates associated with ciguatera fish poisoning in the central Red Sea. Harmful Algae 68: 206–216. Kohli GS, John U, Figueroa RI, Rhodes LL, Harwood DT, Groth M et al (2015). Polyketide synthesis genes associated with toxin production in two species of Gambierdiscus (Dinophyceae). BMC Genomics 16: 410. Jeong HJ, Lim AS, Jang SH, et al. First report of the epiphytic dinoflagellate Gambierdiscus caribaeus in the temperate waters off Jeju Island, Korea: morphology and molecular characterization. J Eukaryot Microbiol 2012;59:637–650. Chinain M, Darius HT, Ung A, Fouc MT, Revel T, Cruchet P, . . . Laurent D. Ciguatera risk management in French Polynesia: the case study of Raivavae Island (Australes Archipelago). Toxicon 2010;56(5):674–690. Kohli GS, Murray SA, Neilan BA, Rhoden LL, Harwood DT, Smith KF, Meyer L, Capper A, Bret S, Hallegraef GM. High abundance of the potentally maitotoxic dinofagellate Gambierdiscus carpenteri in temperate waters of New South Wales, Auntralia. Harmful Algae 2014;39 (0):134–145. Roeder K, Erler K, Kibler S, et al. Characteristic profiles of Ciguatera toxins in different strains of Gambierdiscus spp. Toxicon 2010;56:731–738. Lu SH, Hodgkiss IJ. Harmful algal bloom causative collected from Hong Kong waters. Hydrobiologia 2004;512:231–238. Praseno DP, Wiadnyana NN. HAB organisms in Indonesian waters. In: Penney RW [eds.] Proceedings of the 5th Canadian Workshop on Harmful Marine Algae. Technical Report on Fish and Aq Sci: 1996. 69–75. Hurbungs MD, Jayabalan N. In: Seasonal distribution of potentially toxic dinoflagellates in the lagoon of Trou Aux Biches, Mauritius, Proceedings of the Fifth Annual meeting of Agricultural Scientists. Lalouette DYB JA, Ed. The Food and Agricultural Research Council: 2002. p. 211–217. Ceballos-Corona JGA. Ańalisis de los dinoflagelados potencialmente nocivos en el fitoplancton de la zona neŕıtica de la Costa Michoacana. Morelia, Mexico; UMSNH: 2006. Grzebyk D, Berland B, Thomassin BA, Bosi C, Arnoux A. Ecology of ciguateric dinoflagellates in the coral-reef complex of Mayotte Island (SW Indian Ocean). J Exp Mar Biol Ecol 1994;178:51–66. Hernandez-Becerril DU, Alonso-Rodriguez R, Alvarez-Gongora C, et al. Toxic and harmful marine phytoplankton and microalgae (HABs) in Mexican Coasts. J Environ Sci Health Part A Toxic/Hazard Subst Environ Eng 2007;42:1349–1363.

360

Hazel Farrell, Gurjeet S. Kohli and Shauna A. Murray

[73] Habermehl GG, Krebs HC, Rasoanaivo P, Ramialiharisoa A. Severe ciguatera poisoning in Madagascar: A case report. Toxicon 1994;32:1539–1542. [74] de Silva E. Contribution a l’etude du microplacton de Daker et des regions maritimes voisines. Bull Institut Fondamental d’Afrique Noire 1956;18:13–14. [75] Taylor FJR. The description of the benthic dinoflagellate associated with maitotoxin and ciguatoxin, including observations on Hawaiian material. In: Taylor D, Seliger H [eds.] Toxic Dinoflagellates blooms. New York; Elsevier Scientific: 1979. 71–77. [76] Aligizaki K, Katikou P, Nikolaidis G In: Toxic benthic dinoflagellates spreading and potential risk in the Mediterranean Sea, 7th International Conference in Molluscan Shellfish Safety; 2009. [77] Aligizaki K, Battocchi C, Penna A, Rodríguez Hernández F, Arsenakis M, Fraga S. In: Diversity of potentially toxic benthic dinoflagellates in southern Europe, 14th International Conference on Harmful Algae. Crete; Intergovernmental Oceanographic Commission of UNESCO: 1–5 November, 2010. p. 25. [78] Lobel PS, Andersen D, Clement-Durand M. Assessment of ciguatera dinoflagellate populations: sample variability and algal substrate selection. Biol Bull 1988;175:94–101. [79] Delgado G, Lechuga-Devéze CH, Popowski G, Troccoli L, Salinas CA. Epiphytic dinoflagellates associated with ciguatera in the northwestern coast of Cuba. Rev Biol Trop 2006;54: 299–310. [80] Okolodkov YB, Campos-Bautista G, Gárate-Lizárraga I, González-González JAG, Hoppenrath M, Arenas V. Seasonal changes of benthic and epiphytic dinoflagellates in the Veracruz reef zone, Gulf of Mexico. Aquat Microb Ecol 2007;47:223–237. [81] Parsons ML, Aligizaki K, Bottein M-YD, et al. Gambierdiscus and Ostreopsis: Reassessment of the state of knowledge of their taxonomy, geography, ecophysiology, and toxicology. Harmful Algae 2012;14:107–129. [82] Berdalet E, Tester PA, Zingone A. Global ecology and oceanography of harmful algal blooms: HABs in Benthic Systems. Paris and Newark; GEOHAB: 2012. 66. [83] Bienfang P, Oben B, DeFelice S, et al. Ciguatera: the detection of neurotoxins in carnivorous reef fish from the coast of Cameroon, West Africa. Afr J Mar Sci 2008;30:533–540. [84] Chinain M, Germain M, Deparis X, Pauillac S, Legrand AM. Seasonal abundance and toxicity of the dinoflagellate Gambierdiscus spp. (Dinophyceae), the causative agent of ciguatera in Tahiti, French Polynesia. Mar Biol 1999;135:259–267. [85] Nakajima I, Oshima Y, Yasumoto T. Toxicity of benthic dinoflagellates in Okinawa. Bull Jap Soc Sci Fish 1981;47:1029–1033. [86] Withers NW In: Ciguatera research in the northwestern Hawaiian Islands: laboratory and field studies on ciguatoxigenic dinoflagellates in the Hawaiian Archipelago, Proceedings of the Second Symposium on Resource Investigations in the Northwestern Hawaiian Islands, University of Hawaii Sea Grant 1984. University of Hawaii Sea Grant: 1984. p. 144–156. [87] Darius H, Ponton D, Revel T, et al. Ciguatera risk assessment in two toxic sites of French Polynesia using the receptor-binding assay. Toxicon 2007;50:612–626. [88] Caillaud A, de la Iglesia P, Darius HT, et al. Update on methodologies available for ciguatoxin determination: perspectives to confront the onset of Ciguatera fish poisoning in Europe. Mar Drugs 2010;8:1838–1907. [89] Ballantine DL, Tosteson TR, Bardales AT. Population dynamics and toxicity of natural populations of benthic dinoflagellates in southwestern Puerto Rico. J Exp Mar Biol Ecol 1988;119:201–212. [90] Lewis RJ, Wong Hoy AW, McGiffin DC. Action of ciguatoxin on human atrial trabeculae. Toxicon 1992;30:907–914.

9 Gambierdiscus, the cause of ciguatera fish poisoning

[91]

[92]

[93]

[94]

[95]

[96]

[97] [98] [99] [100]

[101] [102] [103]

[104]

[105] [106] [107]

[108] [109]

361

Mattei C, Dechraoui MY, Molgo J, Meunier FA, Legrand AM, Benoit E. Neurotoxins targetting receptor site 5 of voltage-dependent sodium channels increase the nodal volume of myelinated axons. J Neurosci Res 1999;55:666–673. Lewis R, Molgo J, Adams DJ. Ciguatera toxins: Pharmacology of toxins involved in ciguatera and related fish poisonings. Seafood Freshwater Toxins: Pharmacol Physiol Detect 2000;103:419. Dickey R. Ciguatera toxins: chemistry, toxicology, and detection. In: Botana LM [ed.] Seafood and Freshwater Toxins: Pharmacology, Physiology and Detection (2nd edition). New York; CRC Press: 2008. 479–496. Legrand A-M, Teai T, Cruchet P, Satake M, Murata K, Yasumoto T. In: Two structural types of ciguatoxin involved in ciguatera fish poisoning in French Polynesia. 8th International conference on Harmful Algae 1998; Reguera B, Blanco J, Fernandez ML, Wyatt T, eds. Xunta de Galicia and IOC/UNESCO. p. 473–475. Murata M, Legrand AM, Ishibashi Y, Fukui M, Yasumoto T. Structures and configurations of ciguatoxin from the moray eel Gymnothorax javanicus and its likely precursor from the dinoflagellate Gambierdiscus toxicus. J Am Chem Soc 1990;112:4380–4386. Lewis RJ, Sellin M, Poli MA, Norton RS, MacLeod JK, Sheil MM. Purification and characterization of ciguatoxins from moray eel (Lycodontis javanicus, Muraenidae). Toxicon 1991;29:1115–1127. Lewis RJ, Holmes MJ. Origin and transfer of toxins involved in ciguatera. Comp Biochem Phys C Pharmacol Toxicol Endocrinol 1993;106:615–628. Yasumoto T, Igarashi T, Legrand AM, et al. Structural elucidation of ciguatoxin congeners by fast-atom bombardment tandem mass spectroscopy. J Am Chem Soc 2000;122:4988–4989. Satake M, Murata M, Yasumoto T. The Structure of CTX3C, A Ciguatoxin congener isolated from cultured Gambierdiscus toxicus. Tetrahedron Lett 1993;34:1975–1978. Satake M, Fukui M, Legrand AM, Cruchet P, Yasumoto T. Isolation and structures of new cigua- toxin analogs, 2,3-dihydroxyCTX3C and 51-hydroxyCTX3C, accumulated in tropical reef fish. Tetrahedron Lett 1998;39:1197–1198. Vernoux JP, Lewis RJ. Isolation and characterisation of Caribbean ciguatoxins from the horseeye jack (Caranx latus). Toxicon 1997;35:889–900. Lewis RJ, Vernoux JP, Brereton IM. Structure of Caribbean ciguatoxin isolated from Caranx latus. J Am Chem Soc 1998;120:5914–5920. Pottier I, Vernoux JP, Jones A, Lewis RJ. Characterisation of multiple Caribbean ciguatoxins and congeners in individual specimens of horse-eye jack (Caranx latus) by high-performance liquid chromatography/mass spectrometry. Toxicon 2002;40:929–939. Pottier I, Hamilton B, Jones A, Lewis RJ, Vernoux JP. Identification of slow and fast acting toxins in a highly ciguatoxic barracuda (Sphyraena barracuda) by HPLC/MS and radiolabelled ligand binding. Toxicon 2003;42:663–672. Hamilton B, Hurbungs M, Jones A, Lewis RJ. Multiple ciguatoxins present in Indian Ocean reef fish. Toxicon 2002;40:1347–1353. Hamilton B, Hurbungs M, Vernoux JP, Jones A, Lewis RJ. Isolation and characterisation of Indian Ocean ciguatoxin. Toxicon 2002;40:685–693. Diogene J, Reverte L, Rambla-Alegre M, Del Rio V, de la Iglesia P, Campas M, et al. Identification of ciguatoxins in a shark involved in a fatal food poisoning in the Indian Ocean. Sci Rep 2017;7:8240. Lehane L, Lewis RJ. Ciguatera: recent advances but the risk remains. Int J Food Microbiol 2000;61:91–125. Murata M, Naoki H, Iwashita T, et al. Structure of maitotoxin. J Am Chem Soc 1993;115:2060– 2062.

362

Hazel Farrell, Gurjeet S. Kohli and Shauna A. Murray

[110] Holmes MJ, Lewis RJ. Purification and characterisation of large and small maitotoxins from cultured Gambierdiscus toxicus. Nat Toxins 1994;2:64–72. [111] Yokoyama A, Murata M, Oshima Y, Iwashita T, Yasumoto T. Some chemical properties of maitotoxin, a putative calcium channel agonist isolated from a marinedinoflagellate. J Biochem 1988;104:184–187. [112] Yasumoto T, Bagnis R, Vernoux JP. Toxicity of the surgeon fishes-II properties of the principal water soluble toxin. Bull Jap Soc Sci Fish 1976;42:359–365. [113] Murata M, Naoki H, Matsunaga S, Satake M, Yasumoto T. Structure and partial stereochemical assignments for maitotoxin, the most toxic and largest natural non-biopolymer. J Am Chem Soc 1994;116:7098–7107. [114] Murata M, Yasumoto T. Structure of maitotoxin, the most toxic and largest natural nonbiopolymer. J Synth Org Chem Jpn 1995;53:207–217. [115] Zheng W, DeMattei JA, Wu J-P, et al. Complete relative stereochemistry of maitotoxin. J Am Chem Soc 1996;118:7946–7968. [116] Satake M, Ishida S, Yasumoto T, Murata M, Utsumi H, Hinomoto T. Structural confirmation of maitotoxin based on complete 13C NMR assignments and the three-dimensional PFG NOESYHMQC Spectrum. J Am Chem Soc 1995;117:7019–7020. [117] Nonomura T, Sasaki M, Matsumori N, Murata M, Tachibana K, Yasumoto T. The complete structure of maitotoxin, part II: configuration of the C135 & C142 side chain and absolute configuration of the entire molecule. Angew Chem Int Ed 1996;35:1675–1678. [118] Van Dolah FM. Diversity of marine and freshwater algal toxins. In: Marcel D [ed.] Food Sci- ence and Technology. New York; Marcel Dekker: 2000. 19–44. [119] Lu X-Z, Deckey R, Jiao G-L, Ren H-F LM. Caribbean maitotoxin elevates [Ca2+]i and activates non-selective cation channels in HIT-T15 cells. World J Diabetes 2013;4(3):70–75. [120] Murata M, Gusovsky F, Sasaki M, Yokoyama A, Yasumoto T, Daly J. Effect of maitotoxin analogues on calcium influx and phosphoinositide breakdown in cultured cells. Toxicon 1991;29:1085–1096. [121] Shimizu Y, Chou HN, Bando H, Van Duyne G, Clardy J. Structure of brevetoxin A (GB-1 toxin), the most potent toxin in the Florida red tide organism Gymnodinium breve (Ptychodiscus brevis). J Am Chem Soc 1986;108:514–515. [122] Draisci R, Ferretti E, Palleschi L, et al. High levels of yessotoxin in mussels and presence of yessotoxin and homoyessotoxin in dinoflagellates of the Adriatic Sea. Toxicon 1999;37: 1187–1193. [123] Rhodes L, McNabb P, De Salas M, Briggs L, Beuzenberg V, Gladstone M. Yessotoxin production by Gonyaulax spinifera. Harmful Algae 2006;5:148–155. [124] Eiki K, Satake M, Koike K, Ogata T, Mitsuya T, Oshima Y. Confirmation of yessotoxin production by the dinoflagellate Protoceratium reticulatum in Mutsu Bay. Fish Sci 2005;71:633–638. [125] Lin -Y-Y, Risk M, Ray SM, et al. Isolation and structure of brevetoxin B from the “red tide” dinoflagellate Ptychodiscus brevis (Gymnodinium breve). J Am Chem Soc 1981;103:6773–6775. [126] Golik J, James JC, Nakanishi K, Lin -Y-Y. The structure of brevetoxin C. Tetrahedron Lett 1982;23:2535–2538. [127] Chou HN, Shimizu Y. Biosynthesis of brevetoxins. Evidence for the mixed origin of the backbone carbon chain and possible involvement of dicarboxylic acids. J Am Chem Soc 1987;109:2184–2185. [128] Rein KS, Borrone J. Polyketides from dinoflagellates: origins, pharmacology and biosynthesis. Comp Biochem Physiol B Biochem Mol Biol 1999;124:117–131. [129] Kellmann R, Stüken A, Orr RJ, Svendsen HM, Jakobsen KS. Biosynthesis and molecular genet- ics of polyketides in marine dinoflagellates. Mar Drugs 2010;8:1011–1048.

9 Gambierdiscus, the cause of ciguatera fish poisoning

363

[130] Kalaitzis JA, Chau R, Kohli GS, Murray SA, Neilan BA. Biosynthesis of toxic naturally-occurring seafood contaminants. Toxicon 2010;56:244–258. [131] Shimizu Y. Microalgal metabolites. Curr Opin Microbiol 2003;6:236–243. [132] Khosla C, Gokhale RS, Jacobsen JR, Cane DE. Tolerance and specificity of polyketide synthases. Annu Rev Biochem 1999;68:219–253. [133] Monroe EA, Van Dolah FM. The toxic dinoflagellate Karenia brevis encodes novel type I-like polyketide synthases containing discrete catalytic domains. Protist 2008;159:471–482. [134] Beauchemin M, Roy S, Daoust P, et al. Dinoflagellate tandem array gene transcripts are highly conserved and not polycistronic. Proc Natl Acad Sci USA 2012;109:15793–15798. [135] Murray SA, Garby T, Hoppenrath M, Neilan BA. Genetic diversity, morphological uniformity and polyketide production in dinoflagellates (Amphidinium, Dinoflagellata). PLoS ONE 2012;7:e38253. [136] Satake M, Ishibashi Y, Legrand AM, Yasumoto T. Isolation and structure of ciguatoxin-4A, a new ciguatoxin precursor, from cultures of dinofagellate Gambierdiscus toxicus and parrotfish Scarus gibbus. Biosci Biotechnol Biochem 1996;60:2103–2105. [137] Gillespie N, Lewis R, Burke J, Holmes M. In: The significance of the absence of ciguatoxin in a wild population of a Gambierdiscus toxicus, Proceedings of the Fifth International Coral Reef Congress, Tahiti 27 May–1 June 1985. 1985. p. 437–442. [138] Yogi K, Oshiro N, Inafuku Y, Hirama M, Yasumoto T. Detailed LC-MS/MS analysis of Ciguatoxins revealing distinct regional and species characteristics in fish and causative alga from the Pacific. Ann Chem 2011;83:8886–8891. [139] Ikehara T, Kuniyoshi K, Oshiro N, Yasumoto T. Biooxidation of ciguatoxins leads to speciesspecific toxin profiles. Toxins (Basel) 2017;9. [140] Roue M, Darius HT, Viallon J, Ung A, Gatti C, Harwood DT, et al. Application of solid phase adsorption toxin tracking (SPATT) devices for the field detection of Gambierdiscus toxins. Harmful Algae 2018;71:40–49. [141] Gudger E. Poisonous fishes and fish poisonings, with special reference to Ciguatera in the West Indies. Amn J Trop Med 1930;10:43–55. [142] Yasumoto T, Kanno K. Toxicity studies on marine snails & occurrence of toxins resembling ciguatoxin, scaritoxin, and maitotoxin in a turban shell. Bull Jap Soc Sci Fish 1976;42:1399– 1404. [143] Hokama Y, Yoshikawa-Ebesu JSM. Ciguatera fish poisoning: A foodborne disease. J Toxicol – Toxin Rev 2001;20:85–139. [144] Lewis RJ. The changing face of ciguatera. Toxicon 2001;39:97–106. [145] Dechraoui MYB, Tiedeken JA, Persad R, et al. Use of two detection methods to discriminate ciguatoxins from brevetoxins: Application to great barracuda from Florida Keys. Toxicon 2005;46:261–270. [146] Laurent D, Yeeting B, Labrosse P, Goudechoux JP. Ciguatera: field reference guide. Noumea. New Calidonia; Secretariat of the pacific community: 2005. [147] Dickey RW, Plakas SM. Ciguatera: A public health perspective. Toxicon 2010;56:123–136. [148] Wong CK, Hung P, Lee KLH, Mok T, Chung T, Kam KM. Features of Ciguatera Fish Poisoning Cases in Hong Kong 2004–2007. Biomed Env Sci 2008;21:521–527. [149] Glaziou P, Martin P. Study of factors that influence the clinical response to ciguatera fish poisoning. Toxicon 1993;31:1151–1154. [150] Adams MJ. An outbreak of ciguatera poisoning in a group of scuba-divers. J Wilderness Med 1993;4:304–311. [151] Lindsay JA. Chronic sequelae of foodborne disease. Emerging Infect Dis 1997;3:443. [152] Ting J, Brown A. Ciguatera poisoning: a global issue with common management problems. Eur J Emer Med 2001;8:295–300.

364

Hazel Farrell, Gurjeet S. Kohli and Shauna A. Murray

[153] Halstead BW. Class osteichthyes: poisonous ciguatoxic fishes. In: Halstead BW. Poisonous and venomous marine animals of the world. Princeton: Darwin Press, Inc., 1978:325–402. [154] Ciguatera fish poisoning. In Marine Biotoxins. FAO Food and Nutrition Paper 80; Food and Agriculture Organization of the United Nations: Rome, Italy, 2004; pp. 185–218. [155] Randall JE. A review of ciguatera, tropical fish poisoning, with a tentative explanation of its cause. Bull Mar Sci 1958;8:236–267. [156] Cruz-Rivera E, Villareal TA. Macroalgal palatability and the flux of ciguatera toxins through marine food webs. Harmful Algae 2006;5:497–525. [157] Caillaud A, Eixarch H, de la Iglesia P, et al. Towards the standardisation of the neuroblastoma (neuro-2a) cell-based assay for ciguatoxin-like toxicity detection in fish: application to fish caught in the Canary Islands. Food Addit Contam 2012;29:1000–1010. [158] Bravo J, Suárez F, Ramírez A, Acosta F. Ciguatera, an emerging human poisoning in Europe. J Aquac Mar Biol 2015;3:00053. [159] Sanchez-Henao JA, Garcia-Alvarez N, Fernandez A, Saavedra P, Silva Sergent F, Padilla D, et al. Predictive score and probability of CTX-like toxicity in fish samples from the official control of ciguatera in the Canary Islands. Sci Total Environ 2019;673:576–584. [160] Otero P, Perez S, Alfonso A, et al. First toxin profile of ciguateric fish in Madeira Arquipelago (Europe). Ann Chem 2010;82:6032–6039. [161] Hokama Y, Banner AH, Boylan DB. Radioimmunoassay for detection of ciguatoxin. Toxicon 1977;15:317–325. [162] Hokama Y, Abad MA, Kimura LH. A rapid enzyme-immunoassay for the detection of ciguatoxin in contaminated fish tissues. Toxicon 1983;21:817–824. [163] Campora CE, Dierking J, Tamaru CS, Hokama Y, Vincent D. Detection of ciguatoxin in fish tissue using sandwich ELISA and neuroblastoma cell bioassay. J Clin Lab Anal 2008;22: 246–253. [164] Hokama Y. A rapid, simplified enzyme immunoassay stick test for the detection of ciguatoxin and related polyethers from fish tissues. Toxicon 1985;23:939–946. [165] Poli MA, Lewis RJ, Dickey RW, Musser SM, Buckner CA, Carpenter LG. Identification of Caribbean ciguatoxins as the cause of an outbreak of fish poisoning among US soldiers in Haiti. Toxicon 1997;35:733–741. [166] Vernoux JP, Elandaloussi SA. Heterogeneity of ciguatoxins extracted from fish caught at the coast of the French Antilles. Biochimie 1986;68:287–291. [167] Granade H, Cheng PC, Doorenbos NJ, Ciguatera I. Brine shrimp (Artemia salina) larval assay for ciguatera toxins. J Pharm Sci 1976;65:1414–1415. [168] Estevez P, Castro D, Pequeno-Valtierra A, Leao JM, Vilarino O, Diogene J, et al. An Attempt to Characterize the Ciguatoxin Profile in Seriola fasciata Causing Ciguatera Fish Poisoning in Macaronesia. Toxins (Basel) 2019b;11. [169] Boada LD, Zumbado M, Luzardo OP, et al. Ciguatera fish poisoning on the West Africa Coast: an emerging risk in the Canary Islands (Spain). Toxicon 2010;56:1516–1519. [170] Pottier I, Vernoux J, Jones A, Lewis R. Analysis of toxin profiles in three different fish species causing ciguatera fish poisoning in Guadeloupe, French West Indies. Food Addit Contam 2002;19:1034–1042. [171] Pérez-Arellano J-L, Luzardo OP, Brito AP, et al. Ciguatera fish poisoning, Canary Islands. Emerging Infect Dis 2005;11:1981. [172] Campora CE, Hokama Y, Tamaru CS, Anderson B, Vincent D. Evaluating the risk of ciguatera fish poisoning from reef fish grown at marine aquaculture facilities in Hawaii. J World Aquacult Soc 2010;41:61–70.

9 Gambierdiscus, the cause of ciguatera fish poisoning

365

[173] Estevez P, Castro D, Manuel Leao J, Yasumoto T, Dickey R, Gago-Martinez A. Implementation of liquid chromatography tandem mass spectrometry for the analysis of ciguatera fish poisoning in contaminated fish samples from Atlantic coasts. Food Chem 2019a;280:8–14. [174] O’Toole AC, Bottein MYD, Danylchuk AJ, Ramsdell JS, Cooke SJ. Linking ciguatera poisoning to spatial ecology of fish: A novel approach to examining the distribution of biotoxin levels in the great barracuda by combining non-lethal blood sampling and biotelemetry. Sci Total Environ 2012;427:98–105. [175] Pottier I, Vernoux J-P, Lewis RJ. Ciguatera fish poisoning in the Caribbean islands and Western Atlantic. In: Ware GW, Nigg HN [eds.] Reviews of environmental contamination and toxicol- ogy. Tuscon, Arizona; Springer: 2001. 99–141. [176] Bagnis R, Barsinas M, Prieur C, Pompon A, Chungue E, Legrand A. The use of the mosquito bioassay for determining the toxicity to man of ciguateric fish. Biol Bull 1987;172:137–143. [177] Nunez-Vazquez EJ, Almazan-Becerril A, Lopez-Cortes DJ, Heredia-Tapia A, HernandezSandoval FE, Band-Schmidt CJ, et al. Ciguatera in Mexico (1984(-)2013). Mar Drugs 2018;17. [178] Lewis RJ, Endean R. Ciguatoxin from the flesh and viscera of the barracuda, Sphyraena jello. Toxicon 1984;22:805–810. [179] Hokama Y. Simplified solid-phase immunobead assay for detection of ciguatoxin and related polyethers. J Clin Lab Anal 1990;4:213–217. [180] Graber N, Stavinsky F, Hoffman R, Button J, Clark N, Martin S, et al. Ciguatera Fish PoisoningNew York City, 2010-2011. In: Moolenaar RL [ed.] Morbidity and Mortality Weekly Report. Centers for Disease Control and Prevention: 2013. 61–65. [181] Hung YM, Hung SY, Chou KJ, et al. Short report: Persistent bradycardia caused by ciguatoxin poisoning after barracuda fish eggs ingestion in southern Taiwan. Am J Trop Med Hyg 2005;73:1026–1027. [182] Stewart I, Eaglesham GK, Poole S, Graham G, Paulo C, Wickramasinghe W, et al. Establishing a public health analytical service based on chemical methods for detecting and quantifying Pacific ciguatoxin in fish samples. Toxicon 2010;56:804–812. [183] Wong CK, Hung P, Lo JY. Ciguatera fish poisoning in Hong Kong--a 10-year perspective on the class of ciguatoxins. Toxicon 2014;86:96–106. [184] Gaboriau M, Ponton D, Darius HT, Chinain M. Ciguatera fish toxicity in French Polynesia: size does not always matter. Toxicon 2014;84:41–50. [185] Wu N, Huan Q, Du K, Hu R, Jiang T. Ciguatera toxins in wild coral reef fish along the southern coast of China. Mar Freshw Res 2015;66:1168. [186] Mak YL, Wai TC, Murphy MB, Chan WH, Wu JJ, Lam JC, et al. Pacific ciguatoxins in food web components of coral reef systems in the Republic of Rep. Kiribati. Environ Sci Technol 2013;47:14070–14079. [187] Legrand AM, Litaudon M, Genthon GN, Bagnis R, Yasumoto T. Isolation and some properties of ciguatoxin. J Appl Phycol 1989;1:183–188. [188] Labrousse H, Matile L. Toxicological biotest on diptera larvae to detect ciguatoxins and various other toxic substances. Toxicon 1996;34:881–891. [189] Lewis RJ, Jones A. Characterization of ciguatoxins and ciguatoxin congeners present in ciguateric fish by gradient reversed-phase high-performance liquid chromatography/mass spec- trometry. Toxicon 1997;35:159–168. [190] Scheuer PJ, Takahashi W, Tsutsumi J, Yoshida T. Ciguatoxin: isolation and chemical nature. Science 1967;155:1267–1268. [191] Hardison DR, Holland WC, Darius HT, Chinain M, Tester PA, Shea D, et al. Investigation of ciguatoxins in invasive lionfish from the greater caribbean region: Implications for fishery development. PLoS One 2018;13:e0198358.

366

Hazel Farrell, Gurjeet S. Kohli and Shauna A. Murray

[192] Farrell H, Murray S, Zammit A, Edwards A. Management of ciguatoxin risk in eastern Australia. Toxins 2017;9(11):367. [193] Park DL, Ayala CE, Guzman-Perez SE, Lopez-Garcia R, Trujillo S. Microbial toxins in foods: algal, fungal, and bacterial. In: Food Toxicology. New York; CRC Press: 2001. 93–135. [194] Lonati D, Gatti CM, Zancan A, Darius HT, Fleure M, Chinain M, et al. Novel ciguatera shellfish poisoning (CSP) cluster after consumption of Tectus niloticus, a gastropod, in Nuku-Hiva, French Polynesia. Clin Toxicol 2015;53:278–278. [195] Pawlowiez R, Darius HT, Cruchet P, Rossi F, Caillaud A, Laurent D, et al. Evaluation of seafood toxicity in the Australes archipelago (French Polynesia) using the neuroblastoma cell-based assay. Food Addit Contam Part A Chem Anal Control Expo Risk Assess 2013;30:567–586. [196] Laurent D, Kerbrat A, Darius H, et al. Ciguatera Shellfish Poisoning, a new ecotoxicological phenomenon from cyanobacteria to human via giant clams. In: Food chain: new research. 2012. 1–44. [197] Darius HT, Drescher O, Ponton D, Pawlowiez R, Laurent D, Dewailly E, et al. Use of folk tests to detect ciguateric fish: a scientific evaluation of their effectiveness in Raivavae Island (Australes, French Polynesia). Food Addit Contam Part A Chem Anal Control Expo Risk Assess 2013;30:550–566. [198] Arnett MV, Lim JT. Ciguatera fish poisoning – impact for the military health care provider. Mil Med 2007;172:1012–1015. [199] Lucas RE, Lewis RJ, Taylor JM. Pacific ciguatoxin-1 associated with a large common-source outbreak of ciguatera in East Arnhem Land, Australia. Nat Toxins 1997;5:136–140. [200] Wong CK, Hung P, Lee KLH, Kam KM. Study of an outbreak of ciguatera fish poisoning in Hong Kong. Toxicon 2005;46:563–571. [201] Oshiro N, Yogi K, Asato S, Sasaki T, Tamanaha K, Hirama M, et al. Ciguatera incidence and fish toxicity in Okinawa, Japan. Toxicon 2010;56:656–661. [202] Loeffler CR, Robertson A, Flores Quintana HA, Silander MC, Smith TB, Olsen D. Ciguatoxin prevalence in 4 commercial fish species along an oceanic exposure gradient in the US Virgin Islands. Environ Toxicol Chem 2018;37:1852–1863. [203] Wong CK, Hung P, Lee KLH, Kam KM. Solid-phase extraction clean-up of ciguatoxincontaminated coral fish extracts for use in the mouse bioassay. Food Addit Contam 2009;26:236–247. [204] Costa P, Estevez P, Castro D, Soliño L, Gouveia N, Santos C, . . . Gago-Martínez A. New insights into the occurrence and toxin profile of ciguatoxins in Selvagens Islands (Madeira, Portugal). Toxins 2018;10(12):524. [205] Lechuga Solid-phas Sierraa Solián AP. Documented case of ciguatera on the Mexican Pacific coast. Nat Toxins 1995;3:415–418. [206] Sierra-Beltran AP, Cruz A, Nunez E, Del Villar LM, Cerecero J, JL O In: An overview of the marine food poisoning in Mexico, 12th World Congress on Animal, Plant and Microbial Toxins, Cuernavaca, Mexico, Sep 21-26, 1997. Cuernavaca, Mexico; 1997. p. 1493–1502. [207] Hossen V, Solino L, Leroy P, David E, Velge P, Dragacci S, et al. Contribution to the risk characterization of ciguatoxins: LOAEL estimated from eight ciguatera fish poisoning events in Guadeloupe (French West Indies). Environ Res 2015;143:100–108. [208] Robertson A, Garcia AC, Quintana HA, Smith TB, Castillo BF 2nd, Reale-Munroe K, et al. Invasive lionfish (Pterois volitans): a potential human health threat for ciguatera fish poisoning in tropical waters. Mar Drugs 2013;12:88–97. [209] Pompon A, Bagnis R. Ciguatera – a rapid procedure for extraction of ciguatoxin. Toxicon 1984;22:479–482. [210] Lewis RJ, Sellin M. Multiple ciguatoxins in the flesh of fish. Toxicon 1992;30:915–919.

9 Gambierdiscus, the cause of ciguatera fish poisoning

367

[211] Oshiro N, Matsuo T, Sakugawa S, Yogi K, Matsuda S, Yasumoto T, et al. Ciguatera fish poisoning on Kakeroma Island, Kagoshima Prefecture, Japan. Trop Med Health 2011;39:53–57. [212] Yogi K, Oshiro N, Matsuda S, Sakugawa S, Matsuo T, Yasumoto T. Toxin profiles in fish implicated in ciguatera fish poisoning in Amami and Kakeroma Islands, Kagoshima Prefecture, Japan. Shoku Mamoru-shi 2013;54:385–391. [213] Larson E, Rothman L. Ciguatera poisoning by the horse-eye jack, Caranx latus, a carangid fish from the tropical Atlantic. Toxicon 1967;5:121–124. [214] Zlotnick BA, Hintz S, Park DL, Auerbach PS. Ciguatera poisoning after ingestion of imported jellyfish – diagnostic application of serum immunoassay. Wilderness Environ Med 1995;6:288–294. [215] Solino L, Widgy S, Pautonnier A, Turquet J, Loeffler CR, Flores Quintana HA, et al. Prevalence of ciguatoxins in lionfish (Pterois spp.) from Guadeloupe, Saint Martin, and Saint Barthelmy Islands (Caribbean). Toxicon 2015;102:62–68. [216] Endean R, Griffith JK, Robins JJ, Llewellyn LE, Monks SA. Variation in the toxins present in ciguateric narrow-barred spanish mackerel, Scomberomorus commersoni. Toxicon 1993;31:723–732. [217] Farrell H, Zammit A, Harwood DT, McNabb P, Shadbolt C, Manning J, et al. Clinical diagnosis and chemical confirmation of Ciguatera Fish Poisoning in New South Wales, Australia. Commun Dis Intell 2016;40. [218] Kohli GS, Haslauer K, Sarowar C, Kretzschmar AL, Boulter M, Harwood DT, et al. Qualitative and quantitative assessment of the presence of ciguatoxin, P-CTX-1B, in Spanish Mackerel (Scomberomorus commerson) from waters in New South Wales (Australia). Toxicol Rep 2017b;4:328–334. [219] Bentur Y, Spanier E. Ciguatoxin-like substances in edible fish on the eastern Mediterranean. Clin Toxicol 2007;45:695–700. [220] Ebesu JSM, Nagai H, Hokama Y. The first reported case of human ciguatera possibly due to a farm-cultured salmon. Toxicon 1994;32:1282–1286. [221] Dinubile MJ, Hokama Y. The ciguatera poisoning syndrome from farm-raised salmon. Ann Intern Med 1995;122:113–114. [222] Park DL. Seafood safety monitoring programme for ciguatera: assessing aquatic product safety. Proc Gulf Caribb Fish Inst 1999;45:270–289. [223] Dalzell P. Management of ciguatera fish poisoning in the South Pacific. Memoirs of the Queensland Museum. Brisbane 1994;34:471–479. [224] Bottein M-YD, Kashinsky L, Wang Z, Littnan C, Ramsdell JS. Identification of ciguatoxins in Hawaiian monk seals Monachus schauinslandi from the Northwestern and Main Hawaiian Islands. Environ Sci Technol 2011;45:5403–5409. [225] Brock VE, Van Heukelem W, Helfrich P. An ecological reconnaissance of Johnston Island and the effects of dredging. Hawaii Institute of Marine Biology, University of Hawaii: 1966. [226] Loeffler CR, Handy SM, Flores Q, Deeds JR. Fish hybridization leads to uncertainty regarding Ciguatera Fish Poisoning risk; confirmation of hybridization and ciguatoxin accumulation with implications for stakeholders. J Mar Sci Eng 2019;7:105. [227] Ha DV, Uesugi A, Uchida H, Ky PX, Minh DQ, Watanabe R, et al. Identification of Causative Ciguatoxins in Red Snappers Lutjanus bohar Implicated in Ciguatera Fish Poisonings in Vietnam. Toxins (Basel) 2018;10. [228] Friedemann M. Ciguatera fish poisoning outbreaks from 2012 to 2017 in Germany caused by snappers from India, Indonesia, and Vietnam. J Verbrauch Lebensm 2018;14:71–80. [229] Mattei C, Vetter I, Eisenblätter A, et al. Ciguatera fish poisoning: a first epidemic in Germany highlights an increasing risk for European countries. Toxicon. 2014;91:76–83. doi:10.1016/j. toxicon.2014.10.016.

368

Hazel Farrell, Gurjeet S. Kohli and Shauna A. Murray

[230] Rajisha, R., Kishore P., Panda, S. K., Ravishankar, C. N. and Kumar, A. K. (2017a) Confirmation of Ciguatoxin Fish Poisoning in Red Snapper, Lutjanus bohar (Forsskål, 1775) by Mouse Bioassay. Fish. Technol. 54: 1–4. [231] Hsieh CH, Hwang KL, Lee MM, Lan CH, Lin WF, Hwang DF. Species identification of ciguatoxincarrying grouper implicated in food poisoning. J Food Prot. 2009;72(11):2375–2379. doi:10.4315/0362-028x-72.11.2375. [232] Chen, T. Y., Chen, N. H., Lin, W. F., Hwang, K. L., Huang, Y. C., & Hwang, D. F. (2010). Identification of Causative Fish for a Food Poisioning in Taiwan by Using SDS-PAGE Technique. Journal of Marine Science and Technology, 18(4), 593–596. [233] Lin, W. F., Lyu, Y. C., Wu, Y. J., Lu, C. H., & Hwang, D. F. (2012). Species identification of snapper: A food poisoning incident in Taiwan. Food Control, 25(2), 511–515. [234] Hoffman PA, Granade HR, McMillan JP. The mouse ciguatoxin bioassay – a dose-response curve and symptomatology analysis. Toxicon 1983;21:363–369. [235] Parrilla-Cerrillo MC, Vázquez-Castellanos JL, Saldate-Castañeda EO, Nava-Fernández LM. Outbreaks of food poisonings of microbial and parasitic origins. Brotes de toxiinfecciones alimentarias de origen microbiano y parasitario 1993;35:456–463. [236] Darius HT, Roue M, Sibat M, Viallon J, Gatti C, Vandersea MW, et al. Toxicological Investigations on the Sea Urchin Tripneustes gratilla (Toxopneustidae, Echinoid) from Anaho Bay (Nuku Hiva, French Polynesia): Evidence for the Presence of Pacific Ciguatoxins. Mar Drugs 2018;16(4):122. [237] Silva M, Rodriguez I, Barreiro A, Kaufmann M, Isabel Neto A, Hassouani M, et al. First Report of Ciguatoxins in Two Starfish Species: Ophidiaster ophidianus and Marthasterias glacialis. Toxins (Basel) 2015;7:3740–3757. [238] Bagnis R, Bennett J, Prieur C, Legrand AM In: The dynamics of three toxic benthic dinoflagellates and the toxicity of ciguateric surgeonfish in French Polynesia, Third International Conference on Toxinc Dinoflagellates, St. Andrews, New Brunswick, Canada, June 8–12, 1985. Andersen D, White A, Baden DG [eds.] St. Andrews, New Brunswick, Canada; Elsevier: 1985. p. 177–182. [239] Boisier P, Ranaivoson G, Rasolofonirina N, et al. Fatal mass poisoning in Madagascar following ingestion of a shark (Carcharhinus leucas) – clinical and epidemiologic aspects and isolation of toxins. Toxicon 1995;33:1359–1364. [240] Lewis RJ, Yang A, Jones A. Rapid extraction combined with LC-tandem mass spectrometry (CREM-LC/MS/MS) for the determination of ciguatoxins in ciguateric fish flesh. Toxicon 2009;54:62–66. [241] Hirama M, Oishi T, Uehara H, et al. Total synthesis of ciguatoxin CTX3C. Science 2001;294:1904–1907. [242] Guzmán-Pérez SE and Park DL, 2000. Ciguatera toxins: Chemistry and detection. In: Seafood and freshwater toxins: Pharmacology, physiology and detection. Ed Botana LM. Marcel Dekker, New York, 401–418. [243] Darius H, Roué M, Sibat M, Viallon J, Gatti C, Vandersea M, . . . Chinain M. Tectus niloticus (Tegulidae, gastropod) as a novel vector of ciguatera poisoning: Detection of Pacific ciguatoxins in toxic samples from Nuku Hiva Island (French Polynesia). Toxins 2018;10(1):2. [244] Larsson, M.E.; Laczka, O.F.; Harwood, D.T.; Lewis, R.J.; Himaya, S.W.A.; Murray, S.A.; Doblin, M.A. Toxicology of Gambierdiscusspp. (Dinophyceae) from Tropical and Temperate Australian Waters. Mar. Drugs 2018, 16, 7.

Mireille Chinain, Clémence M.i. Gatti, Hélène Martin-Yken, Mélanie Roué and H. Taiana Darius

10 Ciguatera poisoning: an increasing burden for Pacific island communities in light of climate change? 10.1 Introduction Harmful algal blooms (HABs) are natural phenomena that have apparently expanded worldwide in recent decades [1–3]. They can cause damage to the environment and wild life, alter marine food webs and threaten seafood safety and human health [4–6], thus causing significant socioeconomic problems most notably among world populations that rely heavily on marine resources for their subsistence [7]. It is suggested that this observed HAB increase is linked to both anthropogenic influences on the biosphere and naturally occurring environmental changes such as climate change [8, and references therein, 9, 10], and that the current projections of water temperature warming will eventually result in even greater problems due to HABs in the future [11, and references therein]. Among the many algal toxins that can find their way through marine food webs to human consumers where they cause various poisoning illnesses, are ciguatoxins (CTXs), potent neurotoxins produced by dinoflagellates in the genera Gambierdiscus and Fukuyoa. CTXs are responsible for ciguatera poisoning (CP), a disease resulting from the consumption of poisonous coral-reef finfish and marine invertebrates [12–14]. With an estimated occurrence of 10,000 to 50,000 cases per year [15], CP is regarded as the most prevalent noninfectious seafood-borne disease worldwide. This poisonous syndrome is particularly well known to communities of the Pacific island countries and territories (PICTs) where the highest incidence rates (IRs) in the world are consistently reported since the 1970s [15–20]. This has prompted extensive studies in various Pacific ciguatera hot spots to assess the genetic diversity and distribution of the causative algal organisms [14, 21–28], the marine species at risk [14, 18, 29–34], the suite of toxins involved [27, 35–43], the socioeconomic impacts on the well-being of local communities [44–49], and so forth. But, despite significant advances in the understanding of this complex eco-toxicological phenomenon, CP outbreaks are still difficult to predict and the prevalence of this disease is dangerously high in PICTs.

Mireille Chinain, Clémence M.i. Gatti, H. Taiana Darius, Laboratory of Marine Biotoxins, Institut Louis Malardé – UMR EIO, Papeete-Tahiti, French Polynesia Hélène Martin-Yken, Mélanie Roué, Institut de Recherche pour le Développement – UMR EIO, Faa’a-Tahiti, French Polynesia https://doi.org/10.1515/9783110625738-010

370

Mireille Chinain et al.

The present paper addresses several ciguatera-related issues specific to PICTs and neighboring countries in the Oceania region. A first section describes the genetic heterogeneity and chemodiversity of Gambierdiscus and Fukuyoa species occurring in the area and how climate change is likely to influence CP risk locally as illustrated by the recent expansion of ciguatera to previously unaffected areas. A second section presents the epidemiology of CP in the region and the survey networks currently in place that have allowed the identification of novel vectors of ciguatera among marine invertebrates highly prized by local populations. A further section deals with ciguatera perception among Pacific island communities and the various adaptive strategies and traditional practices developed to cope with this poisoning risk. The fourth and last section addresses the issue of the implementation of surveillance capabilities and risk management programs in PICTs, by highlighting the potential interest and applicability of several monitoring and detection tools particularly well adapted to the context of the Pacific Islands, and the importance of outreach and communication strategies to modify high risk-taking behaviors among local communities. A concluding section outlines some of the issues that remain to be addressed for increased food safety among the island communities of the Pacific region.

10.2 Ciguatera poisoning in the context of climate change 10.2.1 The Oceania region, a “biodiversity hot spot” for Gambierdiscus species 10.2.1.1 Geographical distribution of Gambierdiscus/Fukuyoa species in the Oceania region Among the major locales presently regarded as CP-endemic areas, the Oceania region is likely the area where the diversity and geographic distribution of Gambierdiscus and Fukuyoa species are best documented. Today, this region continues to be the site of extensive sampling efforts. As a result, five of the eight newly described Gambierdiscus species since 2016 originated from various South Pacific locales, that is, G. cheloniae [50], G. honu [51], G. lapillus [52], G. lewisii and G. holmesii [53], and it is highly likely that additional species will be characterized in the near future, as the genus Gambierdiscus is also comprised of four unidentified clades that are genetically distinct and may constitute new undescribed species [24, 54–56]. Currently 16 Gambierdiscus/ Fukuyoa spp. out of the 21 published species are known in the South Pacific, the five species not yet reported in this region being G. carolinianus, G. jejuensis (formerly known as Gambierdiscus sp. type 2), G. silvae, F. rutzleri and F. yasumotoi. Table 10.1

10 Ciguatera poisoning in the Oceania region

371

Table 10.1: Distribution of Gambierdiscus and Fukuyoa species and phylotypes in the Oceania region. All species were identified using molecular analysis (PCR, sequencing and phylogenetic analyses). Locations

Species or phylotypes

References

Australia

G. belizeanus, G. carpenteri, G. honu, G. lapillus, G. toxicus, G. lewisii, G. holmesii, F. paulensis

[, , , –, –]

Cook Islands

G. australes, G. pacificus, G. polynesiensis, G. cheloniae, G. honu, G. lapillus, G. excentricus

[, , , , –]

Fiji

G. carpenteri, G. toxicus

[]

French Polynesia

G. australes, G. caribaeus, G. carpenteri, G. pacificus, G. polynesiensis, G. toxicus, G. honu

[, , , , –] Smith, pers. comm.]

New Zealand

Mainland : F. paulensis Kermadec Islands : G. australes, G. honu, G. polynesiensis, G. pacificus, G. carpenteri

[ , , , , , , ]

Guam

G. carpenteri, G. toxicus

[, ]

Hawaii

G. australes, G. caribaeus, G. carpenteri, G. toxicus

[, , , ]

Kiribati

G. balechii, G. belizeanus, G. carpenteri, G. pacificus, Gambierdiscus sp. type , 

[, ]

Marshall Islands

G. pacificus, G. toxicus

[]

New Caledonia

G. toxicus

[]

Northern G. carpenteri Mariana Islands

[]

Palau

G. caribaeus, G. toxicus, G. honu

[, ]

Tonga

G. australes, G. honu, G. pacificus, G. polynesiensis, G. cheloniae, G. carpenteri

[, ]

and Figure 10.1 provide an updated summary of the current distribution of Gambierdiscus and Fukuyoa species in the Oceania region. From Figure 10.1, it is clear that several island groups in the Pacific such as French Polynesia, the Cook Islands, the Kingdom of Tonga, the Republic of Kiribati and the Kermadec Islands could represent “biodiversity hot spots” of Gambierdiscus with the report of at least five distinct species in each of these island groups (Table 10.1 and references therein). Of note, up to six Gambierdiscus species were found to coexist within a single sampling location of Rikitea Bay on Mangareva Island, a long-standing ciguatera hot spot in the Gambier archipelago (French Polynesia) [73]. Similarly, eight

372

Mireille Chinain et al.

Northern Mariana Islds. Guam

Hawaii Marshall Islds.

Palau

Kiribati

Cook Islds.

French Polynesia

Fiji

Australia

Tonga

New Caledonia

Kermadec Islds. New Zealand

G. australes G. balechii G. belizeanus G. caribaeus G. carpenteri

G. cheloniae G. excentricus G. holmesii G. honu G. lapillus

G. lewisii G. pacificus G. polynesiensis G. toxicus

Gambierdiscus sp. type 4 Gambierdiscus sp. type 5 F. paulensis

Figure 10.1: Gambierdiscus and Fukuyoa species distribution in the Oceania region (see Table 10.1 for corresponding references). © Institut Louis Malardé.

species are currently known in Australia from tropical Queensland to the more temperate waters of New South Wales. This increase in species number has raised concerns about the future trends of ciguatera in this CP-prone region, as it is believed that the global distribution of CP will dramatically increase in the coming years in the context of climate change and global warming [74–76]. Indeed, average annual increases in sea water temperatures have been reported for both south-eastern Australia and for New Zealand’s coastal waters [77, 78], consistent with occurrence reports of both Fukuyoa and Gambierdiscus in the subtropical northern region of New Zealand and Kermadec Islands [25, 28, 64, 72], and in the more temperate waters of New South Wales, Australia [27, 57, 60]. It has been suggested that blooms of G. carpenteri populations stretching as far as 37°S, south of Sidney (New South Wales) may be explained by currents bringing cells from tropical Queensland to the warming New South Wales waters [57, 79]. It should be noted that recent observations indicate that Gambierdiscus and/or Fukuyoa spp. are also presently established in other temperate-like areas globally, including Korea [59, 80], Japan [81], the northern Gulf of Mexico [82] and the Mediterranean Sea [83, 84]. In particular, G. australes is among the few species that appear to have the widest latitudinal ranges

10 Ciguatera poisoning in the Oceania region

373

from 40.00°N to 35.25°S [65]. But whether this current expansion of the geographic range of Gambierdiscus/Fukuyoa spp. in South Pacific areas may lead to increasing risk of CP outbreaks locally is still a matter of current debate: for example, there is no concurrent report of human health impacts in northern New Zealand, yet [28], while in New South Wales, there has been an apparent increase in CP from both imported fish and Spanish mackerel (Scomberomorus commerson) caught in the region’s coastal waters [85, 86].

10.2.1.2 Chemodiversity in Gambierdiscus/Fukuyoa species from the Oceania region Gaining detailed insights into the suite of toxins produced by Gambierdiscus/Fukuyoa spp. is critical to better ascertain the real threat high abundance of populations of these organisms poses to human health. Not all Gambierdiscus/Fukuyoa species produce CTXs and maitotoxins (MTXs), the main cyclic polyether compounds regarded as the causative agents of CP. It is believed that the production of CTXs is limited to certain genetic species/strains of Gambierdiscus/Fukuyoa [87 for review and references therein]. To date, Pacific-CTX (P-CTXs) congeners have been formally detected, through isolation and/or liquid chromatography coupled to tandem mass spectrometry (LC-MS/MS) analyses, in only two of the 16 Gambierdiscus species reported in the Pacific Ocean (Figure 10.1), that is, in G. polynesiensis strains from French Polynesia, the Cook Islands and the Kermadec Islands [28, 40–43, 50, 88–91], and in G. toxicus strains from French Polynesia [36, 38, 92, 93] (Table 10.2). However, all strain identifications made before the introduction of molecular analyses in 1999 [67] should probably be reassessed, particularly concerning the CTX-producing G. toxicus from French Polynesia, which is most likely a G. polynesiensis strain [94]. For now, none of the P-CTXs searched for by LC-MS/MS (i.e., P-CTX3B, P-CTX3C, P-CTX4A and PCTX4B) were detected in strains of G. australes, G. carpenteri, G. cheloniae, G. holmesii, G. honu, G. lapillus, G. lewisii and G. pacificus from Australia, the Cook Islands, French Polynesia, the Kermadec Islands and the Kingdom of Tonga [25, 27, 28, 41, 42, 50–53, 64, 72, 89, 95], nor in strains of F. paulensis from Australia and New Zealand [28, 42, 60, 72]. Moreover, to the best of our knowledge, Pacific strains of G. balechii, G. belizeanus, G. caribaeus, G. excentricus and Gambierdiscus sp. type 4 and 5 have not been analyzed for the presence of P-CTXs, yet, using LC-MS/MS. However, using the mouse bioassay (MBA), radioactive receptor-binding assay (rRBA), or neuroblastoma cell-based assay (CBA-N2a), CTX-like activities have been evidenced in almost all Pacific species of Gambierdiscus, but at ranges well below those reported for G. polynesiensis strains [87 for review and references therein]. Gambierdiscus is also known to produce several other cyclic polyether compounds, which were characterized from Pacific strains for the majority of them. MTX1, for instance, was first isolated from a French Polynesian strain identified as G. toxicus (potential misidentification) [96], and further detected in numerous G. australes strains from the Cook Islands and the Kermadec Islands by means of LC-

374

Mireille Chinain et al.

Table 10.2: List of the Pacific ciguatoxins detected through isolation or liquid chromatography coupled to tandem mass spectrometry (LC-MS/MS) analyses in Gambierdiscus strains isolated from the South Pacific Ocean. Species (a)

G. polynesiensis

G. toxicus(b)

Geographical origin P-CTXs congeners

References

French Polynesia

P-CTXB P-CTXC M-seco-P-CTXC -hydroxy-P-CTXC P-CTXA P-CTXB

[, , , , ]

Cook Islands

P-CTXB P-CTXC P-CTXA P-CTXB

[, , , ]

Kermadec Islands

P-CTXC

[]

French Polynesia

P-CTXB [, , , ] P-CTXC M-seco-P-CTXC M-seco-P-CTXC methyl acetal -hydroxy-P-CTXC P-CTXA P-CTXB M-seco-P-CTXA/B P-CTX (= -epi--deoxy-P-CTXB) P-CTX (= -deoxy-P-CTXB)

a

Some uncharacterized P-CTX isomers have also been detected in several G. polynesiensis strains [28, 40, 43, 90, 91]. b Potential misidentification of the species, most likely G. polynesiensis [94].

MS/MS [25, 28, 41, 42, 50, 64, 72, 89, 95]. The putative MTX2, whose structure has not yet been elucidated, was reportedly found in cultures of G. toxicus, G. caribaeus and G. pacificus from Australia, Hawaii and French Polynesia [37, 97, 98]. Of note, a third MTX analogue, named MTX4, has been recently described in G. excentricus strains from the Atlantic Ocean [98], and efforts to document its production in Pacific strains as well are currently underway. Gambieric acids (GA)-A, -B, -C and -D, gambierol and gambieroxide were isolated from French Polynesian strains also identified as G. toxicus (potential misidentification) [99–101]. Interestingly, gambierone, which was originally characterized from a Caribbean G. belizeanus strain [102], was recently detected in several G. polynesiensis strains from French Polynesia using LC-MS/MS [43]. Finally, the presence of 44-methylgambierone, formerly known as putative MTX3 [37], and whose structure has been fully characterized only recently [103, 104], has been demonstrated in almost all Gambierdiscus strains/species analyzed from Australia, the

10 Ciguatera poisoning in the Oceania region

375

Cook Islands, French Polynesia, Hawaii, the Kermadec Islands, and the Kingdom of Tonga, that is, G. australes, G. caribaeus, G. carpenteri, G. cheloniae, G. holmesii, G. honu, G. lapillus, G. lewisii, G. pacificus, G. polynesiensis and G. toxicus [25, 27, 28, 41– 43, 50–53, 64, 72, 89, 90, 95, 98], as well as in strains of F. paulensis from Australia and New Zealand [28, 42, 60, 72], thus confirming the ubiquitous character of this compound. Only few temperate clones of G. carpenteri from Australia have been found free of 44-methylgambierone [27, 57, 89], while data are not yet available for Pacific strains of G. balechii, G. belizeanus, G. excentricus and Gambierdiscus sp. type 4 and 5. While CTXs and, to a lesser extent, MTXs are believed to play a prominent role in CP outbreaks, it is not fully understood whether the other metabolites produced by Gambierdiscus and Fukuyoa spp. could also contribute to the clinical features of ciguatera, but most of them are considered as compounds of interest for their bioactivity and/or potential therapeutic applications [103, 105]. For instance, GA-A and gambierol are known to act as functional antagonists of the same binding site than CTXs on voltage-gated sodium channels (VGSCs), while gambierone has been shown to cause VGSCs activation in a similar pattern as CTXs, although with much less potency. In addition, both gambierone and 44-methylgambierone are able to induce a small rise in the cytosolic calcium concentration in human cortical neurons as CTXs [87 for review and references therein].

10.2.2 Ciguatera poisoning occurrences in the context of climate change 10.2.2.1 Influence of environmental factors It is well known that reports of CP events/outbreaks are significantly underestimated because of the failure to recognize its symptoms [49], limited collection of epidemiological data on a global level [20 for review and references therein, 106] and even a generalized reticence to report ciguatera in endemic regions (e.g., within island communities) since many patients tend to rely on traditional medicine for treatment [107, 108]. Moreover, the lack of standardization in data collection methodologies and timelines makes it difficult to compare incidence data across regions. Thereby, it is difficult to conclude whether there has been a global increase in the frequency of CP incidents over the past decades, and even more difficult to link such observations to climate change impacts. Many laboratory studies have investigated Gambierdiscus/Fukuyoa spp. growth responses and toxin production under varying environmental factors [87, 109 for reviews and references therein]. Temperature and pH are some of the environmental drivers most representative of climate change, for example, global ocean warming and acidification [110–114]. The effects of temperature changes are, by far, the best documented

376

Mireille Chinain et al.

in the literature. However, it should be mentioned that many of the previous studies were conducted at a time when Gambierdiscus taxonomy was unresolved, so the extent to which, for example, growth responses may vary across the multiple species now known in Gambierdiscus and Fukuyoa genera remains to be clarified. To date, growth data for the following species/phylotypes: G. australes, G. belizeanus, G. caribaeus, G. carolinianus, G. carpenteri, G. jejuensis G. pacificus, G. polynesiensis, G. scabrosus, G. silvae, F. ruetzleri, and Gambierdiscus spp. types 3, 4 and 5 can be found in the literature [40, 59, 115–119], and results confirm that differences in both tolerance and optimum growth ranges exist not only across species but strains as well [119]. Based on sea surface temperature projections over the coming century, a substantial shift in both the distribution and abundance of ciguatera dinoflagellates is to be expected [120, 121], with some species becoming dominant whereas others will become less prevalent. In the long-term, however, temperatures may get too warm according to Llewellyn [76], thereby hindering Gambierdiscus/Fukuyoa growth and resulting in a lower risk of ciguatera. In any case, as outlined by Tester et al. [65]: “as surface waters warm, range extension (of harmful benthic microalgae) of several degrees of latitude are anticipated, but only where species-specific habitat requirements can be met (e.g., temperature, suitable substrate, low turbulence, light, salinity, pH).” There are very few published data about the influence of pH on the growth and/or toxicity of ciguatera-related organisms. Actually, only one laboratory study has examined the linked effects of pH, temperature and salinity on the cell viability of the marine diatom Phaeodactylum tricornutum and has found that low pH combined with high temperature and salinity showed adverse effects on cultures of P. tricornutum [122]. Concerning the ciguatera-causing dinoflagellates, recent studies on G. polynesiensis suggest lower growth rates but increased ciguatoxicity in cultured strains at low pH values (Longo et al., submitted), which is in marked contrast with data obtained for strains of G. balechii and Gambierdiscus sp. type 5 that showed a growth suppression at elevated pH, and reduced ciguatoxicity at decreasing pH values (Chan, personal comm.). Numerous field observations linking climate change to CP outbreaks in the tropical Pacific can be found in the literature: cyclical weather patterns such as El Niño, associated with unusual warming of Pacific Ocean waters, have resulted in spikes of ciguatera cases in the Republic of Kiribati, Western Samoa, the State of Tuvalu and the Cook Islands [74], consistent with observations by Gingold et al. [123] who found an association between CP incidence and warmer sea surface temperatures in the Caribbean basin. A general consensus is that the massive colonization of dead corals by macroalgae provides more substrate for the settlement of epiphytic Gambierdiscus/Fukuyoa spp. populations [62, 124], as first hypothesized by Randall [125]. This may explain why reef disturbances or degradations linked to extreme climatic events (e.g., hurricanes, heavy rains, coral bleaching) or human activities (e.g., dredging and filling, constructions, military activities) frequently precede ciguatoxic events [13, 126, 127]. In the Cook Islands and French Polynesia, for instance, CP outbreaks were associated with both

10 Ciguatera poisoning in the Oceania region

377

cyclone activity and infestations of crown-of-thorns starfish (Acanthaster planci) [48, 128, 129]. Similarly, while studying the fluctuations of G. toxicus populations in a ciguateric site in Tahiti Island (French Polynesia), Chinain et al. [130] found an increase in both the density and frequency of Gambierdiscus blooms following unusually high water temperatures, concomitant with a severe coral bleaching episode affecting large areas of the study site. Of note, Gambierdiscus spp. cell density peaked approximately 10 months after water temperature started to increase [130]. Further, health impacts (peak number of CP cases) were recorded three months after these peak densities of Gambierdiscus [131], indicating that a time frame of approximately 13 months is required for increased cases of CP to be observed after climate-induced changes to sea water temperature. Such findings substantiate the idea that global warming as a result of climate change may increase the incidence of CP in the South Pacific. Additionally, climate change projections show that the intensity of hurricanes will also increase in the tropical Pacific [112].

10.2.2.2 Geographical expansion of ciguatera to novel areas: the case study of Rapa Island (Australes archipelago, French Polynesia) Ciguatera-endemic regions classically overlay coral development areas between 35°N and 35°S, that is, tropical and subtropical areas of the Caribbean Sea, and Pacific and Indian Oceans [109]. However, recent observations clearly highlight the current expansion of Gambierdiscus spp. and ciguatoxic fish to previously unaffected areas, most notably in temperate locales such as the west coast of Africa [132] and the eastern Atlantic Ocean (Canary Islands, Madeira, Selvagens Islands) where reports of CP endogenous cases are documented since 2004 [133–135]. These observations are consistent with the concurrent detection of toxigenic Gambierdiscus spp. in Macaronesian waters [71, 98, 136–139], and the confirmation of CTX contamination in locally sourced fish and invertebrates [133, 140–147]. As previously mentioned, similar observations are also documented in the southern hemisphere, such as in New South Wales, Australia [85, 86, 148]. The case of Rapa Island (Australes archipelago, French Polynesia) is another emblematic example of the current extension of CP to subtropical areas of the South Pacific. Rapa Island (27°38’ S – 144°20ʹ W) was reputed free of ciguatera until 2008 when six isolated CP cases were first reported [149]. Further, from August 2009 to November 2010, a mass-poisoning outbreak involving 114 local residents (based on declared cases, but it is highly likely that half the population was affected) occurred following community fishing in the rāhui zone, a widely used practice in many PICTs to protect local fishing resources. The unusual severity and magnitude of this outbreak resulted in two fatalities. Clinical records indicated symptoms typical of ciguatera in affected individuals, including cold allodynia, paresthesia and dysesthesia that were reported

378

Mireille Chinain et al.

in 92%, 95% and 86% of affected patients, respectively, at the peak of the epidemic (Table 10.3). Moreover, three fish species were primarily involved in this mass-CP incident, namely, Leptoscarus vaigiensis (seagrass parrotfish), Kyphosus cinerascens (highfin chub) and Seriola lalandi (king fish), which were responsible for nearly 80% of the reported cases between 2009 and 2010. Field investigations revealed the presence of toxic Gambierdiscus populations in several sampling locations around Rapa Island [149]. Additionally, a random test survey of the toxic status of ≈ 250 herbivorous and carnivorous fish collected from the main fishing sites allowed confirmation of CTXs at concentrations well above the US Food and Drug Administration (FDA) advisory level (0.01 ppb) [150] in > 60% of the tested fish [149]. Fortunately, the local population was highly reactive to educational and community outreach interventions, as illustrated by a five-fold reduction in CP cases between 2009 and 2012 (Figure 10.2). Recent data (i.e., mean IR around 1280 cases per 100,000 inhabitants over the past 2 years) indicate that ciguatera is still present in Rapa Island but increased awareness among the local population about this toxic threat has helped maintain a low prevalence as compared to the alarming IR peak of 17,500 cases per 100,000 inhabitants reported in 2009. A careful analysis of the clinical forms shows that these good results are clearly attributed to self-regulating behavior among the population toward avoidance of high-risk fish species and fishing areas, as evidenced

Table 10.3: Clinical signs reported among patients during the 2009–2010 ciguatera poisoning outbreak in Rapa Island (Australes archipelago, French Polynesia). 



Number of declared cases





Gastrointestinal disorders Nausea/vomiting Diarrhea

 (%)  (%)

 (%)  (%)

Neurological and systemic disorders Cold allodynia Paresthesia Dysesthesia Itching Dizziness/headaches Muscular weakness/pains/cramps Joint pains Chills/hypothermia Burning sensation (throat, mouth)/“metallic” taste Urogenital burning Behavorial disorders (agitation, disorientation)

 (%)  (%)  (%)  (%)  (%)  (%)  (%)  (%)  (%)  (%)  (%)

 (%)  (%)  (%)  (%)  (%)  (%)  (%)  (%)  (%)  (%)  (%)

Source of data: Bureau de Veille Sanitaire, Public Health Directorate of French Polynesia and Institut Louis Malardé.

379

Miscellaneous Leptoscarus vaigiensis Kyphosus cinerascens Seriola lalandi Number of declared CP cases

*

100 90 80 70 60 50 40 30 20 10 0

Number of cases

19 20

18 20

17 20

16 20

20 15

14 20

13 20

12 20

11 20

10 20

20

20

09

100 90 80 70 60 50 40 30 20 10 0 08

% of fish involved in CP cases

10 Ciguatera poisoning in the Oceania region

* No information available for 2017

Figure 10.2: Number of cases and percentage of fish implicated in ciguatera poisoning events recorded between 2008 and 2019 in Rapa Island (Australes archipelago, French Polynesia). © Institut Louis Malardé.

by a noticeable shift in the fish species implicated in CP cases in the ensuing years (Figure 10.2). Of note, similar encouraging results following public outreach interventions had been previously obtained in Raivavae Island, another ciguatera hot spot in the Australes archipelago [69]. Further investigations are currently underway in order to ascertain the link between this emerging poisoning incident in Rapa Island and global warming in the area (Chinain et al., submitted).

10.3 Epidemiology of ciguatera poisoning in the Pacific 10.3.1 The Pacific, a region with the highest ciguatera poisoning incidence rates in the world Historically, the first description of ciguatera-related syndromes in the South Pacific was provided in 1606 by Pedro Fernandes de Queirós (1565–1614) who reported a mass-poisoning event among crew members following the consumption of a snapper in the Republic of Vanuatu. Later, in the twentieth century, some Pacific countries such as French Polynesia, New-Caledonia, the Territory of Wallis and Futuna Islands, the Republic of Marshall Islands and the Republic of Kiribati experienced a substantial

380

Mireille Chinain et al.

number of CP outbreaks linked to military activities during and following the World War II (battles, naval ships activities, war material dumping, nuclear test, etc.) [127, 151]. Based on the South Pacific Epidemiological and Health Information Service (SPEHIS) database [152], PICTs are among the nations with the highest IRs in the world as compared to other affected regions of the globe [153]. This was particularly true during the last two decades of the twentieth century. Of note, a survey conducted among Marshallese families in 1982, showed that 56% of them counted at least one family member with a previous CP in the past [153]. In Hao Island (French Polynesia), 43% of its inhabitants were affected by CP, 2 years after the onset of military activities in the atoll [154]. According to Skinner et al. [19], it is estimated that ≈ 500,000 people from 17 PICTs where affected by CP over a period of 35 years, with 39,677 CP cases recorded between 1998 and 2008. This corresponds to an annual IR of 194/100,000 inhabitants, and represents a 60% increase compared to the IR of 104/100,000 recorded for the period 1973–1983. Based on more recent data, the Republic of Fiji, French Polynesia, the Republic of Kiribati, Queensland (Australia), the Republic of the Marshall Islands, the Cook Islands, Hawaii and the Territory of Wallis and Futuna Islands (where CP was not considered as a concern until 2010), cumulated over 31,461 cases between 2000 and 2016 [19, 20, 155–161]. It should be noted that the toxic status in a given country/island may significantly vary from one year to another, as illustrated by the example of Rapa Island (French Polynesia) where IRs went from 1230/100,000 in 2008 up to 17,500/ 100,000 inhabitants in 2009 [155]. Likewise, since the 1980s, IRs reported from several South Pacific countries such as the Republic of Kiribati, the State of Tuvalu, the Tokelau Islands and the Cook Islands may sporadically exceed 1000/100,000 inhabitants [20, 152]. This situation is all the more worrying since it is recognized that CP cases reported in the Pacific likely represent only 20% of the actual cases [19]. Over the past two decades, however, a significant decrease in data collection is observed in the Pacific due to a progressive decline in the documentation of CP incidents in the SPEHIS database by participating countries. To date, fragmented CP data exist for the Oceania region [20, 106] with only four countries that currently have specific records of CP cases: (i) the Republic of Fiji, where CP is currently considered as a “high priority” by health authorities [162]; (ii) French Polynesia, where specific CP epidemiological surveillance tools and strategies are implemented since 2007 (see Section 10.3.2.1); (iii) Hawaii [163]; and (iv) Australia, where a “National Ciguatera Fish Poisoning Research Strategy” has been recently launched with the objective of reducing CP incidence through improved risk management [86]. The high IRs consistently reported in the Oceania region largely result from the strong reliance of local communities on marine resources [7, 18, 69, 164–166]. Indeed, in these developing countries, fish consumption rates are among the highest in the world due to limited range of crops and animal protein, with an average annual fish consumption rate between 10.2–167 kg capita–1 year–1 in PICTs [7, 47] versus a global rate of 18.8 kg capita–1 year–1 in 2011, 25 kg capita–1 year–1 in Europe, and 7 kg capita–1 year–1 in the USA [167]. There is also a strong tradition of

10 Ciguatera poisoning in the Oceania region

381

eating fresh fish in PICTs, with 30–90% of the fish caught by the household [7, 47, 166]. But despite these alarming statistics, ciguatera is still largely ignored by most PICT national governments, mostly due to a lack of resources that are dedicated primarily to the control of noncommunicable diseases that represent an unprecedented public health issue in the region.

10.3.2 Survey networks in the Pacific: the example of French Polynesia French Polynesia (276,300 inhabitants), is composed of 121 high islands and atolls stretching from 134° W to 155° W and 8° S to 28° S [168]. Although it has been suggested that the celebrated Polynesian voyages noted from AD 1000 to 1450 in eastern Polynesia may have been prompted by CP [169], the earliest account of CP in French Polynesian dates back to the eighteenth century and was provided by James Morrison, leading rate on board the HMS Bounty, while mooring in the Society archipelago [[170], as cited in [171]]. Later, in the nineteenth century, accounts of the presence of poisonous fish in the Tuamotu, the Gambier and the Marquesas archipelagoes were successively provided by the explorer Jacques-Antoine Moerenhout, Father Honoré Laval and Father Pierre, respectively [171]. French Polynesia is also the site of nearly five decades of research on ciguatera: following a major poisoning outbreak on Bora Bora Island in 1964, which resulted in three fatalities [172], a dedicated ciguatera research unit was created in 1967 at the Institut Louis Malardé (ILM) to investigate the etiology of this enigmatic poisoning (www.ilm.pf). This initiative eventually led to the identification of the causative agent almost a decade later, the dinoflagellate Gambierdiscus toxicus, named after the Gambier Island where it was first described [173, 174]. In a recent survey conducted on Moorea Island (Society archipelago, French Polynesia), over 50% of households interviewed declared they consume fish six to seven times a week, with 76% of them having at least one member of the household actively involved in local reef fishing activity [175]. This heavy dependence on fish resources for subsistence explains why French Polynesian communities are so highly exposed to CP risk: between 1999 and 2005, CP was responsible for 96% of hospitalized cases registered under the codification T61: “Toxic effect of noxious substances eaten as seafood” (International Statistical Classification of Diseases and Related Health Problem (ICD-10) of the World Health Organization) [176]. Since 2007, CP annual IRs have varied from 75 to 233 cases/100,000 inhabitants (i.e., 202 to 615 cases reported annually), but this number has sporadically reached 17,500/100,000 inhabitants in 2009 as previously reported [177]. Moreover, no archipelago is spared from ciguatera, from the extreme north of the territory that is subject to a tropical climate, to the extreme south exposed to a subtropical climate. The general trend presently observed in French Polynesia as a whole points toward a stable annual IR, although rates can significantly differ between islands and, in a given island, from one year to another (www.ciguatera.pf).

382

Mireille Chinain et al.

10.3.2.1 Since 2007, a country-wide epidemiological surveillance program Despite the high IRs recorded over the past decades in French Polynesia, CP is not a notifiable disease mainly due to the (too) high prevalence of this illness, the absence of a duly validated laboratory diagnosis for clinical cases and the lack of a clearly defined medical protocol for patient’s management [178]. The first CP epidemiological studies in French Polynesia date back to the 1960s [154, 172, 179]. From then, epidemiological data were periodically updated through specific studies [16, 17] until 2007 when local health authorities agreed to set up a CP-dedicated epidemiological surveillance program jointly managed by the Health Monitoring Office (Public Health Directorate of French Polynesia) and the Laboratory of Marine Biotoxins (LBM) of Institut Louis Malardé [69]. Initially, the aim of this program was to take advantage of the existing sentinel network of public health physicians to collect declaration forms filled in by each patient suspected of having CP. Over the years, this network progressively extended to the healthcare workers of the 61 public health medical facilities in French Polynesia (including outlying hospitals, medical centers, dispensaries and infirmaries), and even to some private practitioners. Information gathered on this onepage declaration form, which was updated twice since 2007, concern the patient’s age, gender and island of residence, the context of the poisoning (date, fishing area, fish species), the symptoms, the number of previous CP and so on (Figure 10.3). Data collected through this country-wide surveillance program are centralized at LBM-ILM, which produces an annual report widely disseminated through the Public Health Directorate website, medical newsletters/forum, social media, and local media. In addition, a dedicated e-mail address was created to answer to the questions raised by the general public about CP ([email protected]). Since reporting is made on a voluntary basis, CP prevalence is significantly underestimated throughout the country, especially as many patients tend to forgo medical treatment and/or rely on traditional medicine (see Section 10.4.2.2). It is estimated that ciguatera figures in French Polynesia could quite possibly be at least doubled as, in more than half the reported cases, patients stated they had shared the toxic meal with others at table who had also gone ill but failed to report the poisoning [73]. This situation has prompted the implementation of a dedicated ciguatera website (www.ciguatera.pf) in 2014, in an effort to open up the notification program to private practitioners and the general public. The implementation of this community-based participatory reporting program has allowed the production of dynamic risk maps of French Polynesian lagoons, as well as a list of fish species involved in CP incidents in each reporting island. In an effort to increase community outreach, this website also provides a wealth of general information on ciguatera to web users. A main advantage of this IT-based surveillance program is that alerts about the toxic events as well as prevention messages can be issued quickly, since the majority of CP cases are generally notified within 72 h after the poisoning incident.

10 Ciguatera poisoning in the Oceania region

383

French Polynesia Ciguatera and seafood poisoning Surveillance Network DECLARATION FORM PATIENT years

Age

Sex:

F

M

CONTEXT OF POISONING Date of consumption Local name of marine product responsible for the poisoning Flash

Part(s) consumed

Viscera

Head

Eggs

Specify the fishing sopt (Mark with a cross on the map) Region/town

Island bought at the market / store (specify)

bought on the roadside

CLINICAL DATA INTENSITY Mild

Moderate

High

For information: if the patient presents with fever patient and / or allergic reactions and / or skin rash, the diagnosis of ciguatera must be rejected

CARDIOVASCULAR SIGNS Bradycardia Tachycardia Hypotension Hypertension DIGESTIVE SIGNS Nausea Vomiting Diarrhea GENERAL AND NEUROLOGICAL SIGNS Tingling extremities (hands, feet) Touch, neuro-sensitives disturbances Dysesthesia (in contact with cold / hot) Itching Asthenia Headache Dizziness / Balance or walking disorders (underline) Vision disorders Muscular disorders (pain, cramps, weaknesses ...) Joint pain °C Hyportermia : Temperature Burn / tingling of lips, mouth, throat Orofacial pain (teeth, jaw, face) Dysgueusia (taste alteration) Urogenital discomfort/ itching / burning Hallucinations Other symptoms/Observations Time elapsed between the meal and the onset of symptoms

< 30 min

< 12h

< 2h

> 12h

Number of previous CFP/PSP Number of other person(s) also concerned by the poisoning (in addition to the pateint) MEDICAL STRUCTURE IDENTIFICATION Date of consultation Medical structure

Island/Town First aid post

Infirmary

Medical Center

Hospital

Private

Please send the form to LMT - Institut Louis Malarde BP 30 98713 TAHITI I Phone: (689) 40416411 - Fax: (689) 40416406 I eMail: [email protected] You also have the possibility to make a report directly online at www.ciguatera.pf

Figure 10.3: Declaration form filled in by ciguatera patients as part of the CP surveillance program presently conducted in French Polynesia (www.ciguatera.pf). © Institut Louis Malardé.

384

Mireille Chinain et al.

In case of a major outbreak, the response mechanism to CP incidents currently in place in French Polynesia involves three main entities, each having specific mandates: (i) the Health Monitoring Office that issues CP outbreak alerts; (ii) the Hygiene and Public Health Department that sends requests for confirmatory analyses on suspect fish batches and takes all necessary measures to remove toxic fish from retail outlets; and (iii) the LBM-ILM, in charge of conducting confirmatory analyses on suspect fish samples, and field campaigns to investigate the etiology of mass-poisoning outbreaks in newly affected areas.

10.3.2.2 Contribution of herbivores and marine invertebrates to ciguatera poisoning incidents According to Halstead [29], more than 400 fish species are considered to be potential vectors of CP. Historically, CP was thought of as a “carnivore problem” [18, 180], and in fact, Serranidae, Lutjanidae, Carangidae, Muraenidae, Lethrinidae, Sphyraenidae consistently account for the majority of outbreaks worldwide [82, 85, 141, 181–184]. However, it is well established that the transfer of algal CTXs in the food web requires passage through herbivorous fish [185]. This may explain why in CP-prone reef ecosystems, clear shifts in the types of reef fishes involved in CP were generally observed beginning with herbivores, then followed by carnivores years later [18, 128]. 10.3.2.2.1 Herbivorous fish species are major contributors to ciguatera poisoning events in French Polynesia Unlike what is observed in other CP-endemic areas such as the Indian Ocean, the Caribbean and the eastern Atlantic Ocean where CP events rarely involve herbivorous species, in the Pacific, Scaridae, Acanthuridae and Kyphosidae are frequently reported in CP cases occurring in the Cook Islands [18] and French Polynesia [68, 69, 149]. Of note, from 2008 to 2018, herbivorous species consistently ranked among the top five species most frequently involved in CP cases recorded annually in French Polynesia (www.ciguatera.pf), accounting for approximately 14% to 33.6% of cases. Table 10.4 compares the CTX contents found in herbivorous, omnivorous and carnivorous species (either fish or marine invertebrates) in three Pacific Islands: Raivavae (Australes archipelago, French Polynesia), Nuku Hiva (Marquesas archipelago, French Polynesia) and Marakei (Republic of Kiribati). Results confirm that these three islands should be regarded as ciguatera hot spots as significant levels of CTXs were detected in all trophic levels of the food web (Table 10.4). Another important finding is that, in some cases, the amounts of CTXs found in herbivorous species can largely exceed those measured in carnivores, although the highest concentration, that is, 81.84 ppb P-CTX1, was

Trophic level

Family

Marine invertebrates

Herbivores  

Tegulidae (trochus) Toxopneustidae (sea urchin)





 

 

Lutjanidae Lethrinidae Serranidae Carangidae Labridae Mullidae Holocentridae

Carnivores





Siganidae

Omnivores



N

 



Number of species

 

Herbivores

Tridacnidae (giant clam) Scaridae Kyphosidae Acanthuridae Siganidae

Herbivores

Nuku Hiva Island, Marquesas archipelago, French Polynesia

Fish

Marine invertebrates

Raivavae Island, Australes archipelago, French Polynesia

Marine product

CBA-Na

CBA-Na

rRBA

rRBA

rRBA

CBA-Na

Detection methods

.–.

.–.

0.5 day-1. In each of the plots (d), (e) and (f) the light zones (blue) indicate where the second named species would outgrow the first named species; a value of δ growth = zero indicates where neither species exhibited positive net growth. Reproduced from Flynn et al. [97] under a Creative Commons license.

12 Multifaceted climatic change and nutrient effects

483

(Emiliania huxleyi), a motile cryptophyte (Rhodomonas sp.) and a silicifying diatom (Thalassiosira weissflogii), cultures were maintained in present-day pH conditions (8.2), more acidic conditions expected in the future (7.6), and conditions that were more basic, representing of a dense bloom (8.8). Cultures were also compared when the pH was allowed to drift as growth progressed versus when initial conditions were maintained by adding acid or base. Growth rates of Thalassiosira and Emiliania were almost halved in the alkaline cultures where pH was allowed to drift further upward during growth, whereas their growth rates were similar or slightly enhanced in systems that started the drift from more acidic conditions. There was little difference between Rhodomonas grown in drift systems of different initial pH, at least during the nutrient-replete phase. These results were coupled with models revealing potential windows of opportunity for each competing species under different conditions. Emiliania growth was favored under extant pH, with more acidic and basic conditions being unfavorable. Rhodomonas, which had the lowest maximum growth rate, had its highest competitive scope at elevated pH but only when nutrient-replete; these are conditions likely to characterize eutrophic areas during blooms. Interpretations of this include that, when coupled with effects on grazers, oceanic changes in CO2 and eutrophication have potential to impact the frequency of HABs, including blooms of mixoplanktonic species, which combine phototrophy and phagotrophy for their nutrition (see also later). From a biogeochemical perspective, a bloom-induced rise in pH has various other consequences that are often ignored in models, both conceptually and mechanistically, but which may have consequences for bloom success, especially those in shallow waters, such as lakes or estuaries. As pH increases, the fundamental physical– chemical relationships related to P adsorption–desorption change. Enhancement of sediment P release under elevated water column pH conditions has been observed in eutrophic lakes (i.e., [98–101]) and tidal freshwater/oligohaline estuaries [102, 103]. For example, PO43– flux from the sediment in the Potomac Estuary (Maryland, USA) increased from 160 μmol m–2 h–1 when these high pH values were sustained for 100 h [104]. Consequently, the very existence of high-biomass blooms in waters with a legacy of P enrichment effectively creates a self-sustaining flux of P that can prolong the bloom events. Such feedbacks should not be dismissed in modeling and also affect simple interpretations of nutrient ratios of N:P input in coastal ecosystems.

12.3.3 Climate “whiplash” Clearly changes in temperatures, CO2 levels and associated environmental changes are not occurring regionally along the smooth trajectory of increase indicated by

484

Patricia M. Glibert et al.

long-term global average temperatures or oceanic pCO2 changes [105]. The reality instead is that many climate effects are occurring with substantial gyrations at local scales. The term “climate whiplash” is an apt descriptor of many regional phenomena. Seasonality is changing, leading to earlier onset of blooms, or extended durations later in the growing season. Blooms only occur when growth exceeds the potential control by grazers [106]; changes in the timing of prior seasonal events and weather has the potential to shift HAB events. Precipitation patterns are also changing, including increases in droughts in some regions, and/or increased frequency or intensity of storm events in other regions [41, 107]. With warming of the ocean waters and the atmosphere, storms are becoming stronger and wetter. Warmer waters are the accelerant for hurricanes or typhoons, as warmer atmospheres hold more moisture. Stronger and wetter storms have significant implications for nutrient delivery to coastal systems, via both upwelling and coastal runoff. Extreme heat and precipitation events have been on the rise in recent decades and are projected to increase in frequency and intensity in future decades [8]. In an early review of climate change around the globe, it was found that changes in total precipitation were amplified at the tails, and changes not only in average temperature, but also temperature extremes were observed [108]. Considering both historical observations and a large climate model ensemble, the influence of global warming on extreme heat and precipitation events around the world was quantified, and it was found that the historical climate forcing increased the probability of wettest 5-day period by 41% globally [109]. As but one example, in 2017 numerous tropical storms and hurricanes affected the Gulf of Mexico and Caribbean waters, with maximal impact during the fall season in Florida. As another example, in September 2018, Hurricane Florence in the Atlantic dropped ~68 × 109 m3 (18 trillion gallons) of water on the US East Coast, an amount equal to the volume of Chesapeake Bay [110]. Such downpours inevitably wash nutrients (inorganic and organic) off the land. High-resolution Regional Climate Models, embedded into General Circulation Models (GCMs) to obtain more realistic simulations of storms in future climates, tell a similar story [111, 112]. Using an ensemble of GCMs and scenarios from phases 3 and 5 of the Coupled Model Intercomparison Project (CMIP3 and CMIP5), dynamic downscaling projections of twenty-first-century Atlantic hurricane activity have been made [113]. These projections suggest a significant increase in the frequency of intense storms but also hint at a significant reduction in overall tropical storm frequency. Similar results have been indicated by statistical-dynamic models [114–116]. In contrast, under the latest IPCC RCP8.5 scenario [117], tropical cyclones over most oceans are projected to increase not only in intensity but also in frequency during the twenty-first century. Extra-tropical winter storms are also likely to change in the future [118]. The extent to which harmful algae respond to these extreme events will depend on the timing of the event, and the quantity and quality of nutrients discharged from the associated precipitation.

12 Multifaceted climatic change and nutrient effects

485

Droughts, too, are accelerated, and with them come the danger – and reality – of fires. Fires alter the stability of soil, and loss of vegetation during fire can lead to enhance sediment export and runoff. Depending on the terrain, runoff in the year following a fire can increase as much as 30%, and the associated loss of vegetation further exacerbates erosion [119]. Fire also increases the availability of soil nutrients through combustion and accelerated decomposition processes. Moreover, many fire retardants used in fighting fires are NH3 and P-containing compounds. California applied over 5.5 × 107 liters of retardants in 2017, and in 2016, the US Forest Service applied 7 × 107 L on National Forest systems across the country [120]. Australia has applied 15,000 liters of fire retardants for every aircraft pass across the blazing 10.4 million hectares on fire in 2019 and early 2020. This heavy application of fire retardants makes the soils even more nutrient enriched when they eventually become washed to adjacent waters with overland runoff. These phenomena are rarely captured in models of land-use nutrient export.

12.4 Nutrient complexity 12.4.1 Changing nutrient loads As with climate changes, simple dose-response relationships or monotonic trends that characterize many HAB models do not capture the reality of the complexity of how nutrients and their rapid changes are promoting HABs. Nutrient loads have increased substantially in the past several decades and the sources supporting these increased loads are changing with human demography, human diets and the food production supply chain needed to support them. Nutrients are also changing from changing nutrient use efficiencies in crop systems, recycling, inputs from animal wastes (including aquaculture) and changing treatment regimes in wastewater treatment facilities [121–129]. By the year 2050 human population is projected to be nearly 10 billion, a 30% increase over current population estimates, with more than half of this increase expected to occur in Africa [130]. By then also, it has also been suggested that half the human population will reside in urban areas, and that megacities (>10 million people) will continue to be concentrated in China as well as other parts of Asia and Africa, most of which are along coastlines [130]. The concomitant demands for food, fuel, water and sanitation will only continue to rise [131, 132]. To support this increase in population and the required food supply, it is projected that global use of N-based fertilizer, which has increased about 9-fold since 1970, will continue to rise, especially use of urea, which is projected to double by the year 2050 [133, 134]. Use of P-based fertilizers is also predicted to increase, but likely not as steeply as N [125, 135, 136] and there is concern – as well as considerable

486

Patricia M. Glibert et al.

uncertainty – regarding the long-term availability of readily extractable P reserves [137]. Surface waters have reflected these changes, not only in nutrient supplies but also in shifting supply ratios of N relative to P [63, 138, 139]. Management efforts are countering some of these increases. Regional nutrient reductions, the purview of local management decisions, are highly variable. Since the mid-1980s and 1990s the major industrialized nations began curtailing P use by removing it from detergents and by upgrading sewage-treatment processes that generally are more efficient in removing P than N [140–142]. In the US, more states have undertaken P control than N control, and while several states have increased use of both N or P, more have increased use of N relative to P [9]. There is also evidence that China, where massive blooms have occurred in recent decades in both marine and freshwaters [125], is investing substantially in environmental issues, including nutrient-reduction measures. A complicating problem is that nutrients accumulate in landscapes and may be released decades later. For example, N may be temporarily stored in soils and groundwater [128, 143–145]. P may be absorbed by sediments in lakes and reservoirs and be released when concentrations in the water column decline due to P mitigation and with increases in pH (see also above) [146]. The extent to which these legacies contribute to future nutrient loads and changing nutrient ratios, and the extent to which nutrient reductions will be realized in the future, remains in the realm of speculation. Given the aforementioned regional nature of nutrient management versus the global change in patterns of fertilizer use and trends in nutrient reductions, the N:P stoichiometry of runoff is expected to continue to change, leading to upward trends in N:P ratios in many receiving waters [125, 128, 135, 147, 148]. Given an association between elevated cellular ratios of N:P and changes in toxicity of many HAB taxa [139, 149], while noting that it is the concentration and not the ratio of external nutrients that is of importance for phototrophy [150], these issues underscore the importance of multi-stoichiometric modeling approaches.

12.4.2 Spatially explicit models Spatially explicit global maps of dissolved river N and P export and dominant sources have been derived using various models, including the Global Nutrient Export from WaterSheds (Global NEWS) models [151–154], and next-generation models, such as the Integrated Model to Assess the Global Environment-Global Nutrient Model (IMAGE-GNM [127, 155]) and the dynamic version thereof (IMAGE-DGNM [156–158]). Another spatially explicit approach for estimating nutrient export is the SPARROW model that uses statistical relationships to relate water quality monitoring data to upstream sources and watershed characteristics that affect the fate and transport of nutrients [159]. The SPARROW model has been used extensively in the US to estimate nutrient loads to receiving waters such as the Gulf of Mexico [160–162].

12 Multifaceted climatic change and nutrient effects

487

Collectively, these models have documented large differences between regions and over time globally, as well as differences in nutrient export by element. The nutrient source terms that are considered in both the Global NEWS and IMAGE-GNM models include natural sources such as N2 fixation and P weathering, and anthropogenic sources (nonpoint inputs from fertilizer by crop type, N2 fixation by crops, atmospheric N deposition and manure by animal species; point sources from sewage, as estimated by human population and treatment level [151, 163]. These models also account for hydrological and physical factors including water runoff, precipitation intensity, land use and slope as well as in-water removal processes such as dams and reservoirs and consumptive water use, and ultimately, they estimate nutrient export at the river mouths. Applications of these models showed that in the early twenty-first century throughout much of Asia, western Europe, and eastern North America, fertilizer and manure together constitute the largest source of inorganic N (DIN) export [152]; agricultural sources of inorganic N (including fertilizer, animal manures and agricultural N2 fixation) collectively contribute about half of the total DIN exported globally [152]. On the other hand, for dissolved inorganic P (DIP), human sewage is the largest anthropogenic source throughout much of the world, and inorganic P fertilizers and manures are less significant [153]. The aquaculture nutrient model included in IMAGE-GNM shows large and spatially concentrated emissions of nutrients in both freshwater and marine fish cultures in China [129]. IMAGE-DGNM indicates that in river basins with large reservoir volumes, such as the Mississippi, accumulated P in sediments can be de-sorbed and subsequently can contribute to DIP loads [164]. As noted earlier (Section 4.1), increases in pH during bloom development can enhance P release from sediments. Both the NEWS and the IMAGEGNM and IMAGE-DGNM models reflect the same trends in nutrient stoichiometry as those in the fertilizer data, namely, that N:P ratios are rapidly increasing in waters draining to the world’s oceans [127]. Changing nutrient stoichiometry has numerous effects on HABs. In understanding the intersection of effects of changing temperatures and nutrients, applications of key principles of ecological stoichiometry [165], and metabolic ecology [166, 167], have helped in structuring conceptual ideas. These principles focus on associated changes in the traits of growth, respiration, body size and nutrient stoichiometry [168]. Building on these principles, Meunier et al. [169] predicted that increasing N:P ratios in ecosystems should shift communities toward systems with traits including dominance of higher optimal N:P ratios, higher P affinity, decreasing N retention and increasing P storage. These are traits of many HAB species. For mixoplanktonic species, the stoichiometry of its cell, as well as that of its prey, has consequences for growth [170, 171]. Alterations in the composition of nutrient loads have been correlated with shifts from diatom-dominated to flagellated-dominated algal assemblages in many regions. For example, in the Huanghai Sea region of China, inorganic N:P ratios are now about twice Redfield [172] proportions, and about 4-fold higher than in the 1990s [125, 173] and there has been a corresponding nearly 6-fold increase in HAB

488

Patricia M. Glibert et al.

occurrences and a shift to proportionately more dinoflagellates in comparison to diatoms [42, 125]. Similarly, in the South China Sea region, water column inorganic N:P ratios increased from ~2 in the mid-1980s to > 20 in the early 2000s [173]. In addition to the increase in number of HABs, a shift in species composition has led to increasing dominance of genera such as Chattonella, Karenia, and Dinophysis [174]. In the North Sea, as increasing N relative to P proportions developed in response to disproportionate nutrient decreases, mixoplanktonic dinoflagellates have increased, including many harmful species [175]. Previously, the Global NEWS models were used to explore the relationships between the HAB species, Prorocentrum minimum and nutrient export [176]. Areas with high DIN and DIP yields, including eastern Asia, western Europe and eastern North America are also areas in which P. minimum blooms have been documented. This spatially explicit database also allowed identification of relationships between P. minimum occurrence and dominant form of N, as sewage, fertilizer, manure or fixation of atmospheric sources and the extent to which the export of organic forms of N (DON) was attributable to anthropogenic sources. The continued development and application of these global models suggests that they may be applicable to other HAB species for prediction purposes. While Global-NEWS and SPARROW are statistical models that provide snapshots of annual river nutrient export, new generations of models, such as the IMAGE-GNM and IMAGE-DGNM [156–158] and the Riverstrahler model [177, 178] are dynamic and can capture the seasonal dynamics of nutrient biogeochemistry and export to coastal oceans. By simulating the temporal dynamics of N, P, C and Si, these models can be helpful in predicting the occurrence of various types of blooms in coastal marine systems.

12.4.3 Aquaculture as a growing nutrient source One source of rapidly changing anthropogenic nutrients in coastal and freshwater systems, not typically captured in land-use export models, is aquaculture. As wild stocks of finfish and shellfish have continued to decline, aquaculture has become critically important for sustaining global demands [179]. As of 2016, supply of fish and shellfish for human consumption from aquaculture already exceeded those from (declining) wild fish stocks, this supply is projected to increase by at least 60% by 2030. Using data for annual production for different species, countries and types of environment for the period 1950–2010 (derived from the database FISHSTAT of the FAO), a simple nutrient budget model was developed that describes the major flows of nutrients in fish aquaculture systems [129, 180]. The model approach describes nutrients in feed inputs, feed conversion, nutrients in fish and the outflow in the form of feces (solid forms based on apparent digestibility) and dissolved nutrients,

12 Multifaceted climatic change and nutrient effects

489

as well as retention and recycling in pond and integrated aquaculture systems. Based on these models, global cultured production of finfish and crustacean contributed an estimated 1.7 million tonnes of N and 0.46 million tonnes of P to receiving waters during 2008 [181]. Within the relatively short period from 2000 to 2006, nutrient release from shellfish cultures increased by 2.5- to 3-fold, and much larger increases are predicted in nutrient contributions from shellfish cultures by 2050 [182]. While globally these values are small compared to riverine input, on a regional basis they can be substantial [183]. In particular, where marine aquaculture is confined to relatively small areas in coastal seas, primarily in parts of Asia, the amount of nutrients locally released by aquaculture into these seas can be comparable to the nutrient input by rivers [183], for example in Chinese coastal waters [129]. The need for improved models of aquaculture and aquaculture-related effects will only increase as aquaculture continues to expand. The implications of the expansion in aquaculture with regard to HABs are several [183]. Excretory release products and their forms, and proportions under different mariculture conditions may stimulate growth of different protist plankton (phytoplankton and mixoplankton) functional groups [60, 107]. Of particular concern in both shellfish and finfish mariculture is the change in form of nutrients as a result of excretion and associated microbial remineralization. About 7 to 12% of the dissolved N waste of finfish consists of urea, the remainder being NH4+ [184]. As developed more thoroughly later, there is mounting evidence that chemically reduced N forms (in contrast to oxidized N forms), including urea, differentially stimulate the growth of some HAB species [55, 124, 125, 185], may differentially disfavor those competitor species that may depend on oxidized forms of N to a greater degree [55], or may fuel the production of plankton on which mixoplanktonic HABs may feed [17, 107].

12.4.4 Organic nutrients and brownification Sources of organic nutrients are changing in multiple ways. In addition to the increased use of organic N forms as fertilizer [124, 125], and as aquaculture sources, dissolved organic matter (DOM) loading are increasing with changes in precipitation, riverine discharge, sea level rise, altered land use and the associated loss of many wetlands. With coastal inundation due to sea level rise, coastal wetlands and mangrove swamps become stressed and migrate inland. In addition, stronger storms result in increased turbulence that increases scour near roots and trunks of nearshore vegetation, accelerating dieback. When this happens, organic matter and nutrient delivery is increased and a cascade of responses begins that further accelerates erosion, leading to increasing nutrient delivery from land-based sources [186, 187]. Losses of marshes (including mangroves) accelerate nutrient supplies because the marshes are no longer available to take up nutrients and their decay and decomposition increases the supply of nutrients as organic matter decomposes [188]. Long-term (~10–30 year)

490

Patricia M. Glibert et al.

data records of riverine DOC have documented a significant recent increase in DOM concentrations in European [189], Scandinavian [190–192], United Kingdom [193, 194] and North American rivers [195, 196]. This DOM includes dissolved humic substances that are naturally occurring, colored, high molecular weight organic acids. Increases in humic acids are resulting in the phenomenon of brownification of lakes and estuaries, which is having profound effects on food webs and the propensity for blooms. Comparatively few models are addressing these changes, especially considering HABs. Among the modeling frameworks that are addressing marsh production and accretion is the Delft3D modeling system, including one of its modules, the Delft3DWAQ (water quality) model [197, 198]; however, it should be noted that this model system does not currently include any mixoplanktonic HAB species or grazers. Although classically considered to be biologically refractory in marine environments, humic acids have been shown to bioavailable and used in growth of some algal species such as K. brevis [199, 200]. Summer may be a particularly important time for bioavailable DOM transport to estuaries, even though river discharge and DOM fluxes are low [201]. Riverine DOM bioavailability is typically higher in the summer than during the winter [202, 203], suggesting that its impact on the microbial community –including HABs – and contribution to eutrophication may be greater during this period. The largest fraction of the N in humic material is NH4+, loosely bound, and extractable using solid phase resins [204]. NH4+ can be released from humic acids via photooxidation [205] and by ion exchange reactions, especially as salinities increase (as would occur with sea level rise and coastal inundation). Accordingly, humic acids have been hypothesized to act as a “shuttle” for NH4+ in estuaries, binding NH4+ at lower salinities and releasing it at higher salinities [204]. Humic acids also decrease light penetration and can act as chelating agents that make Fe more available. Mixotrophic harmful algae would have an advantage in waters with such low light conditions, as they can augment energy and C inputs through feeding. Highly buoyant blooms of cyanobacteria would also be favored, both in low light and with higher concentrations of NH4+ and warming waters [55]. Recent work has shown that the brownification of freshwaters synergistically interacts with warming temperatures to enhance cyanobacterial blooms and their toxin [206]. Using a factorial experiment with different temperatures and humic-enrichment levels, Urrutia-Cordero et al. [206] found that temperature effects on Microcystis and its toxin levels were amplified in the brownification treatment. Either more toxins were produced per cell, or more toxin strains were selected for in the combined humic enrichment + higher temperature treatments. These findings were supported by analysis and modeling of lake data, which revealed more toxic blooms under conditions of warm temperatures and high humic acid loads. Interestingly, food web manipulation of the lake decreased these impacts, suggesting that complex food web interactions and related nutrient cycling effects are highly intertwined with these effects, but also perhaps offering potential options for managers to keep blooms under control.

12 Multifaceted climatic change and nutrient effects

491

12.4.5 Mixotrophy Increases in all forms of dissolved organic nutrients provide substrates for osmotrophs, which includes all cyanobacteria as well as protists [24]. There is thus scope for growth of many HAB species, and indeed all microbes, to be promoted by increased dissolved organics [207, 208]. While the role of DOM in the growth or such species has long been appreciated, albeit at much lower level of importance compared with phototrophy, the science of HABs and plankton ecology in general, needs a paradigm shift to emphasize the fact that phototrophy + phagotrophy is common in protists [209, 210]. To differentiate mixotrophy as performed by cyanobacteria and diatoms (i.e., phototrophy + osmotrophy) from that involving phagotrophy, the term mixoplankton has been proposed [24]. Mixoplankton are ubiquitous in the world’s oceans (Figure 12.2).

Figure 12.2: Maps showing the distribution of different types of mixoplankton common to HABs. Panel a shows the plastidic specialists that acquire their chloroplast from specific prey, as for example Mesodinium, while panel b shows endosymbiotic mixoplankton that function like greenhouses with phototrophic species while also feeding on other organisms, as for example Noctiluca. Reproduced from Leles et al. [210] under a Creative Commons license.

492

Patricia M. Glibert et al.

In the context of HABs, there is now considerable interest in identifying and quantifying who eats whom and under what conditions [23, 208]. Many harmful algal taxa (except cyanobacteria and diatoms) are mixoplankton [85, 208], many of which feed on their autotrophic competitors, and many of which themselves become prey for other mixoplankton [36, 211]. Mixoplankton range in size from the very smallest flagellated protists to the largest protists [24]. Their prey, too, ranges in size from bacteria and small cells such as Synechococcus [212, 213], to cells larger than the mixotrophs themselves. Although science has known of mixoplankton for several decades, as for example, the discovery years ago that a freshwater synurophycean, the colonial Dinobryon, voraciously consumes bacteria [214], only recently have the pieces of information come together both to conceptualize and to formulate model structures for mixoplankton [22, 215, 216]. These mixotrophic processes substantially alter growth rates [211, 213, 217], have substantial impacts on food webs and are fundamental for understanding the ecology of many types of HABs [23]. The potential for obtaining nutrients via phagotrophy in mixoplankton species, especially as inorganic nutrient supply changes, must be recognized in models for realistic predictions to be achieved [170, 171]. Mixoplanktonic HABs often thrive in conditions where simultaneous acquisition of light and nutrients become challenging. With light from above, but major nutrient sources typically from below, mixotrophs have been said to perform the grand écart (dancer’s splits), by combining two normally incompatible tasks – a difficult accomplishment at best [218]. Some mixoplanktonic HABs accumulate in deep chlorophyll maxima. Off the French coast, for example, a thin layer of dinoflagellates, including the HAB species Dinophysis cf. acuminata, has been observed in the region of the thermocline [219]. The same has been reported for Dinophysis norvegica in the Baltic Sea, where a 1–2-m-thick layer with up to 80,000 cells L–1 is usually situated between 20 and 25 m depth where light is < 1% of light at the surface [220]. Mixoplankton are not necessarily limited to such deep layers, however, as they can graze prey in the water around them [208]. Mixotrophy in these plankton is not simply additive or substitutional with photosynthesis; rather, it is synergistic [221]. Thus, through mixotrophy, there is a provision of additional nutrients (N, P, Fe) from feeding to support primary production, together with a contribution of C to supplement photosynthesis under conditions of light limitation (including night). Moreover, the ecophysiology (nutritional quality) of the component organisms present in the ecosystem affects not only their own growth potential but also the activities of others; the cellular composition of algae, for instance, has consequences for grazers and their emergent properties, with potential for a positive feedback loop generating ungrazable primary producers [2, 222]. In some instances, toxins are used in the capture of prey and thus are directly related to the mixotrophic capacity of the organisms [149, 223, 224]. Modeling of mixoplanktonic activity is advancing as knowledge of this mode of nutrition in HABs increases. The importance of the functioning of predator-prey

12 Multifaceted climatic change and nutrient effects

493

dynamics in HAB formation has long been appreciated by simulation modelers [222], That appreciation is greater now in the light of the mixoplankton paradigm [24], especially in the context of specialist nonconstitutive mixoplankton, such as Dinophysis, that depend on a series of specific trophic interactions to provide the plastids for their growth [216]. Inclusion of mixoplankton within a simple N-phytoplankton-zooplankton-bacteria-detritus (NPZBD) model results in a substantial difference in planktonic trophic dynamics [38, 225] (Figure 12.3). This is because food web dynamics are not merely a summation of a series of rate processes (and kinetic curves); they are an outcome of both the quantity and quality of the substrate (or food) provided. As yet, a sufficient understanding is lacking to simulate how changes in nutrient conditions and climatic events may impact on these processes in nature, although progress at the physiological level is advancing. Recent modeling, based on experimental data on the mixoplankter Karlodinium veneficum and its prey, Rhodomonas, grown individually under varying temperature, autotrophic conditions and in mixed batch cultures in which N:P stoichiometry (molar N:P of 4, 16 and 32) of both predator and prey, illustrate these variable dynamics. Biomass of K. veneficum was highest when it was parameterized as a mixoplankter, and when it consumed prey under the highest N:P conditions considered [171, 227]. The modeled scenarios under all N:P conditions showed large differences between mixoplanktonic and autotrophic growth, but these differences varied with temperature. When inorganic nutrients were in balanced proportions, lower biomass of the mixoplankton was attained at all temperatures in the simulations, suggesting that natural systems might be more resilient against Karlodinium HAB development in warming temperatures if nutrients were available in balanced proportions. The complexities shown in this model, regarding differential impacts of temperature on the growth of mixoplankton predator and prey, parallel those expected under (de) eutrophication scenarios with ocean acidification [97]. Using N-based models, patterns of niche separation of five types of plankton: phytoplankton, protozooplankton and three different types of mixoplankton were described and quantified based on different nutrient, prey and light conditions [38]. Using such a modeling approach, and adjusting resource conditions, makes it possible to simulate these protist groupings under varying conditions. Fully understanding – and thus predicting – changes in the competitive advantage of mixoplanktonic species under multistressor environments (light, temperature, pH, nutrients, salinity) will require a concerted effort in physiology-modeling research in order to aid ecosystem management.

494

Patricia M. Glibert et al.

Nutrients

N

P

B

N

Mixotroph model

Traditional model

M

(b)

(a)

D Z

P

Nutrients

μZ

B D

mZ Light

PO43–

Plasma membrane Autotrophy

(c) +

NH4

Inorganic nutrient uptake

PC Recycle

ChlC

Respiration or regeneration

Recycle

mC

Heterotrophy

Respiration or regeneration

FChIC Assimilation

FNC

FC

FPC

Ingestion

(C-fixation) Photosynthesis

NC

Assimilation

Inorganic nutrient uptake

NO3–

Algal prey

Digestion Void

Figure 12.3: (Panel a) Schematic of a classic model of nutrients (N)-phytoplankton (P) -bacteria (B) -detritus (D) and zooplankton (Z); (Panel b) schematic of how this model changes with inclusion of mixotrophs (M) and micro and microzooplankton classes (mZ, mZ); (Panel c) Schematic of the structure of the mixotrophy model, showing major flows in and out of state variables (solid arrows and boxes) from the external parameters (NO3-, PO43-, and Light), and the major feedback processes [226] (dashed arrows). Autotrophic growth uses inorganic nutrients and light via the photosystems of the mixotroph (phototrophy; white part). A proportion of activity leading to growth is required to support synthesis of those photosystems. Predation brings algal prey into the food vacuole within the confines of the mixotroph cell (heterotrophy; gray part). Interactions between phototrophic and heterotrophic nutrition (Int1) influence the growth of the mixotrophy. The state variables (yellow boxes) that describe carbon (C), nitrogen (N), phosphorus (P), and chlorophyll (Chl) associated with core mixotroph biomass are mC (C-biomass of the mixotroph), ChlC (chlorophyll C quota), NC (cellular NC quota), and PC (cellular PC quota), while the same constituents (green boxes) associated with the content of food vacuole are FC (food vacuole C content relative to mC), FChlC (food vacuole Chl content relative to mC), FNC (food vacuole N content relative to mC), and FPC (food vacuole P content relative to mC). Panels a and b reproduced from Mitra et al. [225] and panel c from Lin et al. [227] under Creative Commons licenses.

12 Multifaceted climatic change and nutrient effects

495

12.5 Ecological complexity HABs do not grow in isolation, away from competitors, grazers or microbial associations – nor, as emphasized earlier, are many exclusively simple vegetative organisms (cyanobacteria and diatoms being the exceptions). In systems most impacted by eutrophication, a close coupling between life history stages of organisms is likely. This is important not only for HABs species but also for other components of the system [2]. Nutrient changes (either direct from nutrient availability, or indirect via trophic interactions) can stimulate the formation of temporary or long-term resting stages (e.g., encystment, and/or excystment) [228]. Germination and subsequent migration of organisms provides a route for the transference of nutrients as from sediments into the water column. This is a general issue not just for HABs, but for other species in the system that may, or may not, interact directly with the HAB species.

12.5.1 Vertical migrations and other adaptations Many HAB species are flagellates that can swim or can exhibit vertical migration behavior through other mechanisms such as buoyancy regulation. Some cyanobacterial species regulate their vertical position in the water column by synthesis and collapse of gas vesicles inside their cells [52]. Flagellates often move into deeper, more nutrient-rich waters, especially toward the microbially active pycnocline or the sediment-water interface; some species move into the lower water column and even surficial sediments, attracted by higher nutrient concentrations [229, 230]. Under certain environmental conditions, their movement can result in the formation of high-density patches [231–233]. Vertical movement by cells in a stratified environment may help to maximize encounter frequencies for sexual reproduction (for eukaryotic algae), and allow cells to obtain nutrients at depth and light at the surface. For mixoplankton, it also enables them to better encounter prey. These and other changes in behavior have the potential to promote synergistic feedbacks in trophic dynamics [2]. For example, diel vertical migration in response to light-nutrient gradients may drive cells to the surface at high densities due to self-shading [234, 235]; this would be exacerbated by advective processes, promoting sexual reproduction and transmission of viruses and pathogens. Dense accumulations formed by such behaviors would also shade competitive phototrophs, though such arguments take on a different slant if one considers mixoplanktonic activity; competitors equate to potential food. The development of prey rejection through nutrient-stress via promotion of anti-grazing activity has been suggested to offer a route for the establishment of an ungrazed HAB [236]. All of these processes are important because they affect the removal and

496

Patricia M. Glibert et al.

transformations of elements (nutrients) for HAB development, as well as predatorprey interactions – but they are often difficult to incorporate in models.

12.5.2 Biogeochemistry Biogeochemical processes, driven largely by bacteria, are also important influences on HABs through processes such as nitrification and denitrification [2, 237]. These processes link chemical interactions between the water column and sediments that can strongly affect HAB initiation, maintenance and dissipation, but they are often poorly characterized in modeling efforts, even in freshwaters where cyanobacterial blooms have been modeled for decades [238]. Typically, such processes are described empirically, but the drivers for these processes require closer association with those affecting HABs. For example, both bacterial production and respiration can be negatively affected by alkaline pH resulting from high rates of photosynthesis from high-biomass blooms mentioned earlier which, in turn, affects C cycling and energy flow and decreases rates of remineralization [239]. The bacteria Nitrosomonas and Nitrobacter are inhibited by NH3 and their inhibition decreases nitrification. Without nitrification, elevated NH4+ and NH3 are sustained [240, 241]. These conditions both favor harmful algal species with a preference for chemically reduced forms of N, such as cyanobacteria and dinoflagellates, while disfavoring taxa (mainly diatoms) that use NO3- preferentially at least under some conditions [55]. Beyond the empirical description of generic benthic biogeochemical processes, various trophic and life history interactions such as bioturbation, as well as abiotic events such as sediment disturbance from storms or wave action may promote fluxes of nutrients to/from the benthos, and also changes in nutrient form. Because benthic environments are more complex and more difficult to characterize compared to the water column, much less emphasis has been placed on modeling benthic processes compared with pelagic. Nevertheless, some models have attempted to include a range of processes required to capture benthic influences on HABs [242, 243]. Progress is being made in developing biogeochemical models that consider HABs. Models such as ERSEM-NEMO (European Regional Seas Ecosystem Model coupled to the Nucleus for European Modelling of the Ocean [237, 244]), ROMS-RCA (Regional Ocean Modeling System coupled to a Row-Column-Aesop model [245, 246]) and CoSINE (Carbon, Silicate, Nitrogen Ecosystem coupled to a ROMS model [247]) among others, are biogeochemical models allowing for multiple phytoplankton groups, particulate and dissolved nutrients (as C, N, P, Si) and which include biogeochemical process of nitrification, denitrification, interactions with sediment, both an aerobic and an anaerobic layer. DELFT3D-GEM model is another generic ecological model that can be applied to different kinds of water systems to simulate the primary production and phytoplankton composition [206, 248, 249]. An earlier assessment with ERSEM and a predecessor of NEMO was successfully used to

12 Multifaceted climatic change and nutrient effects

497

simulate the impact of nutrients and climate change on the proliferation of HABs in NW Europe, NE and SE Asia [45]. These models are largely open-source, providing modelers with a framework on which HAB dynamics can be layered. It should be noted that the Earth Systems Models, which are used for studying impacts of climate change, for the plankton food web structure typically employ simplified NPZ-style models often with only a single nutrient (N, thus ignoring P). Typically, phytoplankton groups are characterized as “small” versus “large,” or “diatom” versus “nondiatom” or at most three phytoplankton functional types [250]. In such models, the HAB species can be included by applying a rhomboid strategy (characterizing a single species against a community background). Such approaches are advancing. For example, spatially explicit mechanistic models of the temporal and spatial dynamics of the HAB dinoflagellates Prorocentrum minimum and Karlodinium veneficum in Chesapeake Bay have been parameterized using this approach [251] and validated against long-term water quality and plankton measurements. In the case of K. veneficum in Chesapeake Bay, both prey availability and differential stoichiometric conditions of prey and of the mixotroph were shown to impact the timing and spatial extent of the resulting blooms [252] (Figure 12.4). Another example includes incorporation of the Perfect Beast mixoplankton model [226] into the ERSEM structure to represent all mixoplankton functional groups allometrically and functionally [209]. This ecosystem model when validated against the long-term L4 data series shows the importance of mixoplankton in these coastal waters [253].

12.5.3 Microbial interactions An area about which much is yet to be learned (and about much work remains to be done in terms of modeling) is the microbial associations of HABs. It has been proposed that perturbations in factors such as nutrients and light alter the microbial loop so as to create “loopholes” that enable the harmful algal species to escape predation pressure and form blooms [106, 255]. Based on data for the biomass of phytoplankton and microzooplankton (ciliates and heterotrophic dinoflagellates) from 12 geographic regions, it was hypothesized that HABs can escape microzooplankton control at bloom initiation through predation avoidance mechanisms, such as by their larger size, colonies, spines or toxins [106]. It has also been suggested that HABs can develop through a “self-propagating failure of normal predator-prey activity, resulting in the transfer of nutrients into HAB growth at the expense of competing algal species” [222]. The rate limitation of this nutrient transfer provides continual nutrient stress, which results in various grazing deterrent behaviors by the harmful species, protecting them from grazing control. This process may be self-stabilizing as long as nutrient demand exceeds supply, which would be most likely under eutrophic conditions with skewed nutrient ratios.

498

Patricia M. Glibert et al.

(a)

(b)

39.5° N

1 × 106 cells L–1 1 × 107 cells L–1 3 × 107 cells L–1

Susquehanna R.

Atla ntic Oce an

39.0°

CBS.2 input Freshwater

38.5°

N:P

N:P

Seawater input

Karlodinium veneficum 38.0°

Spring freshet

37.5°

37.0°

N:P

(c)

Summer dinoflagellate bloom N:P

N-rich nutrient N-rich prey N-deficient predator

×106 4 39.5

3.5

39

2.5 2

38

1.5

37.5

Karlodinium (cells L–1)

3

38.5

Latitude

Salinity 1-3 month lags N inputs No lag N-rich prey

1

37

0.5

36.5

0 –77

–76.5

–76

Karlodinium (cells/L)

Figure 12.4: (Panel a) Spatial distribution of the harmful dinoflagellate Karlodinium veneficum in Chesapeake Bay for all recorded observations exceeding a cell density of 1 × 106 cells L–1 for the years 2003–2008. Reproduced and modified from Li et al [254] with permission of Elsevier. (Panel b) Conceptual diagram of links between flow, dissolved N and P availability and nutritional condition of prey and of K. veneficum in summer in the mesohaline zone of Chesapeake Bay. These conditions are suggested to lead to late-summer blooms of K. veneficum. Reproduced from Lin et al. [252] under a Creative Commons license. (Panel c) modeled spatial distribution of K. veneficum for a summer condition in Chesapeake Bay based on a fully coupled ROMS-RCAmixotrophy model (see text for details).

Microbial interactions play a critical role in controlling plankton diversity and community dynamics in general [256, 257]. Bacteria in particular have been shown to play major roles in providing limiting and regenerated nutrients [258–260] that, in turn, alter growth [261–263]. The mixoplankton paradigm sees a different, greater level of microbial connectivity than was assumed under the classic phytoplankton paradigm [22, 24]. Bacteria have been proposed to be farmed by mixoplankton [225]

12 Multifaceted climatic change and nutrient effects

499

as a route to obtain essential nutrients such as Fe and P. Alternatively, they may lyse phytoplankton cells, including HAB species [262, 264, 265]. Viruses, the most abundant biological entities in the ocean [266], facilitate the movement of nutrients from organisms to pools of dissolved and nonliving particulate organic matter, a process termed the viral shunt [267]. The viral shunt affects microbial turnover rates and, hence, energy and material fluxes. Viral growth also has profound effects on microbial population dynamics [268, 269]. As an example, while scant data are available on bacterial and viral interactions with K. brevis, the major red tide species of the Gulf of Mexico, there have been some lessons learned. Based on a modeling effort, nutrient limitation of K. brevis appeared to be a precondition for K. brevis susceptibility to bacterial and viral attack, and potentially programed cell death, indicating microbial interactions can play a critical role in bloom termination [16]. Viruses have been linked to lysis of K. brevis cultures [270, 271]. In one study, virus production and lysis of nonaxenic K. brevis cultures was shown following the addition of virus-containing  ×  cells mL– or ≥  mg L– biomass (AL ) >  ×  cells mL– or ≥  mg L– biomass (AL )

Cyanobacteria

Visual inspection weekly

15 From science to policy

619

Table 15.3 (continued ) Drinking water Toxin concentration in finished drinking water unless otherwise stated Cyanobacteria concentration in water reservoir Country

Toxin legislated Maximum limit Recommendation

Minimum Frequency

France

MCs

 µg L–

Germany (no specific regulations)

MCs CYL

 µg L– . µg L–

Hungary (no specific regulations)

Cyanobacteria

−

At least once a year

Italy (no specific regulations)

Algae

−

If presumed risk

Netherland (no specific regulations)

MCs Cyanobacteria

 µg L– < . ×  cells mL–

New Zealand

MCs NOD CYL ANAT-a HomoANAT-a ANAT-a(s) STX

 µg L– MC-LR eq.  µg L–  µg L–  µg L–  µg L–  µg L–  µg L– STX eq.

Fortnight if area susceptible of cyanobacteria blooms

Singapore

MC-LR

 µg L–

Annual

Only when cyanobacteria proliferate

–

South Africa

MC-LR

 µg L

Spain

MCs

 µg L–

Only under suspicion of eutrophization

Turkey (proposed legislation)

MCs Cyanobacteria

 µg L– MC-LR eq. >  ×  cells mL– >  µg L– chlorophyll-a.

Cyanobacterial populations monthly

Uruguay

MC-LR

 µg L–

USA (no federal regulation, health advisories)

MCs

. µg L– MC-LR eq. for children (