&EPA
United States Office of Water EPA-820R15102
Environmental Mail Code 4304T June 2015
Protection Agency
Health Effects Support Document
for the Cyanobacterial Toxin
Microcystins
-------
Health Effects Support Document
for the Cyanobacterial Toxin
Microcystins
U.S. Environmental Protection Agency
Office of Water (4304T)
Health and Ecological Criteria Division
Washington, DC 20460
EPA Document Number: 820R15102
Date: June 15, 2015
Health Effects Support Document for Microcystins - June, 2015
i
-------
FOREWORD
The Safe Drinking Water Act (SDWA), as amended in 1996, requires the Administrator of the U.S.
Environmental Protection Agency (EPA) to establish a list of unregulated microbiological and chemical
contaminants that are known or anticipated to occur in public water systems and that may need to be
controlled with a national primary drinking water regulation. The SDWA also requires that the Agency
make regulatory determinations on at least five contaminants on the list every five years. For each
contaminant on the Contaminant Candidate List (CCL), the Agency will need to obtain sufficient data to
conduct analyses on the extent of occurrence and the risk posed to populations via drinking water.
Ultimately, this information will assist the Agency in determining the appropriate course of action (e.g.,
develop a regulation, develop guidance or make a decision not to regulate the contaminant in drinking
water).
This document presents information, including occurrence, toxicology and epidemiology data, for the
cyanobacterial toxins microcystins to be considered in the development of a Drinking Water Health
Advisory (DWHA). DWHAs serve as the informal technical guidance for unregulated drinking water
contaminants to assist federal, state and local officials, and managers of public or community water
systems in protecting public health as needed. They are not to be construed as legally enforceable federal
standards.
To develop the Health Effects Support Document (HESD) for microcystins, a comprehensive literature
search was conducted from January 2013 to May 2014 using Toxicology Literature Online (TOXLINE),
PubMed component and Google Scholar to ensure the most recent published information on microcystins
was included. The literature search included the following terms: microcystin, microcystin congeners,
congeners, human toxicity, animal toxicity, in vitro toxicity, in vivo toxicity, occurrence, environmental
fate, mobility and persistence. EPA assembled available information on: occurrence; environmental fate;
mechanisms of toxicity; acute, short term, subchronic and chronic toxicity and cancer in humans and
animals; toxicokinetics and exposure.
Additionally, EPA relied on information from the following risk assessments in the development of the
HESD for microcystin.
• Health Canada (2012) Toxicity Profile for Cyanobacterial Toxins
• Enzo Funari and Emanuela Testai (2008) Human Health Risk Assessment Related to Cyanotoxins
Exposure
• Tai Nguyen Duy, Paul Lam, Glen Shaw and Des Connell (2000) Toxicology and Risk
Assessment of Freshwater Cyanobacterial (Blue-Green Algal) Toxins in Water
A Reference Dose (RfD) determination assumes that thresholds exist for certain toxic effects, such as
cellular necrosis, significant body or organ weight changes, blood disorders, etc. It is expressed in terms
of milligrams per kilogram per day (mg/kg/day) or micrograms per kilogram per day ((.ig/kg/day). In
general, the RfD is an estimate (with uncertainty spanning perhaps an order of magnitude) of a daily oral
exposure to the human population (including sensitive subgroups) that is likely to be without an
appreciable risk of deleterious effects during a lifetime.
The carcinogenicity assessment for microcystins includes a formal hazard identification and an estimate
of tumorigenic potency if applicable. Hazard identification is a weight-of-evidence judgment of the
likelihood that the agent is a human carcinogen via the oral route and of the conditions under which the
carcinogenic effects may be expressed.
Health Effects Support Document for Microcystins - June, 2015
-------
Development of this hazard identification and dose-response assessment for microcystins has followed
the general guidelines for risk assessment as set forth by the National Research Council (1983) the EPA's
(2014b) Framework for Human Health Risk Assessment to Inform Decision Making. EPA guidelines used
in the development of this assessment include the following:
• Guidelines for the Health Risk Assessment of Chemical Mixtures (U.S. EPA, 1986a)
• Guidelines for Mutagenicity Risk Assessment (U.S. EPA, 1986b)
• Recommendations for and Documentation of Biological Values for Use in Risk Assessment (U.S.
EPA, 1988)
• Guidelines for Developmental Toxicity Risk Assessment (U .S. EPA, 1991)
• Interim Policy for Particle Size and Limit Concentration Issues in Inhalation Toxicity Studies
(U.S. EPA, 1994a)
• Methods for Derivation ofInhalation Reference Concentrations and Application of Inhalation
Dosimetry (U.S. EPA, 1994b)
• Use of the Benchmark Dose Approach in Health Risk Assessment (U.S. EPA, 1995)
• Guidelines for Reproductive Toxicity Risk Assessment (U .S. EPA, 1996)
• Guidelines for Neurotoxicity Risk Assessment (U.S. EPA, 1998)
• Science Policy Council Handbook: Peer Review (2nd edition) (U.S. EPA, 2000a)
• Supplemental Guidance for Conducting Health Risk Assessment of Chemical Mixtures (U.S.
EPA, 2000b)
• A Review of the Reference Dose and Reference Concentration Processes (U.S. EPA, 2002)
• Guidelines for Carcinogen Risk Assessment (U.S. EPA, 2005a)
• Supplemental Guidance for Assessing Susceptibility from Early-Life Exposure to Carcinogens
(U.S. EPA, 2005b)
• Science Policy Council Handbook: Peer Review (U.S. EPA, 2006a)
• A Framework for Assessing Health Risks of Environmental Exposures to Children (U.S. EPA,
2006b)
• Highlights of the Exposure Factors Handbook (U.S. EPA, 2011)
• Benchmark Dose Technical Guidance Document (U.S. EPA, 2012)
• Child-Specific Exposure Scenarios Examples (U.S. EPA, 2014a)
• Framework for Human Health Risk Assessment to Inform Decision Making (U.S. EPA, 2014b)
Health Effects Support Document for Microcystins - June, 2015
-------
AUTHORS, CONTRIBUTORS AND REVIEWERS
Authors
Lesley V. D'Anglada, Dr.P.H. (Lead)
Joyce M. Donohue, Ph.D.
Jamie Strong, Ph.D.
Office of Water, Office of Science and Technology
Health and Ecological Criteria Division
U.S. Environmental Protection Agency, Washington D.C.
Belinda Hawkins, Ph.D., DABT
Office of Research and Development, National Center for Environmental Assessment
U.S. Environmental Protection Agency, Cincinnati, OH
The following contractor authors supported the development of this document:
Anthony Q. Armstrong, M.S.
Carol S. Wood, Ph.D., DABT
Oak Ridge National Laboratory, Oak Ridge, TN
The Oak Ridge National Laboratory is managed and operated by UT-Battelle, LLC. for the U.S.
Department of Energy under Contract No. DE-AC05-00OR22725.
The following contractor authors developed earlier unpublished drafts that contributed
significantly to this document:
Carrie Fleming, Ph.D. (former Oak Ridge Institute for Science and Education participant)
Oak Ridge National Laboratory, Oak Ridge, TN
Heather Carlson-Lynch, S.M., DABT
Julie Melia (Stickney), Ph.D., DABT
Marc Odin, M.S., DABT
SRC, Inc., North Syracuse, NY
Robyn Blain, Ph.D.
Audrey Ichida, Ph.D.
Kaedra Jones, MPH
William Mendez, Ph.D.
Pamela Ross, MPH
ICF International, Fairfax, VA
Health Effects Support Document for Microcystins - June, 2015
iv
-------
Reviewers
Internal Reviewers
Cesar Cordero, M.S.
Kenneth Rotert, M.S.
Meredith Russell, M.S.
Melissa Simic, M.S.
Neil Chernoff, Ph.D.
Armah A. de la Cruz, Ph.D.
Sally Perreault Darney, Ph.D.
Elizabeth Hilborn, DVM, MPH, DACVPM
Nicole Shao, M.S.
External Reviewers
Lorraine Backer, Ph.D., MPH
Wayne W. Carmichael, Ph.D.
Richard Charron, M.S.
Karen Chou, Ph.D.
Enzo Funari, Ph.D.
Michele Giddings, B.S.
Stephen B. Hooser, DVM, Ph.D., DABVT
Andrew Humpage, Ph.D.
Jeanne M. Manson, Ph.D., MSCE
Ian Stewart, Ph.D.
Donald G. Stump, Ph.D., DABT
Xiaozhong Yu, Ph.D.
Office of Ground Water and Drinking Water, USEPA
Office of Ground Water and Drinking Water, USEPA
Office of Ground Water and Drinking Water, USEPA
Office of Ground Water and Drinking Water, USEPA
Office of Research and Development, USEPA
Office of Research and Development, USEPA
Office of Research and Development, USEPA
Office of Research and Development, USEPA
Office of Research and Development, USEPA
Centers for Disease Control and Prevention
Wright State University
Water and Air Quality Bureau, Health Canada
Michigan State University
Istituto Superiore di Sanita
Water and Air Quality Bureau, Health Canada
Purdue University
University of Adelaide
Private Consultant; retired from ExPonent 2012
South Australian Government's R&D Institute (SARDI)
WIL Research Laboratories, LLC
University of Georgia
Health Effects Support Document for Microcystins - June, 2015
v
-------
TABLE OF CONTENTS
FOREWORD II
AUTHORS, CONTRIBUTORS AND REVIEWERS IV
TABLE OF CONTENTS VI
LIST OF TABLES VIII
LIST OF FIGURES VIII
ABBREVIATIONS AND ACRONYMS IX
EXECUTIVE SUMMARY XII
1.0 IDENTITY: CHEMICAL AND PHYSICAL PROPERTIES 1
1.1 Chemical and Physical Properties 1
1.2 Microcystin Congeners 1
2.0 TOXIN SYNTHESIS AND ENVIRONMENTAL FATE 4
2.1 Cyanotoxin Synthesis 4
2.2 Environmental Factors that Affect the Fate of Cyanotoxins 5
2.2.1 Nutrients 5
2.2.2 Light Intensity 6
2.2.3 Temperature 6
2.2.4 Other Environmental Factors 7
2.3 Environmental Fate of Microcystins 9
2.3.1 Hydrolysis 9
2.3.2 Photolysis 9
2.3.3 Metabolism 9
2.3.4 Transport 9
2.4 Summary 10
3.0 CYANOTOXIN OCCURRENCE AND EXPOSURE IN WATER 11
3.1 General Occurrence of Cyanobacteria in Water 11
3.2 Microcystins Occurrence in Surface Water 11
3.3 Microcystins Occurrence in Drinking Water 14
3.4 Summary 15
4.0 CYANOTOXIN OCCURRENCE IN MEDIA OTHER THAN WATER 17
4.1 Occurrence in Soil and Edible Plants 17
4.2 Occurrence in Fish and Shellfish 17
4.3 Occurrence in Dietary Supplements 22
4.4 Summary 22
5.0 TOXICOKINETICS 23
5.1 Absorption 23
5.1.1 Oral Exposure 23
5.1.2 Inhalation Exposure 23
5.1.3 Dermal Exposure 24
5.2 Distribution 24
5.2.1 Oral Exposure 24
5.2.2 Inhalation Exposure 25
Health Effects Support Document for Microcystins - June, 2015 vi
-------
5.2.3 Other Exposure Routes 25
5.2.4 Liver Tissues - in vitro 26
5.3 Metabolism 28
5.4 Excretion 29
5.5 Pharmacokinetic Considerations 30
6.0 HAZARD IDENTIFICATION 31
6.1 Human Studies 31
6.1.1 Epidemiology and Case Studies of Systemic Effects 31
6.1.2 Other Routes of Exposures 32
6.2 Animal Studies 33
6.2.1 Acute Toxicity 33
6.2.2 Short-Term Studies 37
6.2.3 Subchronic Studies 40
6.2.4 Neurotoxicity 42
6.2.5 Developmental/Reproductive Toxicity 44
6.2.6 Chronic Toxicity 49
6.2.7 Immunotoxicity 51
6.2.8 Hematological Effects 52
6.3 Carcinogenicity 53
6.3.1 Cancer Epidemiology Studies 53
6.3.2 Animal Studies 57
6.4 Other Key Data 59
6.4.1 Mutagenicity and Genotoxicity 59
6.4.2 Physiological or Mechanistic Studies 64
7.0 CHARACTERIZATION 01 RISK 79
7.1 Synthesis and Evaluation of Maj or Noncancer Effects 79
7.1.1 Mode of Action of Noncancer Effects 82
7.1.2 Dose-Response Characterization for Noncancer Effects 83
7.2 Synthesis and Evaluation of Carcinogenic Effects 84
7.2.1 Mode of Action and Implications in Cancer Assessment 85
7.2.2 Weight of Evidence Evaluation for Carcinogenicity 85
7.2.3 Dose Response Characterization for Cancer Effects 86
7.3 Potentially Sensitive Populations 86
7.4 Characterization of Health Risk 87
7.4.1 Choice of Key Study 87
7.4.2 Endpoint Selection 88
7.4.3 RfD Determination 88
8.0 RESEARCH GAPS 91
9.0 REFERENCES 92
Health Effects Support Document for Microcystins - June, 2015 vii
-------
LIST OF TABLES
Table 1-1. Amino Acid Composition of Various Microcystin Congeners (Yuan et al., 1999) 2
Table 1-2. Chemical and Physical Properties of Microcystin-LR 3
Table 3-1. States Surveyed as Part of the 2007 National Lakes Assessment with Water Body Microcystin
Concentrations Above the WHO Advisory Guideline Level for Recreational Water of 10 (ig/L (U.S.
EPA, 2009) 13
Table 4-1. Bioaccumulation Studies of Microcystins in Fish, Shellfish, and Crustaceans 20
Table 6-1. Incidence of Liver Lesions in Mice and Rats After Exposure to Microcystin-LR (Fawell et al.,
1999) 35
Table 6-2. Relative Liver Weights and Serum Enzyme Levels in Rats Ingesting microcystin-LR in
Drinking Water (Heinze, 1999) 38
Table 6-3. Histological Evaluation of the Rat Livers After Ingesting Microcystin-LR in drinking Water
(Heinze, 1999) 38
Table 6-4. Incidence and Severity of Nasal Cavity Lesions in Mice After Inhalation of Microcystin-LR 39
Table 6-5. Serum Biochemistry Results for Mice Treated with Microcystin-LR for 13 Weeks 42
Table 6-6. Liver Histopathology in Male and Female Mice Treated with Microcystin-LR for 13 Weeks. 42
Table 6-7. Serum Hormone Levels and Sperm Analyses From Mice Given Microcystin-LR in the
Drinking Water for 3 or 6 Months 46
Table 6-8. Relative Risk of Colorectal Cancer By Drinking Water Source 55
Table 6-9. Mutagenicity Assays with Microcystins 60
Table 6-10. Genotoxicity of Microcystins In vitro 63
Table 6-11. Genotoxicity of Microcystins In vivo 64
Table 6-12. Protein Phosphatase Inhibition Activity Among Microcystin Congeners 67
Table 7-1. Adverse Effects By Route of Exposure to Microcystins 84
LIST OF FIGURES
Figure 1-1. Structure of Microcystin (Kondo etal., 1992) 2
Figure 1-2. Structure of the Amino Acids Adda and Mdha (Harada et al., 1991) 3
Figure 2-1. Environmental Factors Influencing Cyanobacterial Blooms 8
Figure 6-1. Sites of Surface Water Treatment Service Areas and Control Ground Water Treatment Service
Areas 56
Health Effects Support Document for Microcystins - June, 2015
viii
-------
ABBREVIATIONS AND ACRONYMS
A Alanine
Adda 3-amino-9-methoxy-2, 6, 8,-trimethyl-10-phenyldeca-4, 6-dienoic acid
ADHD Attention deficit hyperactivity disorders
AFA Aphanizomenon flos-aquae
ALDH2 Aldehyde dehydrogenase 2
ALT Alanine aminotransferase
ALP Alkaline phosphatase
AST Aspartate aminotransferase
AWWARF American Water Works Association Research Foundation
BGAS Bluegreen algae supplements
BSO Buthionine sulfoximine
BUN Blood urea nitrogen
BW Body weight
CAS Chemical Abstracts Service
CEGLHH Center of Excellence for Great Lakes and Human Health
CCL Contaminant Candidate List
CHO Chinese hamster ovary
CI Confidence Interval
CYP2E1 Cytochrome P45 0 2E1
CYP450 Cytochrome P450
DMBA Dimethylbenzanthracene
DNA Deoxyribonucleic acid
DW Dry Weight
DWHA Drinking Water Health Advisories
ELISA Enzyme-linked immunosorbent assay
EPA U.S. Environmental Protection Agency
ERK Extracellular signal-regulated protein kinase
ETC Electron transport chain
F Phenylalanine
Fe Iron
FEL Frank effect level
FSH Follicle stimulating hormone
FT3 Free triiodothyronine
FT4 Free thyroxin
g Gram
GD Gestation day
GC/MS Gas chromatograph/mass spectrometry
GFR Glomerular filtration rate
GGT y-Glutamyl transpeptidase
GI Gastrointestinal
GIS Geographical information system
GSH Glutathione
GST Glutathione S-transferase
Health Effects Support Document for Microcystins - June, 2015
ix
-------
GST-P
Glutathione ^-transferase placental form-positive
HA Health advisory
HAB Harmful algal bloom
HEK Human embryonic kidney cells
HPLC High-performance liquid chromatography
IARC International Agency for Research on Cancer
i.p. Intraperitoneal
i.v. Intravenous
JNK c-Jun N-terminal protein kinase
kg Kilogram
L Leucine
LC/MS Liquid chromatography/mass spectrometry
LDH Lactate dehydrogenase
LH Luteinizing hormone
LOAEL Lowest-observed-adverse-effect level
LPS Lipopoly saccharides
MAPK Mitogen-activated protein kinase
MDA Malondialdehyde
Mdha Methyldehydroalanine
MERHAB-LGL Monitoring and Event Response to Harmful Algal Blooms in the Lower
Great Lakes
(.ig Microgram
(.iM Micromole
mg Milligram
mL Milliliter
MMP Metalloproteinase
Mn Manganese
MOA Mode of action
MPT Mitochondrial permeability transition
mRNA Messenger RNA
nm Nanometer
nM Nanomole
N Nitrogen
N/A Not Applicable
NDEA N-nitrosodiethylamine
NLA National Lakes Assessment
NMR Nuclear magnetic resonance
NOAA National Oceanic and Atmospheric Administration
NOAEL No-observed-adverse-effect level
NRC National Research Council
NRPS Nonribosomal peptide synthetase
Health Effects Support Document for Microcystins - June, 2015
x
-------
OATp
Organic acid transporter polypeptides
OR
Odds ratio
OXPHOS
Oxidative phosphorylation
P
Phosphorus
PCE
Polychromatic erythrocyte
PKS
Polyketide synthase
PMN
Polymorphonuclear leukocyte
PP2A
Protein phosphatase 2A
PP1
Protein phosphatase 1
PP4
Protein phosphatase 4
R
Arginine
RfD
Reference dose
RNA
Ribonucleic acid
ROS
Reactive oxygen species
RR
Relative risk
qPCR
Quantitative polymerase chain reaction
SDWA
Safe Drinking Water Act
SE
Standard error
SOD
Superoxide dismutase
SRR
Standardized rate ratios
TEF
Toxicity equivalency factors
TH
Thyroid hormone
TNF-a
Tumor necrosis factor-alpha
Tra
TH receptor
TOXLINE
Toxicology Literature Online
TUNEL
Terminal deoxynucleotidyl transferase-mediated dUTP-biotin nick end-
labeling assay
UF
Uncertainty factor
USGS
United States Geological Survey
uv
Ultraviolet
w
Tryptophan
WHO
World Health Organization
Y
Tyrosine
Health Effects Support Document for Microcystins - June, 2015
xi
-------
EXECUTIVE SUMMARY
Microcystins are toxins produced by various cyanobacterial species, including members of Microcystis,
Anabaena, Nodularia, Planktothrix, Fischerella, Nostoc, Oscillatoria, and Gloeotrichia. Structurally, the
microcystins are monocyclic heptapeptides that contain seven amino acids joined end-to-end and then
head to tail to form cyclic compounds that are comparatively large, (molecular weights ranging from ~
800 to 1,100 g/mole).
Microcystin congeners vary based on their amino acid composition and through methylation or
demethylation at selected sites within the cyclicpeptide. The variations in composition and methylation
account for the large number of toxin congeners (approximately 100). The microcystins are named based
on their variable amino acids. For example, microcystin-LR, the most common congener, contains leucine
(L) and arginine (R). The preponderance of toxicological data on the effects of microcystins is restricted
to the microcystin-LR congener.
Microcystins are the most common cyanotoxins found worldwide and are relatively stable in the
environment as they are resistant to hydrolysis at near neutral pH. In the presence of full sunlight,
photochemical breakdown can occur in as little as two weeks or longer than six weeks, depending on the
microcystin congener. They are susceptible to degradation by aquatic bacteria found naturally in rivers
and reservoirs. In aquatic environments the toxin tends to remain contained within the cyanobacterial cell
and is released in substantial amounts only upon cell lysis. Microcystins have been reported to remain
potent even after boiling. Microcystins may adsorb onto naturally suspended solids and dried crusts of
cyanobacteria and can precipitate out of the water column and reside in sediments for months.
Concentrations associated with blooms in surface waters in the U.S. and Europe typically range from very
low levels (detection limit) and have been measured as high as 150,000 (ig/L.
Drinking water is an important source of potential exposure to cyanotoxins. Exposure to cyanobacteria
and their toxins may also occur by ingestion of toxin-contaminated food, by inhalation and dermal contact
during bathing or showering, and during recreational activities in waterbodies with the toxins. However,
these types of exposures are considered minimal due to various factors including lack of biomagnification
and biodilution via food. Due to the seasonality of cyanobacterial blooms, exposures are usually not
chronic. Symptoms reported after acute recreational exposure to cyanobacterial blooms (including
microcystin-producing genera) included skin irritations, allergic reactions or gastrointestinal illnesses.
Limited data in humans and animals demonstrate that the absorption of microcystins from the intestinal
tract into liver, brain, and other tissues, and the export from the body, requires facilitated transport using
receptors belonging to the organic acid transporter polypeptide (OATp) family. Data in humans and
animals suggests that the liver is a primary site for binding these proteins (i.e., increased liver weight in
laboratory animals and increased levels of serum enzymes in laboratory animals and humans). Once
inside the cell, these toxins covalently bind to cytosolic proteins (PP1 and PP2) resulting in their retention
in the liver. Limited data are available on the metabolism of microcystins, but most of the studies show
that conjugation with glutathione and cysteine increases solubility and facilitates excretion.
Human data on the oral toxicity of MC-LR are limited by lack of quantitative information and by
potential co-exposure to other cyanobacterial toxins and microorganisms. Acute, short-term and
subchronic experimental studies all provide evidence of hepatotoxicity, and chronic studies, that are
limited by lack of evaluation of comprehensive endpoints and comprehensive reporting, support these
findings. Several studies of microcystin-LR reported findings of lesions in the testes and decreased sperm
counts and motility.
EPA estimated a reference dose (RfD) for microcystins of 0.05 (ig/kg/day based on increased liver
weight, slight to moderate liver lesions with necrosis with hemorrhages, and increased enzyme levels in
Health Effects Support Document for Microcystins - June, 2015
xii
-------
rats from the study by Heinze (1999). This study identified a LOAEL of 50 (ig/kg/day, based on these
effects. The drinking water route of exposure and shorter duration of the study (28 days) closely match
potential short-term exposure scenarios that are the focus of a Ten-day health advisory for microcystin.
The composite uncertainty factor includes application of a 10 for intraspecies variability, 10 for
interspecies variability, 3 (10I/2) for converting a LOAEL to a NOAEL, and 3 (1014) for uncertainties in the
database.
Applying the Guidelines for Carcinogen Risk Assessment, there is inadequate evidence to determine the
carcinogenicity of microcystins. The few available epidemiological studies suggest an association
between liver or colorectal cancers and microcystin exposures, but are limited by their ecological study
design, lack of individual exposure measurements, potential co-exposure to other microbial or chemical
contaminants and, in some cases, failure to control for known liver and colorectal risk factors. No long
term animal studies designed to evaluate dose-response for tumorigenicity of microcystin following
lifetime exposures were available. Other studies evaluating the tumor promotion potential of microcystin
following pretreatment with a potent initiator such as NDEA or N-methyl-N-nitroso urea, found an
increase in the number and/or size of GST-P positive foci observed (Nishiwaki-Matsushima et al., 1992;
Ohta et al., 1994; Falconer and Humpage, 1996; Sekijima et al., 1999; Humpage et al., 2000; Ito et al.,
1997b). In two promotion studies, MC-LR alone showed no initiating activity (Nishiwaki-Matsushima et
al., 1992; Ohtaetal., 1994).
Health Effects Support Document for Microcystins - June, 2015
xiii
-------
1.0 IDENTITY: CHEMICAL AND PHYSICAL PROPERTIES
1.1 Chemical and Physical Properties
Cyanobacteria, formerly known as blue-green algae (Cyanophyceae), are a group of bacteria containing
chlorophyll-a that can carry out the light and dark phases of photosynthesis (Castenholz and Waterbury,
1989). In addition to chlorophyll-a, other pigments such as carotene, xanthophyll, blue c phycocyanin and
red c phycoerythrin are also present in cyanobacteria (Duy et al., 2000). Most cyanobacteria are aerobic
photoautotrophs, requiring only water, carbon dioxide, inorganic nutrients and light for survival, but
others have heterotrophic properties and can survive long periods in complete darkness (Fay, 1965). Some
species also are capable of nitrogen fixation (i.e., diazotrophy) (Duy et al., 2000) producing inorganic
nitrogen compounds to synthesize nitrogen-containing biomolecules, such as nucleic acids and proteins.
Cyanobacteria can form symbiotic associations with animals and plants, such as fungi, bryophytes,
pteriodophytes, gymnosperms and angiosperms, supporting their growth and reproduction (Sarma, 2013;
Hudnell, 2008; Hudnell, 2010; Rai, 1990).
Cyanobacteria can be found in unicellular, colony and multicellular filamentous forms. The unicellular
form occurs when the daughter cells separate after binary fission reproduction. These cells can aggregate
into irregular colonies held together by a slimy matrix secreted during colony growth (WHO, 1999). The
filamentous form occurs when repeated cell divisions happen in a single plane at right angles to the main
axis (WHO, 1999). Reproduction is asexual.
Cyanobacteria are considered gram-negative, even though the peptidoglycan layer is thicker than most
gram-negative bacteria. However, studies using electron microscopy show that cyanobacteria possess
properties of both gram-negative and gram-positive bacteria (Stewart et al., 2006a). Compared to
heterotrophic bacteria, the cyanobacterial lipopolysaccharides (LPS) have little or no 2-keto-3-deoxy-D-
manno-octonic acid, and they lack phosphate groups, glucosamine and L-glycero-D-mannoheptose.
Cyanobacteria also have long-chain saturated and unsaturated fatty acids.
Under the optimal pH, nutrient availability, light and temperature conditions, cyanobacteria can reproduce
quickly forming a bloom. Studies of the impact of environmental factors on cyanotoxin production are
ongoing, including such factors as nutrient (nitrogen, phosphorus and trace metals) concentrations, light,
temperature, oxidative stressors and interactions with other biota (viruses, bacteria and animal grazers), as
well as the combined effects of these factors (Paerl and Otten 2013a; 2013b). Fulvic and humic acids also
have been reported to encourage cyanobacteria growth (Kosakowska et al., 2007).
Cyanobacteria can produce a wide range of bioactive compounds, some of which have beneficial or
therapeutic effects. These bioactive compounds have been used in pharmacology, as dietary supplements
and as mood enhancers (Jensen et al., 2001). Other cyanobacteria can produce bioactive compounds that
may be harmful, called cyanotoxins. The most commonly recognized bioactive compounds produced by
cyanobacteria fall into four broad groupings: cyclic peptides, alkaloids, amino acids and LPS.
Microcystins are produced by several cyanobacterial species, including species of Anabaena, Nodularia,
Nostoc Oscillatoria, members of Microcystis, Fischerella, Planktothrix, and Gloeotrichia echinulata
(Duy et al., 2000; Codd et al., 2005; Stewart et al., 2006a; Carey et al., 2012).
1.2 Microcystin Congeners
The cyclic peptides include six congeners of nodularins and around 100 congeners of microcystins.
Figure 2-1 provides the structure of microcystin, a monocyclic heptapeptide, where X and Y represent
Health Effects Support Document for Microcystins - June, 2015
1
-------
variable amino acids as presented in Table 1-1. Although substitutions mostly occur in positions X and Y,
other modifications have been reported for all of the amino acids (Puddick et al., 2015). The amino acids
are joined end-to-end and then head to tail to form cyclic compounds that are comparatively large,
(molecular weights ranging from -800 to 1,100 g/mole). Table 2-1 lists only the most common
microcystin congeners, of which currently around 100 different congeners have been identified.
N ^ ^CH2
H3C, X
H H
H C02H
H3Cv/
H H CH3 h h-
H CH3 CH3 H
O H CO?H
Figure 1-1. Structure of Microcystin (Kondo et al., 1992).
Nodularin has a similar structure to microcystin and a similar mode of toxicity (McElhiney et al., 2005).
Nodularins show hepatotoxic effects through the inhibition of protein phosphatases just like microcystins
and some have suggested carcinogenic potential of nodularins (Nishiwaki-Matsushima et al., 1992; Ohta
et al., 1994). However, there are no published animals studies evaluating the health effects associated
with exposure to nodularin.
Microcystin congeners vary based on their amino acid composition and through methylation or
demethylation at selected sites within the cyclicpeptide (Table 1-1; Duy et al., 2000). The variations in
composition and methylation account for the large number of toxin congeners. The microcystins are
named based on their variable amino acids, although they have had many other names (Carmichael et al.,
1988). For example, microcystin-LR, the most common congener, contains leucine (L) and arginine (R).
The letters used to identify the variable amino acids are the standard single letter abbreviations for the
amino acids found in proteins. The variable amino acids are usually the L-amino acids as found in
proteins. There has been at least one microcystin where the leucine was D-leucine (Carmichael, 1992).
Table 1-1. Amino Acid Composition of Various Microcystin Congeners (Yuan et al., 1999)
Microcystins Congener
Amino Acid in X
Amino Acid in Y
microcystin-LR
Leucine
Arginine
microcystin-RR
Arginine
Arginine
microcystin-YR
Tyrosine
Arginine
microcystin-LA
Leucine
Alanine
microcystin-LY
Leucine
Tyrosine
microcystin-LF
Leucine
Phenylalanine
microcystin-LW
Leucine
Tryptophan
Health Effects Support Document for Microcystins - June, 2015
2
-------
Most research has concentrated on microcystin-LR with lesser amounts of data available for the other
amino acid combinations. Structurally, the microcystins are monocyclic heptapeptides that contain seven
amino acids: two variable L-amino acids, three common D-amino acids or their derivatives, and two
novel D-amino acids (Adda and Mdha). Adda (3S-amino-9S-methoxy-2,6,8S,-trimethyl-10-phenyldeca-
4,6-dienoic acid) is characteristic of all toxic microcystin structural congeners and is essential for their
biological activity (Rao et al., 2002; Funari and Testai, 2008). Mdha (methyldehydroalanine) is the
second unique component of the microcystins. It plays an important role in the ability of the microcystins
to inhibit protein phosphatases. Figure 1-2 illustrates the structures of the two unique amino acid
microcystin components.
Microcystins are water soluble. In aquatic environments, the cyclic peptides tend to remain contained
within the cyanobacterial cell and are released in substantial amounts only upon cell lysis. Microcystins
are most frequently found in cyanobacterial blooms in fresh and brackish waters (WHO, 1999). Table 1-2
provides chemical and physical properties of microcystin-LR.
Adda Mdha
Figure 1-2. Structure of the Amino Acids Adda and Mdha (Harada et al., 1991).
Table 1-2. Chemical and Physical Properties of Microcystin-LR
Property
Microcystin-LR
Chemical Abstracts Registry (CAS) #
101043-37-2
Chemical Formula
C49H74N10O12
Molecular Weight
995.17 g/mole
Color/Physical State
Solid
Boiling Point
N/A
Melting Point
N/A
Density
1.29 g/cm3
Vapor Pressure at 25°C
N/A
Henry's Law Constant
N/A
Kow
N/A
Koc
N/A
Solubility in Water
Highly
Other Solvents
Ethanol and methanol
Sources: Chemical Book, 2012; TOXLINE, 2012
Health Effects Support Document for Microcystins - June, 2015
3
-------
2.0 TOXIN SYNTHESIS AND ENVIRONMENTAL FATE
2.1 Cyanotoxin Synthesis
Toxin production varies between blooms and within an individual bloom over time (Duy et al., 2000).
Cyanotoxins can be produced by more than one species of cyanobacteria and some species may produce
more than one toxin at a time, resulting in blooms with different cyanotoxins (Funari and Testai, 2008).
The toxicity of a particular bloom is complex, determined by the mixture of species and the variation of
strains with toxic and nontoxic genotypes involved (WHO, 1999). Generally, toxins in cyanobacteria are
retained within the cell unless conditions favor cell wall lysis (ILS, 2000).
The synthesis of cyanotoxins is the focus of much research with evidence suggesting that the production
and accumulation of toxin(s) correlates with cyanobacterial growth rate, with the highest amount being
produced during the late logarithmic phase (Funari and Testai, 2008). For example, Long et al. (2001)
described a positive linear relationship between the content of microcystins in cells and their specific
growth rate.
Evidence suggests that the environmental conditions in which a bloom occurs may alter the levels of toxin
produced. Several culture experiments have suggested that the biosynthesis of microcystin is regulated by
environmental and nutritional factors including light intensity, temperature, and nutrients such as
nitrogen, phosphorus, and iron (Neilan et al., 2007). However, the physiological function of iron is still
unclear. Studies on the effect of different light intensities on microcystin production have yielded
contradictory conclusions (Neilan et al., 2007). The effects of environmental conditions on bloom growth
and toxin production are discussed in more detail in section 2.2.
Although there is little information on the genetic regulation of microcystin production, Dittman et al.
(1997) showed that peptide synthetase genes are responsible for microcystin production. Studies
conducted by Kaebernick et al. (2000) on Microcystis aeruginosa suggest that microcystin is produced
nonribosomally through large multifunctional enzyme complexes consisting of both nonribosomal
peptide synthetase (NRPS) and polyketide synthase (PKS) modules coded by the mcyS (microcystin)
gene cluster. According to Gewolb (2002), most NRPSs are made up of a series of four to 10 modules,
each of which is responsible for specific steps of activation, modification, and condensation during the
addition of one specific amino acid or other compound to the growing linear peptide chain that is then
cyclized to produce microcystin. The sequence of modules in an enzyme determines the type of
microcystin produced (Gewolb, 2002).
The difference in toxicity of microcystin congeners depends on the amino acid composition (Falconer,
2005). Stoner et al (1989) administered by intraperitoneal (i.p.) purified microcystin congeners (-LR, -LA,
-LY and -RR congeners) into ten or more adult male and female Swiss albino mice. Necropsies were
performed to confirm the presence of the pathognomonic hemorrhagic livers. The authors reported LD50
doses of 36 ng/g-bw for microcystin-LR, 39 ng/g-bw for microcystin -LA, 91 ng/g-bw for microcystin -
LY and 111 ng/g-bw for microcystin -RR. Similarly, Gupta et al., (2003) determined LD50s for the MC
congeners LR, RR and YR in female mice using DNA fragmentation assay and histopathology
examinations of the liver and lung. The acute LD50 determinations showed that the most toxic variant
was microcystin -LR (43.0 |_ig/kg). followed by microcystin-YR (110.6 j^ig/kg) and microcystin-RR
(235.4 |_ig/kg). The most toxic microcystins are those with the more hydrophobic L-amino with one or
two hydrophobic amino acids (-LA, -LR, -and -YM) and the least toxic are those with hydrophilic amino
acids, such as microcystin-RR. The Adda group is also important since its removal or saturation of its
double bonds greatly reduces toxicity.
Health Effects Support Document for Microcystins - June, 2015
4
-------
2.2 Environmental Factors that Affect the Fate of Cyanotoxins
Cyanotoxin production is strongly influenced by the environmental conditions that promote growth of
particular cyanobacterial species and strains. Nutrient concentrations, light intensity, temperature, and
other environmental factors affect growth and the population dynamics of cyanobacteria production, as
described below. Although environmental conditions affect the formation of blooms, the number of
cyanobacteria and the concentration of toxins produced are not always closely related. Cyanotoxin
concentrations depend on the dominance and diversity of strains within the bloom along with
environmental and ecosystem influences on bloom dynamics as shown in Figure 2-1 below (Hitzfeld et
al., 2000; WHO, 1999).
2.2.1 Nutrients
Nutrients are key environmental drivers that influence the proportion of cyanobacteria in the
phytoplankton community, the cyanobacterial biovolume, toxin production, and the impact that
cyanobacteria may have on ecosystem function and water quality (Paerl et al., 2011). Cyanobacteria
production and toxin concentrations are dependent on nutrient levels (Wang et al., 2002); however,
different cyanobacteria species use organic and inorganic nutrient forms differently. Loading of nitrogen
(N) and/or phosphorus (P) to water bodies from agricultural, industrial and urban sources influence the
development of cyanobacterial blooms and may be related to cyanotoxin production (Paerl et al., 2011).
Nitrogen loading can enhance the growth and toxin levels of Microcystis sp. blooms and microcystin
synthetase gene expression (Gobler et al., 2007; O'Neil et al., 2012). Gobler et al. (2007) suggest that
dominance of Microcystis sp. blooms during summer is linked to N loading, which stimulates growth and
toxin synthesis. This may cause the inhibition of grazing by mesozooplankton and further accumulation
of cyanobacterial cells.
Optimal concentrations of total and dissolved phosphorous (Wang et al., 2002) and soluble phosphates
and nitrates (ILS, 2000; Paerl and Scott, 2010; Wang et al., 2010; O'Neil et al., 2012) may also result in
the increased production of microcystins. Some studies have observed a decrease in toxicity of
Microcystis sp. after removal of N or inorganic carbon, but no changes was observed when P was
removed from a cyanobacteria culture media (Codd and Poon, 1988). Similarly, Sivonen (1990) found a
relationship between high toxicity and high N concentration, but no effect at higher concentrations of
phosphorus.
Smith (1983) first described a strong relationship between the relative amounts of N and P in surface
waters and cyanobacterial blooms. Smith proposed that cyanobacteria should be superior competitors
under conditions of N-limitation because of their unique capacity for N-fixation. While the dominance of
N-fixing cyanobacteria at low N:P ratios has been demonstrated in mesocosm- and ecosystem-scale
experiments in prairie and boreal lakes (Schindler et al., 2008) the hypothesis that low N:P ratios favor
cyanobacteria formation has been debated and challenged for its inability to reliably predict
cyanobacterial dominance (Downing et al., 2001). Eutrophic systems already subject to bloom events are
prone to further expansion of these blooms due to additional N inputs, especially if these nutrients are
available from internal sources. As the trophic state increases, aquatic systems absorb higher
concentrations of N (Paerl and Huisman, 2008; Paerl and Otten, 2013b). Recent surveys of cyanobacterial
and algal productivity in response to nutrient pollution across geographically diverse eutrophic lakes,
reservoirs, estuarine and coastal waters, and in different experimental enclosures of varying sizes
demonstrate that greater stimulation is routinely observed in response to both N and P additions. Further,
this evidence suggests that nutrient colimitation is widespread (Elser et al., 2007; Lewis et al., 2011; Paerl
et al., 2011). These results strongly suggest that reductions in both N and P inputs are needed to stem
eutrophication and cyanobacterial bloom expansion.
Health Effects Support Document for Microcystins - June, 2015
5
-------
Analysis of observational data collected at larger spatial scales support the idea that controlling Total
Phosphorus (TP) and Total Nitrogen (TN) could reduce the frequency of high MC events by reducing the
biomass of cyanobacteria in the system (Yuan et al., 2014, Orihel et al., 2012; Scott et al., 2013). Some of
these analyses have also found that TN concentrations are the strongest predictors of high MC across
large spatial scales, but the causal mechanisms forthis correlation are still not clear (Scott et al., 2013;
Yuan et al., 2014). Subsequent experiments should manipulate N:P ratios at scales relevant to ecosystem
management to further develop/evaluate the need for a dual nutrient strategy as discussed in Paerl et al.
(2011) and Paerl and Otten (2013b).
2.2.2 Light Intensity
Sunlight availability and turbidity have a strong influence on the cyanobacteria species that predominate,
as well as the depth at which they occur (Falconer et al., 2005; Carey et al., 2012). For example,
Microcystis aeruginosa occurs mostly at the surface with higher light intensities and in shallow lakes. The
relationship of light intensity to toxin production in blooms is somewhat unclear and continues to be
investigated (Duy et al., 2000). Some scientists have found evidence that toxin production increases with
high light intensity (Watanabe and Oishi, 1985) while others have found little variation in toxicity at
different levels of light intensity (Codd and Poon, 1988; Codd, 1995).
Kosten et al. (2011) surveyed 143 shallow lakes along a latitudinal gradient (between 5-55°S and
38-68°N) from subarctic Europe to southern South America). Their analyses found a greater proportion of
the total phytoplankton biovolume attributable to cyanobacteria in lakes with high rates of light
absorption. Kosten et al. (2011) could not establish cause and effect from these field data, but other
controlled experiments and field data have demonstrated that light availability can affect the competitive
balance among a large group of shade-tolerant species of cyanobacteria, mainly Oscillatoriales and other
phytoplankton species (Smith, 1986; Scheffer et al., 1997). Overall, results from Kosten et al. (2011)
suggest that higher temperatures interact with nutrient loading and underwater light conditions in
determining the proportion of cyanobacteria in the phytoplankton community in shallow lakes.
2.2.3 Temperature
The increasing body of laboratory and field data (Weyhenmeyer, 2001; Huisman et al., 2005; Reynolds,
2006; De Senerpont Domis et al., 2007; Jeppesen et al., 2009; Wagner and Adrian, 2009; Kosten et al.,
2011; Carey et al., 2012) suggest that an increase in temperature may influence cyanobacterial dominance
in the phytoplankton community. Cyanobacteria may benefit more from warming than other
phytoplankton groups due to their higher optimum growth temperatures. The optimum temperatures for
microcystin production range from 20 to 25°C (WHO, 2003). The increase in water column stability
associated with higher temperatures also may favor cyanobacteria (Wagner and Adrian, 2009; Carey et
al., 2012). Kosten et al. (2011) demonstrated that during the summer, the percentage of the total
phytoplankton biovolume attributable to cyanobacteria increased steeply with temperature in shallow
lakes sampled along a latitudinal transect ranging from subarctic Europe to southern South America.
Furthermore, warmer temperatures appear to favor the growth of toxigenic strains of Microcystis over
nontoxic ecotypes (Dziallas and Grossart, 2011; Paerl and Otten, 2013b).
Indirectly, warming also may increase nutrient concentrations by enhancing mineralization (Gudasz et al.,
2010; Kosten et al., 2009 and 2010) by temperature- or anoxia-mediated sediment phosphorus release
(Jensen and Andersen, 1992; Sondergaard et al., 2003). Thus, temperature may indirectly increase
cyanobacteria biomass through its effect on nutrient concentrations. Others have suggested that warmer
conditions may raise total phytoplankton biomass through an alteration of top-down regulation by
selective grazing that favors larger size phytoplankton species and cyanobacteria blooms (Jeppesen et al.,
2009, 2010; Teixeira-de Mello et al., 2009). The relationship between temperature and cyanobacterial
Health Effects Support Document for Microcystins - June, 2015
6
-------
dominance may be explained not only by temperature effect on the competitive advantage of
cyanobacteria, but also factors such as the percent area covered and the volume of the lake taken up by
submerged macrophytes (Kosten et al., 2011; Carey et al., 2012).
Rising global temperatures and changing precipitation patterns may stimulate cyanobacteria blooms.
Warmer temperatures favor surface bloom-forming cyanobacterial genera because they are heat-adapted
and their maximal growth rates occur at relatively high temperatures, often in excess of 25°C (Robarts
and Zohary 1987; Reynolds, 2006). At these elevated temperatures, cyanobacteria routinely out-compete
eukaryotic algae (Elliott, 2010; Paerl et al., 2011). Specifically, as the growth rates of the eukaryotic taxa
decline in response to warming, cyanobacterial growth rates reach their optima. Warmer surface waters,
especially in areas of reduced precipitation, are prone to intense vertical stratification. The strength of
vertical stratification depends on the density difference between the warm surface layer and the
underlying cold water which is influenced by amount of precipitation. As temperatures rise due to climate
change, stratification is expected to occur earlier in the spring and persist longer into the fall (Paerl and
Otten, 2013b). The increase in water column stability associated with higher temperatures and climate
change may therefore favor cyanobacteria production and possibly the prevalence of cyanotoxins such as
microcystins (Wagner and Adrian, 2009; Carey et al., 2012).
2.2.4 Other Environmental Factors
Cyanobacteria blooms have been shown to intensify and persist at pH levels between six and nine (WHO,
2003). When these blooms are massive or persist for a prolonged period, they can become harmful.
Kosten et al. (2011) noted the impact of pH on cyanobacteria abundance in lakes along a latitudinal
transect from Europe to southern South America. The percentage of cyanobacteria in the 143 shallow
lakes sampled was highly correlated with pH, with an increased proportion of cyanobacteria at higher pH.
Cyanobacteria have a competitive advantage over other phytoplankton species because they are efficient
users of carbon dioxide in water (Shapiro, 1984; Caraco and Miller, 1998). This characteristic is
especially advantageous for cyanobacteria under conditions of higher pH when the concentration of
carbon dioxide in the water column is diminished due to photosynthetic activity. Although this could
explain the positive correlation observed between pH and the proportion of cyanobacteria, the high
proportion of cyanobacteria at high pH could be the result of an indirect nutrient effect as described
previously (see discussion in Temperature Section). As photosynthesis intensifies, pH increases due to
carbon dioxide uptake by algae, resulting in a shift in the carbonic buffer equilibrium and a higher
concentration of basic forms of carbonate. Thus, higher water column pH may be correlated with a higher
proportion of cyanobacteria because of higher photosynthetic rates, which can be linked with high
nutrient concentrations (Duy et al., 2000) that stimulate phytoplankton growth and bloom formation. High
iron concentrations (more than 100 (.iM) have also been shown to increase cell density and chlorophyll
content in Microcystis aeruginosa (Kosakowska et al., 2007).
Most phytoplankton-cyanobacteria blooms occur in late summer and early fall when deeper lakes or
reservoirs are vertically stratified and phytoplankton species may be stratified as well. Vertical
phytoplankton biomass structure and cyanotoxin production can be influenced by seasonal changes as
well as severe weather conditions (e.g., strong wind or rainfall), and also by runoff. At times, the
hypolimnion (bottom layer of the water column) can have a higher phytoplankton-cyanobacteria biomass
and display different population dynamics than the epilimnion (upper layer of the water column).
Conversely, seasonal effects of increasing temperatures and changes in wind patterns may favorably
influence the upper water column cyanobacterial community. This vertical variability is common and
attributed to four causes, each of which may occur at different times, including: (a) sinking of dead/dying
cells; (b) density stratification of the water column, especially nutrient concentrations and light, which
affects all aspects of cyanobacteria growth; (c) increased nutrient supply from organic-rich bottom
Health Effects Support Document for Microcystins - June, 2015
7
-------
sediment (even when the water body is not density-stratified), encouraging cyanobacteria growth at or
near the bottom sediment; and (d) species-specific factors such as the tendency to form surface scums in
the case ofM aeruginosa or the presence of resting spores in the sediment in the case of N spnmigena
(Drake et al., 2010). In addition, there are microbial interactions that may occur within blooms, such as
competition and adaptation between toxic and nontoxic cyanobacterial strains, as well as impacts from
viruses. Each of these factors can cause fluctuations in bloom development and composition. When the
composition of the cyanobacteria bloom changes, so do the toxins present and their concentrations (Honjo
et al., 2006: Paerl and
Positive
1 High P (High N for some)
' Low N (DIN, DON) (only
applies to N2 fixers)
1 Low N:P Ratios
1 Low turbulence
1 Low water flushing-Long
water residence time
1 High (adequate) light
1 Warm temperatures
1 High dissolved organic
matter
1 Sufficient Fe (+ trace
metals)
1 Low grazing rates
w
CD
"5
ac
Cyanos
CO
CD
¦*—>
03
cc
Diversity
Modulating factors
• Strong biogeochemical gradients (e.g.
persistent stratification, stable benthos)
• Heterogeneous and diverse habitats (e.g.
reefs, seagrasses, marshes, sediments,
aggregates)
• Selective grazing
• 'Toxin" production??
Figure 2-1. Environmental Factors Influencing Cyanobacterial Blooms.
Negative
1 High DIN/total N (only
applies to N2 fixers)
• Low P (DIP)
1 High N:P ratios
1 High turbulence & vertical
mixing
1 High water flushing-Short
water residence time
1 Low light (for most taxa)
1 Cool temperatures
1 Low dissolved organic
matter
• Low Fe (+ trace metals)
1 High grazing rates
1 Viruses (cyanophages)
• Predatory bacteria
Otten, 2013b). The concentration of cyanotoxins observed in a water body when a bloom collapses, such
as from cell aging or algaecide treatment, depends on dilution of the toxin due to water column mixing,
the degree of adsorption to sediment or particulates and the rate of toxin biodegradation (Funari and
Testai, 2008).
In summary, there is a complex interplay of environmental factors that dictates the spatial and temporal
changes in the concentration of cyanobacteria cells and their toxins with respect to the dominant species
as illustrated in Figure 2-1 (Paerl and Otten, 2013b). Factors such as the N:P ratio, organic matter
availability, temperature, and light attenuation, as well as other physico-chemical processes, can play a
role in determining harmful algal bloom (HAB) composition and toxin production (Paerl and Huisman,
2008; Paerl and Otten, 2013b). Dynamics of microflora competition as blooms develop and collapse can
also impact cyanotoxin concentrations in surface waters. In addition, impacts of climate change, including
potential warming of surface waters and changes in precipitation, could result in changes in ecosystem
dynamics that lead to more frequent formation of cyanobacteria blooms and their associated toxins (Paerl
and Huisman, 2008; Paerl et al., 2011; Paerl and Otten, 2013b).
Health Effects Support Document for Microcystins - June, 2015
8
-------
2.3 Environmental Fate of Microcystins
2.3.1 Hydrolysis
Microcystins are relatively stable and resistant to chemical hydrolysis or oxidation at near neutral pH.
Elevated or low pH or temperatures above 30°C may cause slow hydrolysis. They have been reported to
remain potent toxins even after boiling (Rao et al., 2002). In natural waters, microcystins may persist for
between 21 days and 2-3 months in solution and up to 6 months in dry scum when kept in the dark
(Rapala et al., 2006; Funari and Testai, 2008).
2.3.2 Photolysis
In the presence of full sunlight, microcystins undergo slow photochemical breakdown, but this varies by
microcystin congener (WHO, 1999; Chorus et al., 2000). The presence of water-soluble cell pigments, in
particular phycobiliproteins, enhances this breakdown. Breakdown can occur rapidly in as few as two
weeks or longer than six weeks, depending on the concentration of pigment and the intensity of the light
(Tsuji et al., 1993; 1995). Microcystin-LR was photodegraded with a half-life (time it takes half of the
toxin to degrade) of about 5 days in the presence of 5 mg/L of extractable cyanobacterial pigment. Humic
substances can also act as photosensitizers and increase the rate of microcystin breakdown with sunlight.
In deeper or turbid water, the breakdown rate is slower.
2.3.3 Metabolism
Microcystins are susceptible to degradation by aquatic bacteria found naturally in rivers and reservoirs
(Jones etal., 1994). Bacteria isolates of Arthrobacter, Brevibacterium, Rhodococcus, Paucibacter, and
various strains of the genus Sphingomonas (Pseudomonas) have been reported to be capable of degrading
MC-LR (de la Cruz et al., 2011; Han et al., 2012). These degradative bacteria have also been found in
sewage effluent (Lam et al., 1995), lake water (Jones et al., 1994; Cousins et al., 1996; Lahti et al.,
1997a), and lake sediment (Rapala et al., 1994; Lahti et al., 1997b). Lam et al (1995) reported that the
biotransformation of microcystin-LR followed a first-order decay with a half-life of microcystin
biotransformation of 0.2 to 3.6 days (Lam et al., 1995). Jones et al. (1994) evaluated degradation of
microcystin-LR in different natural surface waters. Microcystin-LR persisted for 3 days to 3 weeks before
being degraded. However, degradation was fairly rapid with more than 95% loss happening within 3 to 4
days (Jones et al., 1994). A study by Christoffersen et al. (2002), reported half-lives of microcystin-LR in
the laboratory and in the field of approximately 1 day, which were driven largely by bacterial aerobic
metabolism. These researchers found that approximately 90% of the initial amount of microcystin
disappeared from the water phase within 5 days, irrespective of the starting concentration. Other
researchers (Edwards et al., 2008) have reported longer half-lives of 4-14 days, with longer half-lives
associated with streams and shorter half-lives associated with lakes.
2.3.4 Transport
Microcystins may adsorb onto naturally suspended solids and dried crusts of cyanobacteria. They can
precipitate out of the water column and reside in sediments for months (Han et al., 2012: Falconer,
1998). A study by United States Geological Survey (USGS) and the University of Central Florida
determined that microcystins did not sorb in sandy aquifers and were transported along with groundwater
(O'Reilly et al., 2011). The authors suggested that the removal of microcystins was due to biodegradation.
Health Effects Support Document for Microcystins - June, 2015
9
-------
2.4 Summary
Microcystins are produced by a variety of cyanobacteria. Currently around 100 different congeners of
microcystins have been identified, with Microcystin-LR the most common and best known congener
worldwide. Environmental conditions such as nutrients, pH, light intensity and temperature influence the
growth of these cyanobacteria and could encourage toxin production. Microcystins are water soluble and
tend to remain contained within the cyanobacterial cell until the cell lyses (dies) and they are released in
substantial amounts into the water. They are stable and resistant to chemical hydrolysis or oxidation at
near neutral pH. Slow hydrolysis may occur at elevated or low pH, or temperatures above 30°C.
Microcystins remain potent even after boiling for 15 minutes. In the dark, microcystins may persist from
21 days to 3 months in solution and up to 6 months in dry scum. In the presence of full sunlight,
microcystins undergo slow photochemical degradation which varies by microcystin congener and could
take about one to a few weeks, or longer than six weeks to degrade. The presence of water-soluble cell
pigments, in particular phycobiliproteins, enhances this breakdown. Half-lives vary from 4 to 14 days.
Microcystins are susceptible to degradation by aquatic bacteria found naturally in rivers and reservoirs
(Jones et al., 1994). Half-lives vary from 0.2 to 14 days.
Health Effects Support Document for Microcystins - June, 2015
10
-------
3.0 CYANOTOXIN OCCURRENCE AND EXPOSURE IN WATER
The presence of detectable concentrations of cyanotoxins in the environment is closely associated with
blooms of cyanobacteria. Cyanobacteria flourish in various natural environments including salty, brackish
or fresh water, cold and hot springs, and in environments where no other microalgae can exist, including
desert sand, volcanic ash and rocks (Jaag, 1945; Dor and Danin, 1996). Cyanobacteria also form
symbiotic associations with aquatic animals and plants, and cyanotoxins are known to bioaccumulate in
common aquatic vertebrates and invertebrates (Ettoumi et al. 2011).
Currently, there is no national database recording freshwater harmful algal bloom (HAB) events. Instead,
state and local governments document HAB occurrences in various ways depending on the monitoring
methods used and the availability of laboratories capable of conducting algal toxin analyses.
Human exposure to cyanotoxins, including microcystin, may occur by direct ingestion of toxin-
contaminated water or food, and by inhalation and dermal contact during bathing, showering or during
recreational activities in waterbodies contaminated with the toxins. Microcystins can be dissolved in
drinking water either by the breakdown of a cyanobacterial bloom or by cell lysis. Exposure can occur via
drinking water as some water treatment technologies are not designed for removal of cyanotoxins.
Because children consume more water per unit body weight than do adults, children may potentially
receive a higher dose on a per body-weight basis. Exposure through drinking water can occur if there are
toxins in the water source and the existing water treatment technologies were not designed for removal of
cyanotoxins. Because children consume more water per unit body weight than do adults, children
potentially may receive a higher dose than adults. Exposures are usually not chronic; however, they can
be repeated in regions where cyanobacterial blooms are more extensive or persistent. Exposure to
microcystin from ambient surface waters is more likely to be acute or subacute as is most likely to occur
during a bloom. People, particularly children, recreating close to lakes and beach shores also can be at
potential risk from exposure to nearshore blooms.
Livestock and pets potentially can be exposed to higher concentrations of cyanobacterial toxins than
humans because they are more likely to consume scum and mats while drinking cyanobacteria-
contaminated water (Backer et al., 2013). Dogs are particularly at risk as they may lick cyanobacteria
from their fur after swimming in a water body with an ongoing bloom.
3.1 General Occurrence of Cyanobacteria in Water
Species of cyanobacteria are predominantly found in eutrophic (nutrient-rich) water bodies in freshwater
and marine environments (ILS, 2000), including salt marshes. Most marine cyanobacteria of known
public health concern grow along the shore as benthic vegetation between the low- and high-tide water
marks. The marine planktonic forms have a global distribution. They also can be found in hot springs
(Castenholz, 1973; Mohamed, 2008), mountain streams (Kann, 1988), Arctic and Antarctic lakes
(Skulberg, 1996) and in snow and ice (Laamanen, 1996).
A visibly colored scum formed by floating cells may contain more than 10,000 cells/mL (Falconer, 1998).
The floating scum, as in the case of Microcystis spp., may be concentrated by prevailing winds in certain
surface water areas, especially at the shore.
3.2 Microcystins Occurrence in Surface Water
The microcystins are the most common cyanotoxins found worldwide and have been reported in surface
waters in most of the U.S. and Europe (Funari and Testai, 2008). Dry-weight concentrations of
microcystins in surface freshwater cyanobacterial blooms or surface freshwater samples reported
worldwide between 1985 and 1996 ranged from 1 to 7,300 jj.g/g. Water concentrations of extracellular
Health Effects Support Document for Microcystins - June, 2015
11
-------
plus intracellular microcystins ranged from 0.04 to 150,000 (ig/L. The concentration of extracellular
microcystins ranged from 0.02 to a high of 1,800 (ig/L, which occurred following treatment of a large
bloom with algaecide (WHO, 1999). A concentration of 150,000 (ig/L total microcystins was reported by
the USGS in a lake in Kansas (Graham et al., 2012).
According to a survey conducted in Florida in 1999 between the months of June and November, the most
frequently observed cyanobacteria werq Microcystis (43.1%), Cylindrospermopsis (39.5%), and
Anabaena spp (28.7%) (Burns, 2008). Of 167 surface water samples taken from 75 waterbodies, 88
samples were positive for cyanotoxins. Microcystin was the most commonly found cyanotoxin in water
samples collected, occurring in 87 water samples.
In 2002, the Monitoring and Event Response to Harmful Algal Blooms in the Lower Great Lakes
(MERHAB-LGL) project evaluated the occurrence and distribution of cyanobacterial toxins in the lower
Great Lakes region (Boyer, 2007). Analysis for total microcystins was done using Protein Phosphatase
Inhibition Assay (PPIA). Microcystins were detected in at least 65% of samples, mostly in Lake Erie,
Lake Ontario, and Lake Champlain. The National Oceanic and Atmospheric Administration (NOAA)
Center of Excellence for Great Lakes and Human Health (CEGLHH) continues to monitor the Great
Lakes and regularly samples algal blooms for microcystin in response to bloom events.
A 2004 study of the Great Lakes found high levels of cyanobacteria during the month of August
(Makarewicz et al., 2006). Microcystin-LR was analyzed by PPIA (limit of detection of 0.003 (ig/L) and
was detected at levels of 0.084 |ig/L in the nearshore and 0.076 |ig/L in the bays and rivers. The study
reported higher levels of microcystin-LR (1.6 to 10.7fxg/L) in smaller lakes in the Lake Ontario
watershed.
In 2006, the USGS conducted a study of 23 lakes in the Midwestern U.S. in which cyanobacterial blooms
were sampled to determine the co-occurrence of toxins (Graham et al., 2010). The study reported that
microcystins were detected in 91% of the lakes sampled. Mixtures of all the microcystin congeners
measured (LA, LF, LR, LW, LY, RR, and YR) were common and all the congeners were present in the
blooms. Microcystin-LR and -RR were the dominant congeners detected with mean concentrations of
104 and 910 (ig/L, respectively.
EPA's National Aquatic Resource Surveys (NARS) generate national estimates of pollutant occurrence
every 5 years. In 2007, the National Lakes Assessment (NLA) conducted the first-ever national
probability-based survey of the nation's lakes, ponds and reservoirs (U.S.EPA, 2009). This baseline study
of the condition of the nation's lakes provided estimates of the condition of natural and man-made
freshwater lakes, ponds, and reservoirs greater than 10 acres and at least one meter deep. A total of 1,028
lakes were sampled for the NLA during summer 2007. The NLA measured microcystins using Enzyme
Linked Immunosorbent Assays (ELISA) with a detection limit of 0.1 (ig/L, as well as cyanobacterial cell
counts and chlorophyll-a concentrations, as indicators of the presence of cyanobacterial toxins. Samples
were collected in open water at mid-lake; no samples were taken nearshore or other areas where scums
were present.
A total of 48 states were sampled in the NLA and states with lakes reporting microcystin levels above the
WHO's moderate risk1 threshold in recreational water (>10 j^ig/L) are shown in Table 3-1. Microcystins
were present in 30% of the lakes sampled nationally, with sample concentrations that ranged from the
limit of detection (0.1 (ig/L) to 225 (ig/L. Two states (North Dakota and Nebraska) had 9% of the samples
1 The WHO established guideline values for recreational exposure to cyanobacteria using a three-tier approach: low
risk (<20,000 cyanobacterial cells/ml corresponding to <10 |ig/L of MC-LR); moderate risk (20,000-100,000
cyanobacterial cells/ml corresponding to 10-20 |ig/L of MC-LR); and high risk (>100,000 cyanobacterial cells/ml
corresponding to >20 |ig/L for MC-LR) (WHO, 1999).
Health Effects Support Document for Microcystins - June, 2015
12
-------
above 10 (ig/L. Other states, including Iowa, Texas, South Dakota, and Utah also had samples that
exceeded 10 (ig/L. Several samples in North Dakota, Nebraska, and Ohio exceeded the WHO high risk
threshold value for recreational waters of 20 (ig/L (192, 225 (ig/L, and 78 (ig/L, respectively). EPA
completed a second survey in 2012 but data have not yet been published.
Table 3-1. States Surveyed as Part of the 2007 National Lakes Assessment with Water Body
Microcystin Concentrations Above the WHO Advisory Guideline Level for Recreational Water of
10 |jg/L(U.S. EPA, 2009)
State
Number of
Sites
Sampled
Percentage of Samples with Detection of
Microcystins >10 |jg/L
Maximum Detection of
Microcystins
North Dakota
38
9.1%
192 |jg/L
Nebraska
42
9.1%
225 |jg/L
South Dakota
40
4.9%
33 |jg/L
Ohio
21
4.5%
78 |jg/L*
Iowa
20
4.5%
38 |jg/L*
Utah
26
3.6%
15 |jg/L*
Texas
51
1.8%
28 |jg/L *
* Single Sample
Microcystins have been detected in most of the states of the U.S. and over the past several years and
many studies have been conducted to determine their occurrence in surface water. USGS, for example,
did a study in the Upper Klamath Lake in Oregon in 2007 and detected total microcystin concentrations
between 1 (ig/L and 17 (ig/L (VanderKooi et al., 2010). USGS also monitored Lake Houston in Texas
from 2006 to 2008, and found microcystins in 16% of samples, with concentrations less than or equal to
0.2 (ig/L (Beussink and Graham, 2011). In 2011, USGS conducted a study on the upstream reservoirs of
the Kansas River, a primary source of drinking water for residents in northeastern Kansas, to characterize
the transport of cyanobacteria and associated compounds (Graham et al., 2012). Concentrations of total
microcystins were low in the majority of the tributaries with the exception of Milford Lake, which had
higher total microcystin concentrations, some of which exceeded the Kansas recreational guidance level
of 20 (ig/L. Upstream from Milford Lake, a cyanobacterial bloom was observed with total microcystin
concentration of 150,000 (ig/L. When sampled a week later, total microcystin concentrations were less
than 1 (ig/L. This may have been due either to dispersion of microcystins through the water column or to
other areas or settling of cyanobacteria out of the water column. Samples taken during the same time from
outflow waters contained a total microcystin concentration of 6.2 (ig/L.
In 2005, Washington State Department of Ecology developed the Ecology Freshwater Algae Program to
focus on the monitoring and management of cyanobacteria in Washington lakes, ponds, and streams
(WSDE, 2012). The data collected have been summarized in a series of reports for the Washington State
Legislature (Hamel, 2009, 2012). Microcystin levels ranged from the detection limit (0.05 (ig/L) to 4,620
(ig/L in 2008, 18,700 (ig/L in 2009, 853 (lg/L in 2010, and 26,400 (ig/L in 2011.
Other surveys and studies have been conducted to determine the occurrence of microcystin in lakes in the
United States. A survey conducted during the spring and summer of 1999 and 2000 in more than 50 lakes
in New Hampshire found measureable microcystin concentrations in all samples (Haney and Ikawa,
2000). Microcystins were analyzed by ELISA and were found in all of the lakes sampled with a mean
concentration of 0.1 (ig/L. In 2005 and 2006, a study conducted in New York, including Lake Ontario,
found variability in microcystin-LR concentrations within the Lake Ontario ecosystem (Makarewicz et
al., 2009). Of the samples taken in Lake Ontario coastal waters, only 0.3% of the samples exceeded the
WHO provisional guideline value for drinking water of 1 (.ig/L. However, 20.4% of the samples taken at
upland lakes and ponds within the Lake Ontario watershed, some of them sources of drinking water,
Health Effects Support Document for Microcystins - June, 2015
13
-------
exceeded 1 (.ig/L. During 2008 and 2009, a study was done in Kabetogama Lake, Minnesota to measure
microcystin concentrations associated with algal blooms (Christensen et al., 2011). Microcystins were
detected in 78% of bloom samples. Of these, 50% were above 1 |ig/L in finished drinking water and two
samples were above the high risk WHO recreational level of 20 |ig/L.
A study from 2002 evaluated water quality, including chlorophyll-a concentrations, cyanobacterial
assemblages, and microcystin concentrations in 11 potable water supply reservoirs within the North
Carolina Piedmont during the dry summer growing season (Touchette et al., 2007). Microcystin
concentrations were assessed using ELISA. The study found that cyanobacteria were the dominant
phytoplankton community, averaging 65-95% of the total cells. Although microcystin concentrations
were detected in nearly all source water samples, concentrations were <0.8 (ig/L.
Since 2007, Ohio EPA (OHEPA, 2012) has been monitoring inland lakes for cyanotoxins. Of the 19 lakes
in Ohio sampled during the NLA, 36% had detectable levels of microcystins. In 2010, OHEPA sampled
Grand Lake St. Marys for anatoxin-a, cylindrospermopsin, microcystins, and saxitoxin. Toxin levels
ranged from below the detection limit (<0.15 |_ig/L) to more than 2,000 (ig/L for microcystins. Follow-up
samples taken in 2011 for microcystins indicated concentrations exceeding 50 (ig/L in August. During the
same month, sampling in Lake Erie found microcystins levels to exceed 100 (ig/L.
In 2008, NOAA began monitoring cyanobacterial blooms in Lake Erie using high temporal resolution
satellite imagery. Between 2008 and 2010, Microcystis cyanobacterial blooms associated with water
temperatures above 18°C were detected (Wynne et al., 2013). Using the Great Lakes Coastal Forecast
System (GLCFS) hydrodynamic model, forecasts of bloom transport are created to estimate the trajectory
of the bloom and these are distributed as bulletins to local managers, health departments, researchers and
other stakeholders. To evaluate bloom toxicity, the Great Lakes Environmental Research Laboratory
(GLERL) collected samples at six stations each week for 24 weeks, measuring toxin concentrations as
well as chlorophyll biomass and an additional 18 parameters (e.g., nutrients) to improve future forecasts
of these blooms. In 2014, particulate toxin concentrations, collected from 1 meter depth, ranged from
below detection to 36.7 (ig/L. Particulate toxin concentrations peaked in August, 2014 at all sites, with the
Maumee Bay site yielding the highest toxin concentration for the entire sampling period. Dissolved toxin
concentrations were collected at each site from September until November when the field season
ended. During the final months of sampling (October-November) dissolved toxin concentrations were
detected with peak concentrations of 0.8 |ig/L (mean: 0.28 +/- 0.2 |ig/L) whereas particulate toxin
concentrations were below detection limits on many dates indicating that a majority of the toxins (mean:
72% +/- 37%) were in the dissolved pool as the bloom declined in intensity.
Concentrations of microcystins were detected during sampling in 2005 and 2006 in lakes and ponds used
as a source of drinking water within the Lake Ontario watershed (Makarewicz et al., 2009). A
microcystin-LR concentration of 5.07 (ig/L was found in Conesus Lake, a source of public water supply
that provides drinking water to approximately 15,000 people. Microcystin-LR was also detected at 10.716
(ig/L in Silver Lake, a public drinking water supply for four municipalities.
3.3 Microcystins Occurrence in Drinking Water
The occurrence of cyanotoxins in drinking water depends on their levels in the raw source water and the
effectiveness of treatment methods for removing cyanobacteria and cyanotoxins during the production of
drinking water. Currently, there is no program in place to monitor for the occurrence of cyanotoxins at
surface-water treatment plants for drinking water in the U.S. Therefore, data on the presence or absence
of cyanotoxins in finished drinking water are limited.
Health Effects Support Document for Microcystins - June, 2015
14
-------
The American Water Works Association Research Foundation (AWWARF) conducted a study on the
occurrence of cyanobacterial toxins in source and treated drinking waters from 24 public water systems in
the U.S. and Canada in 1996-1998 (AWWARF, 2001). Of 677 samples tested, microcystin was found in
80% (539) of the waters sampled, including treated waters. Only two samples of finished drinking water
had microcystin concentrations above 1 (.ig/L. A survey conducted in 1999 in Florida (Burns, 2008)
reported that microcystins were the most commonly found toxin in pre- and post-treated drinking water.
Finished water concentrations ranged from below detection levels to 12.5 (ig/L.
A study from 2002 conducted during the dry summer growing season, evaluated the water quality and
environmental parameters, including phytoplankton chlorophyll a concentrations, cyanobacterial
assemblages, and microcystin concentrations in 11 potable water supply reservoirs within the North
Carolina Piedmont (Touchette et al., 2007). The study found that cyanobacteria were the dominant
phytoplankton community, averaging 65-95% of the total cells. Although microcystin concentrations
were detected in nearly all samples, microcystin-LR was detected below l^g/L.
During the summer of 2003, a survey was conducted to test for microcystins in 33 U.S. drinking water
treatment plants in the Northeast and Midwest (Haddix et al., 2007). Microcystins were detected at low
levels ranging from undetectable (<0.15 j^ig/L) to 0.36 (ig/L in all 77 finished water samples.
Concentrations of microcystin-LR have been detected during sampling in 2005 and 2006 in lakes and
ponds used as a source of drinking water within the Lake Ontario watershed (Makarewicz et al., 2009). A
Microcystin-LR concentration of 5.070 (ig/L was measured in Conesus Lake, a source of public water
supply that provides drinking water to approximately 15,000 people. Microcystin-LR was also detected at
10.716 (ig/L in Silver Lake, a public drinking water supply for four municipalities.
In August 2014, the city of Toledo, Ohio issued a "do not drink or boil advisory" to nearly 500,000
customers in response to the presence of total microcystins in the city's finished drinking water at levels
up to 2.50 (ig/L. The presence of the toxins was due to a cyanobacterial bloom near Toledo's drinking
water intake located on Lake Erie. The advisory was lifted two days later, after treatment adjustments led
to the reduction of the cyanotoxin concentrations to concentrations below the WHO guideline value of 1
(.ig/L in all samples from the treatment plant and distribution system.
3.4 Summary
Microcystin-producing cyanobacteria occur in freshwater systems worldwide. No national database
recording freshwater microcystins is available. Microcystin monitoring efforts in surface waters and
drinking water is being conducted by states and others, including USGS, EPA, and NOAA. A survey
done by USGS in 2006 of 23 lakes in the Midwestern U.S., found that microcystin was detected in all the
blooms. Mixtures of all the microcystin congeners measured (LA, LF, LR, LW, LY, RR, and YR) were
common, and all the congeners were present in the blooms. The 2007 EPA National Lakes Assessment
found microcystin in about one third of the lakes sampled with concentrations ranging from the limit of
detection (0.05 j^ig/L) to 225 (ig/L. Sampling done in 2014 in Lake Erie by NOAA reported microcystin
concentrations ranging from below detection limits to 36.7 (ig/L. The U.S. Geological Survey (USGS)
reported a concentration of 150,000 (ig/L total microcystins, in a lake in Kansas (Graham et al., 2012).
Microcystins have been found in raw and in finished drinking water. In a study done in 2007 in 33 lakes
across the U.S., microcystins exceeded 1 (ig/L levels in 7% of the raw water samples. A survey conducted
in 1999 in Florida found microcystins concentrations in finished water ranging from below detection
levels to 12.5(ig/L.
Exposure to microcystin from contaminated drinking water sources may occur mostly via oral exposure
(e.g. ingestion of contaminated drinking water), dermal exposure (contact of exposed parts of the body
Health Effects Support Document for Microcystins - June, 2015
15
-------
with water containing toxins) and inhalation exposure. Exposure to microcystins during recreational
activities may occur through direct contact, inhalation and/or ingestion. Exposures are usually not chronic
with the exception of regions with extensive and persistent cyanobacterial blooms. Since children
consume more water per unit body weight than do adults, children may potentially receive a higher dose.
Pets, livestock and wildlife are also potentially exposed to microcystin when consuming scum and mats,
and drinking cyanobacteria-contaminated water.
Health Effects Support Document for Microcystins - June, 2015
16
-------
4.0 CYANOTOXIN OCCURRENCE IN MEDIA OTHER THAN WATER
4.1 Occurrence in Soil and Edible Plants
Cyanobacteria are highly adaptable and have been found to colonize infertile substrates, such as volcanic
ash and desert sand (Jaag, 1945; Dor and Danin, 1996; Metcalf et al., 2012). They also have been found in
soil, at the surface or several centimeters below the surface, where they play a functional role in nutrient
cycling. Cyanobacteria are known to survive on rocks or tree trunks, and in snow and ice (Adhikary,
1996). They have been reported in deeper soil layers, likely transported by percolating water or burrowing
animals. Some freshwater species are halotolerant (salt tolerant) and have been found in saline
environments such as salt works or salt marshes (WHO, 1999). Cyanobacterial cells can bioaccumulate in
zooplankton (Watanabe et al., 1992). As a result of higher trophic level grazing, the damaged or residual
cyanobacterial cells may settle out of the water column and accumulate in sediment where breakdown by
sediment bacteria and protozoa can release their toxins (Watanabe et al., 1992).
Cyanobacterial cells and toxins can contaminate spray irrigation water and subsequently be associated
with crop plants after spray irrigation with contaminated water (Corbel et al., 2014). Water contaminated
with cyanobacterial cells and toxins used for spray irrigation of crop plants may cause food chain
contamination since low levels of cyanotoxins could be absorbed by roots, migrate to shoots, and then be
translocated to grains and or fruits. Cyanotoxins can be transmitted to food plants from irrigation water
when cyanotoxins are deposited on the plants leaves. A study was conducted with lettuce plants grown
with spray irrigation containing M. aeruginosa at levels ranging from 0.094 to 2.487 jxg/g dw. Cyanotoxin
levels detected in lettuce leaf extracts 10 days after irrigation indicated microcystin-LR equivalents up to
2.49 jxg/g dw (Codd et al., 1999). Extracts from rape and rice seedlings were exposed to water with
concentrations of microcystin-LR up to 3 mg/L (Chen et al., 2004a). The study found concentrations of
microcystin-LR of 651 ng/g in extracts from rape and 5.4 ng/g in rice. These studies and others with high
concentrations of cyanotoxins found that concentrations at these levels are able to inhibit plant growth
causing visible toxic effects on the plant such as leaf withering. The microcystin concentrations detected
in rice grains were very low. Studies with seedlings exposed to cyanotoxin concentrations typically found
in natural surface waters (1-10 |ig/L) reported microcystins at low levels in broccoli roots (0.9 to 2.4 ng
microcystin-LR/g fresh weight) and mustard roots (2.5 to 2.6 ng microcystin-LR/g fresh weight)
(Jarvenpaa et al., 2007).
Uptake of microcystins was measured in vegetables grown with irrigated contaminated groundwater in
Saudi Arabia (Mohamed and Al Shehri, 2009). The concentration of total microcystins was highly
variable in the plants but positively correlated with concentrations in groundwater. Radishes had the
highest concentration (1.2 jj.g/g fresh weight) and cabbages had the lowest amount (0.07 jj.g/g fresh
weight). Lettuce, parsley, arugula, and dill also had measurable concentrations. Generally, roots
accumulated more than the leaves.
Water contaminated with cyanotoxins used for spray irrigation of crop plants will inhibit plant growth and
will induce visible toxic effects such as the appearance of brown leaves (Funari and Testai, 2008).
Therefore, according to the authors, affected plants and crops will most likely not be used for eating
purposes. Further investigation is needed to understand the uptake and fate of microcystins and other
cyanobacterial toxins by food plants.
4.2 Occurrence in Fish and Shellfish
Cyanotoxins can bioaccumulate in common aquatic vertebrates and invertebrates, including fish, snails
(Carbis et al., 1997; Beattie et al., 1998; Berry et al., 2012) and mussels (Eriksson et al., 1989; Falconer et
al., 1992; Prepas et al., 1997; Watanabe et al., 1997; Funari and Testai, 2008). Human exposure to
Health Effects Support Document for Microcystins - June, 2015
17
-------
cyanotoxins may occur if fish are consumed from reservoirs with existing blooms of toxin-producing
cyanobacteria (Magalhaes et al., 2001).
The health risk from fish and shellfish consumption depends on the bioaccumulation of toxins in edible
fish tissue compared to toxins in organs such as the liver. Numerous authors have found that microcystins
accumulate to a lesser extent in the edible parts of aquatic organisms, such as muscle (Xie et al., 2005;
Zimba et al., 2006; Song et al., 2009; Wilson et al., 2008; Deblois et al., 2011; Vareli et al., 2012;
Gutierrez-Praena et al., 2013). In a survey of microcystin in water and fish in two temperate Great Lakes
(Erie and Ontario), the highest microcystin concentrations in fish muscle observed Lake Erie were for
alewives (20.0-37.5 |ig/kg) and northern pike (1.6-25.8 |ig/kg): and for Lake Ontario: walleye (5.3-41.2
|ig/kg). white bass (4.2-27.1 |ig/kg) and smallmouth bass (1.5-43.6 |ig/kg) (Poste et al., 2011). Muscle
tissue microcystin concentrations in yellow perch collected during a toxic bloom were lower in
comparison (0.12- 0.02 ng toxin/g dw) (Wilson et al., 2008). Nevertheless, concentrations of microcystin
in edible tissues have been reported to be greater than 0.1 jj.g/g for fish, crab, mussels and shrimp
(Magalhaes et al., 2001; Mohamed et al., 2003; Xie et al., 2005; Vareli et al., 2012).
Microcystins have been shown to bioaccumulate in the liver and hepatopancreas of decapod crustaceans
(Williams et al., 1997), but there was not strong evidence for biomagnification (Ibelings et al., 2005; Xie
et al., 2005; Ibelings and Havens, 2008; Papadimitriou et al., 2012). Because fish are generally more
tolerant of cyanobacterial toxins than mammals, they tend to accumulate them over time (ILS, 2000).
In a survey by Xie et al. (2005) microcystin-LR content in muscle was highest in carnivorous and
omnivorous fish and was lowest in phytoplanktivorous and herbivorous fish. Chen et al. (2009) also
found highest total microcystin levels in liver and muscle from omnivorous fish compared with other
types of feeders. Berry et al. (2011) found the highest levels in phytoplanktivores and omnivores with no
microcystins detected in predominantly zooplanktivorus fish. Microcystin-LR was not detected in livers
from northern pike and white sucker fish collected from a lake in Canada following peak seasonal
microcystin levels measured in the water (Kotak et al., 1996).
After fish are exposed, concentrations of microcystins decrease with time as a result of detoxification and
depuration processes (Tencalla and Dietrich, 1997; Xie et al., 2005; Mohamed and Hussain, 2006; Wood
et al., 2006; Gutierrez-Praena et al., 2013). Researchers have also suggested that biodilution may occur
given the observations of depuration and toxin elimination within organisms (Ibelings and Havens, 2008,
Poste et al., 2011). It has also been raised that biotransformation of microcystin by aquatic organisms to
covalently-bound forms may complicate the complete measurement of total microcystin content in tissues
(Williams et al., 1997; Wilson et al., 2008; Dyble et al., 2011).
Levels of microcystins found in tissues of aquatic species potentially consumed by humans are shown in
Table 5-1. Unless specified, levels are reported as microcystin-LR equivalents. Most studies have
concentrated on levels in fish, although limited data show measurable amounts of microcystin-LR in
mussels, shrimp, and crayfish. Recent reviews emphasize that microcystin levels in edible fish and
shellfish are highly variable depending on trophic level, bloom conditions, and potential for depuration
(Ibelings and Chorus, 2007; Ferrao-Filho et al., 2011, and Kozlowsky-Suzuki, 2011). Soares et al (2004)
reported that microcystins could still be found in the fish muscle several days after the end of a toxic
bloom. In fish, higher concentrations were consistently measured in liver compared with muscle, which is
a significant dietary contribution in small fish consumed whole. Reports of deaths of marine mammals
from microcystin intoxication related to trophic transfer through marine invertebrates have been reported
(Miller et al., 2010). Deaths of 21 southern sea otters close to river mouths contaminated with
microcystins were related to intoxication after consuming farmed and free-living marine clams, mussels
and oysters in the area showing significant biomagnification (up to 107 times ambient water levels).
There have been no documented cases of microcystin toxicity in humans following ingestion of fish or
shellfish that have been exposed to microcystins (Mulvenna et al., 2012). Since food web exposures to
Health Effects Support Document for Microcystins - June, 2015
18
-------
blooms can vary greatly between geographical regions, it is unlikely to have year-round exposure in
humans that may consume aquatic organisms from water bodies susceptible to cyanobacterial blooms.
Data regarding microcystin elimination in fish are limited. A study of common carp (Cyprinus carpio)
and Silver Carp (Hypophtalmichthys molitrix) in Europe found that microcystins were completely
eliminated within one to two weeks from muscle and hepatopancreas after transferring the fish to clean
water (Adamovsky et al., 2007). The mean elimination half-lives ranged from 0.7 to 2.8 days in silver
carp muscle and from to 3.5 to 8.4 days in common carp liver. However, slower elimination (15 to 40
days after the end of the accumulation period), was reported in silver carp and Nile tilapia by Soares et al.
(2004).
Health Effects Support Document for Microcystins - June, 2015
19
-------
Table 4-1. Bioaccumulation Studies of Microcystins in Fish, Shellfish, and Crustaceans
Species/tissue
Tissue Concentration
Sampling Conditions
Average Water: Tissue
Correlations
Reference
Fish
Tilapia
Muscle
Liver
Viscera
0.002-0.337 |jg/g ww
0-31.1 |jg/g ww
0-71.6 |jg/g ww
3-year sampling from coastal lagoon;
seston concentrations ranged from 0-980
|jg/L during the study period
19.6 |jg/L:0.02|jg/g muscle
17 |jg/L:0.03|jg/g muscle
4.7 |jg/L:0.03|jg/g muscle
Magalhaes et al.,
2001
Tilapia
Muscle
Liver
0.007-0.06 |jg/g
0.092-0.28 |jg/g
Average levels from laboratory feeding of
isolated cells
14.6 [jg/fish/day (28 days):0.08
|jg/g muscle (peak)
Soares et al.,
2004
Fish - muscle
0.0396 |jg/g ww
Peak level in samples from bay over 11
months
0.78 |jg/L:0.0396 |jg/g muscle
Magalhaes et al.,
2003
Cyprinus carpio
Muscle
Hepatopancreas
0.038 |jg/g fresh wt
0.261 |jg/g fresh wt
Laboratory feeding bloom scum at 50 [jg/kg
body weight for 28 days
See previous columns
Li et al., 2004
Corydoras paleathus and
Jenynsia multidentata
Muscle
Liver
Gill
0.04-0.11 |jg/g ww
1.62-19.63 |jg/g ww
0.56-1.40 |jg/g ww
Laboratory exposure to 50 |jg microcystin-
RR/L for 24 hours
See previous columns
Cazenave et al.,
2005
Odontesthes bonariensis
Muscle
Liver
Gill
(average/maximum)
0.05/0.34 |jg/g ww
0.16/1.01 |jg/g ww
0.03/0.10 |jg/g ww
Wild caught from cyanobacteria containing
reservoir; cellular microcystin-RR = 41.59
|jg/g (wet season) and 9.65 |jg/g (dry
season)
Maximum tissue levels
correlated to wet season
Cazenave et al.,
2005
8 species
Muscle
Liver
Intestine
1.81 |jg/g dw
7.77 |jg/g dw
22 |jg/g dw
Wild caught from lake during bloom; 240
|jg/g dry weight of bloom sample
Water not sampled; ingestion
by fish possible
Xie et al., 2005
Yellow perch
Muscle
Liver
0.00012-0.004 |jg/g dw
0.017-1.182 |jg/g dw
Wild caught from lake during summer
months;
0.00016 - 4.28 |jg/L in seston
Data presented graphically;
positive correlation
Wilson et al.,
2008
4 species
Muscle
Liver
0.002-0.027 |jg/g dw
0.003-0.150 |jg/g dw
Wild caught from lake during August; Total
MC (-RR, -YR, -LR) in scum = 328 |jg/g dry
weight
See previous columns; tissue
concentrations varied by
species
Chen et al., 2009
2 species
Muscle
Liver
0.005-0.157 |jg/g
0.094-0.867 |jg/g
Commercial catch from lake with bloom;
0.02-0.36 |jg/L in seston; 0.16-0.19 |jg/L in
water
Samples not matched to fish
Berry et al., 2011
Multiple
Muscle
Whole
0.0005-1.917 |jg/g ww
0.0045-0.215 |jg/g ww
Multiple temperate and tropical lakes; 0.1-
57.1 |jg/L in water for all lakes
See paper, multiple fish
samples from all lakes
Poste et al.,
2011
Health Effects Support Document for Microcystins - June, 2015
20
-------
Species/tissue
Tissue Concentration
Sampling Conditions
Average Water: Tissue
Correlations
Reference
3 species
Muscle
Liver
-------
4.3 Occurrence in Dietary Supplements
Extracts from Arthrospira (Spirulina spp.) and Aphanizomenon flos-aquae (AFA) have been used as
dietary bluegreen algae supplements (BGAS) (Funari and Testai, 2008). These supplements are reported
to have beneficial health effects including supporting weight loss, and increasing alertness, energy and
mood elevation for people suffering from depression (Jensen et al., 2001). In children, they have been
used as an alternative, natural therapy to treat attention deficit hyperactivity disorders (ADHD).
Studies suggest that BGAS can be contaminated with microcystins ranging from 1 jxg/g up to 35 jxg/g
(Dietrich and Hoeger, 2005). Heussner et al. (2012) analyzed 18 commercially available BGAS for the
presence of toxins. Neither anatoxin-a nor cylindrospermopsin were found in any of the supplements.
However, all products containing AFA tested positive for microcystins at levels < 1 (ig microcystin-LR
equivalents/g dw. The microcystin (microcystin-LR with traces of microcystin-LA) was assumed to be
the result of contamination.
The levels of algal toxins in food supplements are unregulated at the federal level in the United States.
Therefore, it is difficult to appropriately evaluate the actual exposure to cyanobacterial supplements.
4.4 Summary
Microcystins have been detected in soil, at the surface or several centimeters below the surface, where
they play a functional role in nutrient cycling. They have also been found in sediments, edible plants, and
aquatic animals. Cyanobacterial cells and toxins can contaminate spray irrigation water and subsequently
be transmitted to food plants. Since water contaminated with cyanotoxins used for spray irrigation of crop
plants will inhibit plant growth and will induce visible toxic effects (e.g. brown leaves), affected plants
and crops will most likely not be used for eating purposes. Further investigation is needed to understand
the uptake and fate of microcystins and other cyanobacterial toxins by food plants. Bioaccumulation in
aquatic animals occurs mostly in the liver of fish, shellfish and crustaceans, but microcystins have also
been detected in fish tissue. After fish are exposed, concentrations of microcystins decrease with time as a
result of detoxification and depuration processes. The health risk from consumption depends on the
bioaccumulation of toxins in edible fish tissue compared to toxins in organs such as the liver. Currently,
cases of microcystin toxicity in humans following ingestion of fish or shellfish exposed to microcystins
have not been documented. Microcystin-LR has been detected in algal supplements at levels at or lower
than l(ig microcystin-LR equivalents/g dw.
Health Effects Support Document for Microcystins - June, 2015
22
-------
5.0 TOXICOKINETICS
5.1 Absorption
No data were available that quantified the intestinal, respiratory or dermal absorption of microcystin.
Most of the available evidence indicates that absorption from the intestinal tract and into liver, brain, and
other tissues requires facilitated transport using receptors belonging to the Organic Acid Transporter
polypeptide (OATp) family. The OATp family transporters are part of a large family of membrane
receptors that facilitate cellular, sodium-independent uptake and export of a wide variety of amphipathic
compounds including bile salts, steroids, drugs, peptides and toxins (Cheng et al., 2005; Fischer et al.,
2005; Svoboda et al., 2011). OATps are located in the liver, brain, testes, lungs, kidneys, heart, placenta
and other tissues of rodents and humans (Cheng et al., 2005; Svoboda et al., 2011). Only a few of the
OATps have been characterized at their functional, structural, and regulatory levels. In mice, males often
express OATps in tissues to a greater extent than females (Cheng et al., 2005).
For this document the abbreviation for the Organic Acid Transporter polypeptides will be written as
OATp rather than differentiating the animal versions from the human versions by using lower case letters
for the animals and upper case letters for humans.
5.1.1 Oral Exposure
An in situ study in rats indirectly studied the oral bioavailability of microcystin-LR using isolated
intestinal loops (Dahlem et al., 1989). After receiving a single 5 mg/kg infusion of microcystin-LR (>95%
pure) into the ileum, the rats showed clinical signs, including labored breathing and circulatory shock, as
well as evidence of liver toxicity within 6 hours. When an infusion of a similar dose was given into a
jejunal loop, a lower degree of liver toxicity was observed. The authors suggested site-specificity in
microcystin-LR intestinal absorption although the authors did not consider differences in absorptive
surface area when their hypothesis on differences in absorptive capacity was proposed.
A study done in swine demonstrated oral absorption of 3H-dihydromicrocystin (75 |_ig/kg) using ileal loop
exposure (Stotts et al., 1997a,b). The maximum blood concentration of the toxin occurred 90 minutes
after dosing.
Oral absorption of microcystin-LR (purified from an algal bloom sample) after a single gavage dose of
500 |ig/kg was examined by Ito et al. (1997a) and Ito and Nagai (2000). Microcystin-LR was absorbed
primarily in the small intestine, although some absorption was observed in the stomach as demonstrated
by targeted immunostaining (Ito and Nagai, 2000). The authors observed an erosion of the surface
epithelial cells of the small intestine villi facilitating perhaps the uptake of the toxin into the bloodstream
(Ito and Nagai, 2000; Ito et al., 1997a).
5.1.2 Inhalation Exposure
Microcystins can be present as aerosols in surface waters and drinking water after they are generated by
the wind or during showering or swimming providing contact with the respiratory epithelium. After an
intratracheal instillation in mice of a 50 j^ig/kg sublethal dose or a 100 j^ig/kg lethal dose, pulmonary
absorption of microcystin-LR (purified from an algal bloom sample) observed as immunostaining of the
lung occurred within 5 minutes (Ito et al., 2001). After the lethal dose was administered, a lag period of
60 minutes occurred and staining was observed in the liver after 7 hours of the sublethal dose
administration. This observation demonstrated the possibility of uptake from the lungs into systemic
circulation.
Health Effects Support Document for Microcystins - June, 2015
23
-------
Low levels of total microcystins (detection limit = 0.08 ng/m3) were detected in air samples collected
above a lake bloom, indicating that inhalation exposure was possible (Backer et al., 2008). However,
recreational users of the lake at the time of the bloom had no detectable microcystin in their blood and did
not report an increase in symptoms after spending time on the lake.
5.1.3 Dermal Exposure
In vivo or in vitro studies to determine the dermal absorption of microcystin have been identified. Skin
patch testing was done on 19 human volunteers using lyophilizedM aeruginosa (Stewart et al., 2006b).
Up to 170 ng of cyanotoxin was applied to filter paper discs applied to the back of each volunteer; patches
were removed after 48 hours and the exposed skin was scored after 48 and 96 hours. No individual
developed clinically detectable skin reactions.
5.2 Distribution
Facilitated transport is apparently necessary for both uptake of microcystins into organs and tissues as
well as for their export. In the liver, microcystins compete with bile acid uptake such that blocking this
transport system also prevents microcystin-LR uptake and toxicity in hepatocytes (Thompson and Pace,
1992). In vitro or in vivo exposures have shown that inhibition of microcystin uptake by its OATp
transporter could eliminate or reduce the toxicity in the liver (Runnegar et al., 1981, 1995a; Runnegar and
Falconer, 1982; Hermansky etal., 1990a,b).
In a study done by Fischer et al. (2005) human OATplA2, OATplBl, and OATplB3 demonstrated the
ability to mediate the transport of 3H-dihydromicrocystin-LR in Xenopus laevis oocytes. Inhibition of the
uptake was done by sulfobromophthalein and taurocholate. In addition, various in vitro studies have
shown that cells without microcystin-competent OATp do not absorb microcystin and that the
introduction of OATps to these cells will allow them to absorb microcystin (Komatsu et al., 2007,
Jasionek et al., 2010, Feurstein et al., 2010, Fischer et al., 2010). Another study by Fischer et al. (2010),
found that the role of OATp in microcystin uptake varies by congener and that highest uptake rates were
observed in MC-LW and MC-LF in comparison with microcystin-LR and microcystin-RR.
A study done by Lu et al. (2008) used OATplb2 null mice to demonstrate the importance of the OATp
system in transporting microcystin-LR into the liver The authors found severe hepatotoxicity and death
that was caused in wild-type mice after the intraperitoneal (i.p) administration of 120 |ig microcystin-
LR/kg. Fischer et al. (2010) used primary human hepatocytes and compared OATp-transfected HEK293
cells and control vector HEK293 cells (resistant to microcystin cytotoxicity) to show the need for
microcystin-competent OATp for transporting of microcystin across the cellular membrane. The primary
human hepatocytes were an order of magnitude more sensitive than the OATp-transfected HEK293 cells,
probably because HEK293 cells only have OATp lb 1 and lb3, while other OATps that contribute to the
uptake of the microcystin congeners may be in the primary human hepatocytes. Another study observed
similar results (Komatsu et al., 2007), however, microcystin-LR accumulation in OATp-transfected
HEK293 cells increased in a dose-dependent manner, which was not observed in the control vector
HEK293 cells.
5.2.1 Oral Exposure
The distribution of microcystin-LR (purified from an algal bloom sample) following oral gavage
administration to mice (500 j^ig/kg) was investigated using immunostaining methods (Ito and Nagai,
2000). Microcystin-LR was detected in large amounts in the villi of the small intestine. Erosion of the
villi was observed, which may have enhanced absorption of the toxin into the bloodstream. Microcystin-
LR was also present in the blood plasma, liver, lungs, and kidneys.
Health Effects Support Document for Microcystins - June, 2015
24
-------
Once inside the cell, microcystins covalently bind to cytosolic proteins, resulting in their retention in the
liver. The hepatic cytosolic proteins that bind microcystin have been identified as the protein phosphatase
enzymes 1 and 2A (PP1 and PP2A). Covalent adducts of microcystin-LR, microcystin-LA, and
microcystins-LL with both enzymes were identified by reverse-phase liquid chromatography. In contrast,
the dihydromicrocystin-LA analog did not form covalent bonds with PP1 and PP2A which suggests a role
for the double bonds of Adda in covalent binding. However, the dihydromicrocystin analog was able to
inhibit the enzyme activity, supporting a role for electrostatic interactions in the mode of action (MOA)
for enzyme inhibition as well as covalent binding; the IC50 was similar for microcystin-LR and the
dihydro-analog (Craig et al., 1996).
Nishiwaki et al. (1994) demonstrated that the distribution of 3H-dihydromicrocystin-LR in mice differs by
route of exposure. When 3H-dihydromicrocystin-LR (11.4 jj.Ci/2.4 mmol/0.2 mL saline) is administered
by intraperitoneal injection (i.p.), rapid and continuous uptake by the liver is observed, with around 72%
of the dose in the liver after 1 hour of administration. Total radiolabel in small percentages was observed
in various organs: 1.4% in the small intestine; 0.5% in the kidney and gallbladder; 0.4% in the lungs;
0.3% in the stomach. When 3H-dihydromicrocystin-LR (22.8 jj.Ci/2.1 (imol/0.2 mL saline) was
administered orally, much lower concentrations were observed in the liver, with less than 1% of the dose
in the liver at either 6 hours or 6 days after administration. Approximately 38% of the dose was found in
the gastrointestinal contents.
Microcystin-LR was not detected in the milk of dairy cattle exposed to M. aeruginosa cells either
administered by drinking water (detection limit= 2 ng/L) (Orr et al., 2001), or by ingestion of a gelatin
capsule with the cells (detection limit= 0.2 ng/L) (Feitz et al., 2002). Microcystins were not detected in
the blood plasma and only 10-39% of the total ingested microcystin-LR was found in the liver of beef
cattle given M aeruginosa cells via drinking water for 29 days (Orr et al., 2003). However, these studies
were limited by study design and data reporting (e.g. lack of controls, low number of cows and exposure
doses, or no concentrations of microcystin reported).
5.2.2 Inhalation Exposure
The organ distribution after intratracheal instillation of a lethal dose (100 j^ig/kg) of microcystin-LR
purified from an algal bloom was assessed by using immunostaining methods (Ito et al., 2001). The
kidney, liver, lung, and small intestine were positively stained for microcystin-LR. After 5 minutes of
instillation, intense staining was observed in the lung, in the kidney after 10 minutes, in the small intestine
after 45 minutes, and in the liver after an hour. Bleeding began around the hepatic central vein after 90
minutes of instillation. According to the authors, the pathological changes in the liver were the same as
those seen following oral or i.p. injection exposure routes. After intratracheal instillation of a sublethal
dose of 50 |_ig/kg. the authors observed immunostaining of the liver, kidney, lung, cecum and large
intestine but no obvious pathological changes were observed (Ito et al., 2001).
5.2.3 Other Exposure Routes
Studies in female rats have investigated the organ distribution of the i.v. administration of 2 |_ig of 125I-
labeled heptapeptide toxin (MW 1019) isolated from M. aeruginosa (Falconer et al., 1986; Runnegar et
al., 1986). High-performance liquid chromatography (HPLC) was used to purify the heptapeptide toxin
prior to reaction with 125I in the presence of Nal and lactoperoxidase. After 30 minutes, the liver and
kidney showed the highest tissue concentrations; 21.7% in the liver and 5.6% in the kidneys. The authors
reported 7% of the dose administered in the gut contents, and 0.9% cleared in the urine, with no
significant accumulation in other organs or tissues (Falconer et al., 1986).
Extensive liver uptake in mice was reported by Brooks and Codd (1987) after i.p. injection of 125 j^ig/kg
of a 14C-labelled toxin extracted fromM aeruginosa strain 7820. After 1 minute, 70% of the radiolabel
Health Effects Support Document for Microcystins - June, 2015
25
-------
was found in the liver, and after 3 hours increased to almost 90%. The kidneys, lungs, ileum, heart, large
intestine, and spleen also showed radiolabeled accumulation.
Robinson et al. (1989) determine the distribution of 3H-dihydromicrocystin-LR (>95% pure) after i.p.
injection of a sublethal dose of 45 j^ig/kg, or a lethal dose of 101 j^ig/kg in mice. Similar tissue distribution
of radiolabel (as % of total radioactivity) was observed after administration of both doses and after 60
minutes, accumulation in the liver from both doses reached a maximal value of 60%. For the lethal dose
(101 (.ig/kg). the radiolabel accumulation was 56% in the liver, 7% in the intestine, and 0.9% in the
kidney. Less than 1% was found in the heart, spleen, lung and skeletal muscle. In another study, Robinson
et al. (1991) observed distribution of microcystin-LR in mice within one minute of a sublethal i.v.
injection of 35 j^ig/kg to the liver, intestines, kidneys, plasma, and carcass (body minus the liver, gut,
kidney, heart, lung, and spleen). After one hour, the liver had around 67% of the dose, which remained
the same for the 6 days of the study even though 24% of the dose was eliminated in the urine and feces.
After one hour of the administration, small percentages were found in the intestines (8.6%), the carcass
(6%), the kidneys (0.8%), and trace amounts were found in the plasma. Within 3 minutes, levels in the
lung were high but after 10 minutes they were not detected. There was measurable radiolabel in the
spleen.
The subcellular distribution of radioactivity in the liver demonstrated that approximately 70% of the
hepatic radiolabel was present in the cytosol. In vitro experiments showed that radiolabeled microcystin
in the liver was bound to high molecular weight cytosolic proteins (Robinson et al., 1991). The nature of
the binding was demonstrated to be covalent, saturable and specific for a protein with a molecular weight
of approximately 40,000. Binding was inhibited by okadaic acid (a potent inhibitor of serine/threonine
phosphatases [1 and 2A]), suggesting that the target protein is protein phosphatase 1 or 2A. Binding
proteins for microcystin-LR were found in cytosol derived from several different organs, suggesting that
liver specificity is not due to limited distribution of target proteins. Covalent binding to hepatic proteins
may be responsible for the long retention of microcystin in the liver.
Rapid uptake of pure microcystin-LR into the serum was observed after i.p. injection of 35 |_ig/kg
(sublethal dose) to 24 mice (Lin and Chu, 1994). The samples were analyzed by direct competitive
ELISA and found that by 2 hours of administration, microcystin-LR reached a maximum concentration in
the serum, and after 12 hours in the liver cytosol, bound to liver cytosolic proteins. The kinetics of
binding was correlated by the authors with inhibition of protein phosphatase 2A activity. A maximum
decrease in enzyme activity was observed after 6 to 12 hours of dose injection.
Data from humans accidentally exposed to microcystin from dialysis water indicates that a large
proportion of microcystin in the serum and liver is bound to protein (Y uan et al., 2006). Three methods
were compared to detect microcystin in stored sera and liver samples from the exposed dialysis patients:
1) direct competitive ELISA using a polyclonal antibody against microcystin, which detects free
microcystin in a supernatant fraction; 2) liquid chromatography-mass spectrometry (LC/MS) after
oxidation and solid phase extraction to detect bound microcystin in a protein pellet fraction; and 3) gas-
chromatography-MS (GC/MS) after oxidation and solid phase extraction to detect total microcystin in a
sera or liver homogenate.
5.2.4 Liver Tissues - in vitro
Many researchers have examined the distribution to the liver using perfused liver and hepatic cell
cultures. Pace et al. (1991) demonstrated significant accumulation of 3H-dihydromicrocystin-LR in
isolated perfused liver despite a low overall extraction ratio (16% in liver, 79% in perfusate). In the liver,
radiolabel corresponding to microcystin-LR (15%) and a more polar metabolite (85%) was primarily
found in the cytosolic fraction.
Health Effects Support Document for Microcystins - June, 2015
26
-------
Primary rat hepatocytes in suspension and isolated perfused rat liver were used to evaluate the cellular
uptake of 3H-dihydromicrocystin-LR (Eriksson et al., 1990; Hooser et al., 1991a). Eriksson et al. (1990)
measured the uptake by scintillation counting of washed cells of a mixture of unlabeled microcystin-LR
and 3H-dihydromicrocystin-LR. Uptake was specific for freshly isolated rat hepatocytes and was inhibited
by the bile salts cholate and taurocholate, and by bile acid transport inhibitors such as antamanide,
sulfobromophthalein and rifampicin. Using both rat hepatocyte suspensions (four replicates from two rats,
two from each rat), and the isolated perfused rat liver (two rats), Hooser et al. (1991a) found that for the
first 5 to 10 minutes, the uptake of 3H-dihydromicrocystin-LR was rapid, followed by a plateau. The
uptake of 3H-dihydroMCLR was measured as radioactivity in fractionated cells versus radioactivity in
medium. At 0°C, the uptake was inhibited by incubation of suspended rat hepatocyte, probably by
involvement of an energy-dependent process. Inhibition of uptake was also observed by preincubation of
hepatocytes with rifampicin, a competitive inhibition of the bile acid transporter.
The dose level and exposure time in isolated rat hepatocytes on the uptake of 125I- microcystin-YM was
measured by Runnegar et al. (1991). Uptake was measured as radioactivity in centrifuged cell pellet.
Initially, hepatocyte uptake was rapid but after 10 minutes a plateau in the uptake rate was observed. In
the first minute of exposure, initial uptake rate increased with increasing concentration, however
cumulative uptake stopped at a dose causing plasma membrane blebbing.
Runnegar et al. (1995a), studied the microcystin-YM uptake by isolated rat hepatocytes using cell
associated radioactivity and assays for protein phosphatase inhibition in cell lysates. The authors found
that uptake was temperature-dependent and inhibited around 20-60% by in vitro preincubation with bile
acids or bile acid transport inhibitors such as trypan blue, taurocholate, cholate, cyclosporine A,
sulfobromophthalein, trypan red and rifamycin. This result indicates that uptake of microcystin happens
by carrier mediated transport. The pretreatment with protein phosphatase inhibitors such as okadaic acid
and calyculin A, inhibited both the uptake of microcystin-YM and the protein phosphatase, suggesting
that the protein phosphatase may have impacted the conformation or membrane presence of the OATp
transporter. Serine phosphorylation is involved in the regulation of hepatocyte OATplAl's transport
function (Svoboda et al., 2011).
After 2 to 3 days of being maintained in culture, the primary cultures of liver cells cease to express the
OATps. As a result, established liver cell lines are generally not suitable to evaluate microcystin toxicity
(Eriksson and Golman, 1993; Heinze et al., 2001). This was also observed by Chong et al. (2000) who
evaluated microcystin toxicity in eight rodent, primate and human permanent cell lines, and found that
after microcystin-LR exposure, only two showed cytotoxicity: a human oral epidermoid carcinoma KB
cells, and a rat Reuber H35 hepatoma H-4-II-E cells. Toxic response in these cells was most evident when
microcystin-LR was added after cells were seeded. Those cells more resistant to microcystin toxicity were
established monolayers cells.
Hooser et al. (1991a) also evaluated the subcellular distribution of 3H-dihydromicrocystin-LR in primary
rat hepatocytes in suspension and the isolated perfused rat liver. The authors found that after protein
precipitation with trichloroacetic acid, 50% of the 3H-dihydromicrocystin-LR was localized in the
cytosolic fraction and bound to cytosolic proteins, and 50% was found as free toxin. The authors
suggested that since 3H-dihydromicrocystin-LR did not bind significantly to actin or other cytoskeletal
proteins, little of the radiolabel was in the insoluble pellet containing insoluble actin and other elements
(Hooser et al., 1991a).
Studies on the binding of subcellular protein of 3H-dihydromicrocystin-LR in rat liver homogenates found
that around 80% of the radiolabeled toxin was bound to cytosolic proteins (Toivola et al., 1994). 3H-
dihydromicrocystin-LR shown to bind to both PP1 and PP2A. PP2A was detected primarily in the cytosol
and PP1 was found in the membrane proteins (mitochondrial and post-mitochondrial particulate fraction).
Health Effects Support Document for Microcystins - June, 2015
27
-------
5.3 Metabolism
Limited data are available on the metabolism of microcystins. Most of the studies discussed below
indicate that there is minimal if any catabolism (process of breaking down molecules into smaller units to
release energy). The microcystins can be conjugated with glutathione and cysteine to increase their
solubility and facilitate excretion (Kondo et al., 1996). It is not clear whether CYP450-facilitated
oxidation precedes conjugation. Stotts et al. (1997a,b) found that after i.v. injection or ileal loop exposure
in swine, 3H-dihydromicrocystin-LR was not metabolized in the liver and was primarily present in hepatic
tissues as the parent compound.
Some metabolism of microcystin-LR was shown to occur in mice and in isolated perfused rat liver
(Robinson et al., 1991; Pace et al., 1991). Male CD-I mice were administered 3H-dihydromicrocystin-LR
as an i.v. dose of 35 |_ig/kg and monitored for up to six days. Over the 6-day interval, 9.2% and 14.5% of
the dose was excreted in the urine and feces, respectively, of which -60% was parent compound. High-
performance liquid chromatography analysis for urinary and fecal metabolites revealed several minor
peaks of lower retention times. Analysis of liver cytosol preparations revealed that 83% of the radiolabel
was bound to a high molecular weight cytosolic protein after six hours and that amount decreased to 42%
by day 6 (Robinson et al., 1991). Pace et al. (1991) also demonstrated binding of both the parent toxin
(3H-dihydromicrocystin-LR) and a more polar metabolite to cytosolic proteins in isolated perfused rat
liver. Of the hepatic cytosol radiolabeled, 60 to 85% were polar metabolites. No characterization of
metabolites of microcystin-LR was done in these studies.
A decrease in the amount of cytochrome b5 and cytochrome P450 in the liver was observed after the
administration of 125 (ig/kg of Microcystis strain 7820 (primarily produces microcystin-LR) to mice
(Brooks and Codd, 1987). The pretreatment of mice with microsomal enzyme (mixed function oxidase)
inducers such as (3-naphthoflavone, 3-methylcholanthrene and phenobarbital, eliminated this effect on
hepatic cytochromes. Pretreatment also extended survival and reduced liver toxicity (i.e., changes in liver
weight). However, no change in cytochrome P450 associated enzyme activity (i.e., metabolism of
aminopyrene and p-nitrophenol) was found in microsomes isolated from mouse liver after animals were
injected with an extract ofM aeruginosa (Cote et al., 1986).
Glutathione and cysteine conjugates have been identified in the liver after i.p. injection of 10 or 20 |_ig
microcystin-RR to mice or 4(ig microcystin-LR to rats (purified from blooms) (Kondo et al., 1992, 1996).
The conjugates were isolated and compared to chemically prepared standards which indicated structural
modification of the Adda and Mdha moieties of the microcystin toxins. The authors postulated that these
moieties could be the sites of CYP oxidation and subsequent conjugation with glutathione or cysteine.
Formation of microcystin-LR glutathione conjugates occurs by glutathione S-transferase (GST) enzymes
found in both liver cytosol and microsomes of rats (Takenaka, 2001). Characterization of glutathione
conjugation of microcystin-LR (>95% pure isolated fromM aeruginosa) has been done by five
recombinant human GSTs (Al-1. A3-3, Ml-1, P101, and Tl-1) (Buratti et al., 2011). Although with
different dose-responses, all five GSTs catalyzed the conjugation. The authors also determined that the
spontaneous reaction for microcystin-LR conjugation with glutathione (GSH) was dependent on GSH
concentration, temperature and pH.
Based on LD50 estimates, Kondo et al. (1992) found that glutathione and cysteine conjugates of
microcystin-LR and microcystin-YR were less toxic than the parent compounds, however, they
demonstrated that these conjugates were toxic (LD50 values ranged from 217 to 630 |ig/kg in mice).
Metcalf et al. (2000) also demonstrated in vitro that glutathione, cysteine-glycine and cysteine conjugates
were less toxic in the mouse bioassay than the parent compounds demonstrating that conjugates were also
weaker inhibitors of protein phosphatases 1 and 2A. After intratracheal instillation in mice, the
distribution of glutathione and cysteine conjugates of microcystin-LR start in the kidney and continue in
Health Effects Support Document for Microcystins - June, 2015
28
-------
the intestine suggesting that in vivo, the lower toxicity of glutathione and cysteine conjugates may be
related to the distribution through excretory organs and elimination of metabolites (Ito et al., 2002a).
Ito et al. (2002b) synthesized glutathione and cysteine conjugates of microcystin-LR and administered
them by intratracheal instillation in mice. The metabolites were demonstrated to be less toxic than the
parent compound as shown by lethal doses about 12-fold higher than the microcystin-LR lethal dose. The
metabolites were distributed primarily to the kidney and intestine, as opposed to the liver (Ito et al.,
2002b).
Several studies have investigated the role of glutathione homeostasis and lipid peroxidation in
microcystin-induced liver toxicity (Ding et al., 2000a; Gehringer et al., 2004; Bouai'cha and Maatouk,
2004). Ding et al. (2000a) indicated that microcystin exposure in isolated hepatocytes resulted in an initial
increase in glutathione synthesis followed by a later depletion of glutathione. Gehringer et al. (2004)
suggest that increased lipid peroxidation induced by microcystins is accompanied by an increase in
glutathione peroxidase, transcriptional regulation of glutathione-S-transferase and glutathione peroxidase
and de novo synthesis of glutathione. Bouai'cha and Maatouk (2004) found that 2 ng/mL of microcystin-
LR in primary rat hepatocytes caused an initial increase in ROS formation and an increase in glutathione.
Additional details of the oxidative stress reaction to microcystins are given in section 6.4.5 Physiological
or Mechanistic Studies.
5.4 Excretion
Biliary excretion has been shown in both in vivo and in vitro studies. Falconer et al. (1986) administered
to female albino rats an i.v. dose of 2 |_ig of a peptide extracted from M. aeruginosa. The authors
demonstrated a biphasic blood elimination curve, with the first component with a half-life of 2.1 minutes
and a second component with a half-life of 42 minutes. After 120 minutes, the authors observed 1.9% of
the administered dose in the urine and 9.4% in the intestinal contents, suggesting biliary excretion of the
toxin. Pace et al. (1991) also observed biliary excretion in isolated perfused rat liver after 1.7% of
radiolabeled microcystin-LR was recovered in the bile after a 60-minute perfusion. Seventy-eight percent
of the radiolabel in the bile collected during the perfusion was associated with the parent toxin while the
rest of the radiolabel was associated with more polar metabolites (Pace et al., 1991).
In a study by Robinson et al. (1991), male VAF/plus CD-I mice were administered an i.v. dose of 35
(ig/kg of radiolabeled microcystin-LR. A biexponential plasma elimination curve was observed with
plasma half-lives of 0.8 and 6.9 minutes for the first and second phase of elimination, respectively. A total
of approximately 24% of the administered dose was eliminated in the urine (9%) and feces (15%) during
the 6-day study monitoring period. Around 60% of the excreted radiolabel in both urine and feces,
measured at 6 and 12 hours following injection, was present as the parent compound.
Elimination in swine was evaluated following i.v. injection or ileal loop exposure (Stotts et al., 1997a,b).
3H-dihydromicrocystin-LR was detected in the bile as early as 30 minutes after i.v. injection of 75 |_ig/kg.
After ileal loop exposure to the same dose, the toxin concentration in the portal venous blood was
consistently higher as compared to peripheral blood. The labeled microcystin-LR was rapidly eliminated
and followed a biphasic pattern in both the i.v. and ileal loop exposures, suggesting that the liver removes
the toxin rapidly from the blood. At higher dose levels, removal from the blood is slower, likely due to
the liver toxicity and circulatory shock observed at high doses.
Health Effects Support Document for Microcystins - June, 2015
29
-------
5.5 Pharmacokinetic Considerations
The blood half-life in female rats was measured following i.v. administration of a 125I-labelled
heptapeptide toxin extracted from M. aeruginosa (MW 1019, assumed to be a microcystin) (Falconer et
al., 1986). A biphasic blood elimination curve was demonstrated, with a half-life of 2.1 minutes in the
first component and a half-life of 42 minutes in the second component.
Microcystin-LR excretion was also evaluated in mice (Robinson et al., 1991). A biexponential plasma
elimination curve was demonstrated after i.v. injection of a 35 j^ig/kg sublethal dose of 3H-
dihydromicrocystin-LR. A plasma half-life of 0.8 minutes was observed in the first phase of elimination
and 6.9 minutes was reported for the second phase.
Stotts et al. (1997a,b) evaluated the toxicokinetics of 3H-dihydromicrocystin in swine following i.v.
injection and ileal loop exposure. Elimination of labeled microcystin-LR was rapid and followed a
biphasic pattern, suggesting that the liver rapidly removes the toxin from the blood. Clearance from the
blood is slower at higher dose levels, presumably due to the liver toxicity and circulatory shock observed
at high doses. It is important to take into consideration that tritium radiolabeling may alter the microcystin
molecule's ability to bind with protein phosphatases, thus altering microcystin protein binding and tissue
distribution profile (Hilborn et al., 2007).
No physiologically based toxicokinetic models have been developed for microcystins.
Health Effects Support Document for Microcystins - June, 2015
30
-------
6.0 HAZARD IDENTIFICATION
6.1 Human Studies
6.1.1 Epidemiology and Case Studies of Systemic Effects
Analysis of hepatic enzyme levels from a group of patients served by a public water supply contaminated
with a bloom of M. aeruginosa were compared with levels in patients living in areas served by other
water supplies not contaminated with the bloom (Falconer et al., 1983). Although 871 individual records
were examined, the number of exposed and unexposed were not reported. The authors used as the study
population those patients referred to a single hospital laboratory for liver function tests before, during and
after a bloom of M. aeruginosa in the Malpas Dam reservoir of Australia. The patients were classified as
those that used the reservoir for drinking water supply (Armidale residents), and residents of neighboring
towns with independent water supplies. Analysis of plasma enzymes (y-glutamyltransferase (GGT),
aspartate aminotransferase (AST), alanine aminotransferase (ALT), and alkaline phosphatase (ALP)) was
conducted. Liver function for each group was classified based on the liver enzyme testing date: during 5
weeks before the first signs of the bloom appeared; during the 3-weeks of the bloom or the 2 weeks after
copper sulfate treatment of the bloom; and during 5 weeks after the bloom. Copper sulfate addition was
identified as the high-risk time interval due to the cell lysis and subsequent toxin release. Differences in
enzyme levels between comparison groups and between times within comparison groups was analyzed
using analysis of variance. The authors observed a significant increase in GGT levels in residents of
Armidale during the bloom period, and an increase in ALT levels in the same group, although not
statistically significant.
Although the authors observed a difference in enzyme levels between the groups, the finding was
attributed to the imprecise method of selecting study participants (Falconer et al., 1983). Several of the
enzyme measurements for the Armidale residents were associated with one participant with chronic
kidney disease requiring a repeat of the analysis. Alcoholism was reported to occur in about the same
proportion, 7 to 10% in both groups assessed before and during the bloom, although in lower proportion
in the post-bloom group of Armidale residents. Alcoholism has been associated with an increase in GGT
levels. However, the authors concluded that these changes in GGT among Armidale residents before and
during the bloom period might potentially be associated with exposure to drinking water contaminated
with a M. aeruginosa bloom.
Turner et al. (1990) reported an outbreak among army recruits who had consumed reservoir water during
canoe exercises. The reservoir contained a bloom of cyanobacteria, primarily M. aeruginosa. Two
recruits, both 16 years old, had detailed case reports with history of malaise, sore throat, blistering around
the mouth, dry cough, pleuritic pain and abdominal pain. One of them experienced vomiting and diarrhea.
After physical examination, both patients presented fever, abdominal tenderness, and left basal pulmonary
consolidation (pneumonia). Within 24 hours and after treatment with antibiotics, temperature returned to
normal. Low platelet counts in both patients but no increases in liver enzymes was detected in blood tests.
Testing of various pathogens such as Leptospira, Legionella, Chlamydia, Coxiella, Mycoplasma and
influenza and adenovirus was negative. Sixteen soldiers that participated in the same canoe exercises also
reported similar symptoms including diarrhea, vomiting, sore throat, dry cough, headache, abdominal
pain, and blistered mouth.
Microcystins, including microcystin-LR was detected in a sample of the bloom taken the day after the
patients were admitted into the hospital (Turner et al., 1990). After two weeks, high levels of Escherichia
coli were also found in reservoir water. The authors suggested that exposure to microcystin may have
been related to the pulmonary consolidation and low platelet count of the two patients, citing evidence
from studies in mice. The potential role of other toxins in this event was not addressed.
Health Effects Support Document for Microcystins - June, 2015
31
-------
A cross-sectional study was done to evaluate the relationship between liver damage in children and
microcystin levels in drinking water and aquatic food (carp and duck) in China (Li et al., 201 la).
Microcystin concentrations were measured in three sources of drinking water used by local residents in
the Three Gorges Reservoir Region in China: a community well rarely contaminated with microcystin
(unexposed), a lake with an occasional cyanobacterial bloom (Lake 1), and a lake with regular
cyanobacterial blooms over the previous 5 years (Lake 2). Children from 5 schools were selected to
participate and those served by water from the wells for more than 5 years and rarely ate fish or duck
from the lakes (145 participants) were considered to have no exposure. Those with low exposures were
the children served by Lake 1 (183 participants) and those with high exposure were the children served
from Lake 2 (994 participants). A questionnaire was administered to the participants and blood samples
from approximately 50 children per exposure group were obtained for analysis of ALT, AST, GGT, ALP,
and mean serum microcystin levels.
Concentrations of microcystin were found to be below detection limit in the well water in all but one of
the six years tested (Li et al., 201 la). Only one year detected microcystins at 0.1 |ig microcystin-LR
equivalents/L. The average microcystin-LR equivalents/L over the 5 years in Lake 1 was 0.24 |ig/L and in
Lake 2 was 2.58 |ig/L. Levels of microcystin-LR (fish and ducks) were higher in the aquatic food from
Lake 2 than Lake 1. Based on consumption of drinking water and aquatic food, the authors estimated that
children served by Lake 1 (low exposure) consumed 0.36 (ig/day of microcystin-LR, while children in
Lake 2 (high-exposure) consumed 2.03 (ig/day. Mean serum levels of microcystin in the groups were
below detection in the unexposed group, 0.4 |ig microcystin-LR equivalents/L in the low-exposure group,
and 1.3 |ig microcystin-LR equivalents/L in the high-exposure group. The respective serum detection
rates were 1.9%, 84.2%, and 91.9% in the unexposed, low-exposed, and in the high-exposed groups,
respectively.
Exposure to microcystin in drinking water was associated with increases in AST and ALP, but no
increases in ALT or GGT were observed. The odds ratio (OR) for liver damage associated with
microcystin exposures was 1.72 with 95% confidence intervals (95% CI) of 1.05-2.76 (dichotomous
based on two or more abnormally elevated liver enzyme assays). According to the authors, Hepatitis B
infection, based on serum measurements of antigens and/or antibodies, was a greater risk for liver damage
than microcystin exposure among these children.
6.1.2 Other Routes of Exposures
In February 1996, there was an outbreak of acute liver failure in patients at a renal dialysis clinic in
Caruaru, Brazil (Carmichael et al., 2001, Jochimsen et al., 1998). One hundred and sixteen of 130 patients
who received their routine hemodialysis treatment at that time experienced headache, eye pain, blurred
vision, nausea and vomiting. Subsequently, 100 of the affected patients developed acute liver failure and,
of these, 76 died. A cohort study was conducted as well as an evaluation of the center water supply;
patient's serum, and postmortem liver tissue were analyzed for microcystin. Analysis of the carbon, sand,
and cation/anion exchange resin from in-house filters in the clinic's water treatment for microcystins and
cylindrospermopsin demonstrated the presence of both cyanotoxins (Azevedo et al., 2002). Analyses of
blood, sera, and liver samples from the patients revealed microcystins, but not cylindrospermopsin. The
method used to extract cylindrospermopsin from the samples may have been inadequate. Based on a
comparison of patient's symptoms and liver pathology with data from animal studies of microcystins and
cylindrospermopsin, the authors concluded that the major contributing factor to death of the dialysis
patients was intravenous exposure of microcystins.
Blood samples collected from 51 patients of the renal dialysis clinic in Caruaru, Brazil were analyzed
using ELISA (Hilborn et al., 2005; 2007). Microcystin concentrations ranged from less than 0.16 (ig/L
(limit of detection) to 28.8 (ig/L in serum samples. Additional analysis using GC/MS, in 6 serum samples
Health Effects Support Document for Microcystins - June, 2015
32
-------
found microcystin oxidized to MMPB (2-methyl-3- methoxy-4-phenylbutyric acid) ranging from 45.7 to
112.9 ng/mL. ELISA analysis of these serum samples detected free microcystin concentrations ranging
from 6.7 to 26.3 (ig/L. The authors concluded that both free and protein-bound microcystins were found
in human serum.
In another contamination event at a dialysis center in Rio de Janeiro, Brazil in 2001, microcystin
concentrations of 0.32 (ig/L were measured in the activated carbon filter used in an intermediate step for
treating drinking water to prepare dialysate (Soares et al., 2005). A concentration of 0.4 (ig/L was
detected in the drinking water. Serum samples were collected 31 to 38 days after microcystin-LR was
detected in water samples and patients were monitored for eight weeks. The presence of microcystins
indicated that 44 dialysis patients were potentially exposed to microcystin from contaminated dialysate
(Hilborn et al., 2013). A longitudinal study to characterize the clinicopathological outcomes among 13
dialysis patients was conducted and serum microcystin concentrations were quantified with ELISA.
Although the biochemical outcomes varied among the patients, markers of hepatic cellular injury
chlolestasis (elevations of AST, ALT bilirubin, ALP and GGT) in serum during weeks one to eight after
treatment frequently exceeded normal values. Concentrations of microcystin-LR in the serum ranged
from 0.46 to 0.96 ng/mL (Soares et al., 2005). Since microcystin was not detected during weekly
monitoring after the first detection, the authors suggested that the patients were not continuously exposed
to the toxin and that the toxin detected in the serum after eight weeks may have been present in the form
of bound toxin in the liver (Soares et al., 2005). Results were consistent with mild to moderate mixed
liver injury. Although the patients in the study had pre-existing diseases, the direct intravenous exposure
to dialysate prepared from surface drinking water supplies made them at risk for cyanotoxin exposure and
resultant adverse effects (Hilborn et al., 2013).
6.2 Animal Studies
6.2.1 Acute Toxicity
6.2.1.1 Oral Exposure
Fitzgeorge et al. (1994) administered microcystin-LR via gavage to newly weaned CBA/BALBc mice
weighing 20±1 g. Sex of the mice and the number used per dose group were not reported. Deaths were
recorded within two hours of dosing. The commercially-obtained compound was described only as
"suitably purified". The LD50 was estimated to be 3,000 |ig/kg. and increases in liver (43%) and kidney
(5.9%) weights were reported. The authors reported that there was no change in lung or spleen weight;
dose-response data and other endpoints were either not examined or not reported.
Acute oral toxicity of purified microcystin-LR (>95% pure by HPLC) in female BALB/c mice was
evaluated by Yoshida et al. (1997). Previous studies using doses of 16.8 and 20 mg/kg resulted in death
within 160 minutes in two mice. Therefore, to determine the LD50, the authors administered via gavage to
seven 6-week-old mice 0, 8.0, 10.0 and 12.5 mg/kg doses of microcystin-LR in saline solution. Within 24
hours, the mortality was 0/2 in controls, 0/1 at 8 mg/kg, 0/2 at 10 mg/kg and 2/2 at 12.5 mg/kg. The oral
LD50 was identified as 10.0 mg/kg.
Light microscopy was used to examine the liver, kidneys and lung and electron microscopy was used to
identify apoptotic cells in the livers of treated mice (Y oshida et al., 1997). Histopathological analysis was
performed on the remaining tissues. The only effects observed were in the liver and kidneys; no effects
were observed on the stomach, intestine, skin or organs after histophatological evaluation were observed.
In those animals that died, liver effects included centrilobular hemorrhage and hepatocyte degeneration.
In those mice administered doses greater than 12.5 mg/kg, free hepatocytes in the veins of mice were
Health Effects Support Document for Microcystins - June, 2015
33
-------
observed. In the previous study, those mice receiving doses of 16.8 and 20 mg/kg, showed proteinaceous
eosinophilic materials in the Bowman's spaces in the kidneys.
One of the surviving mice at 10.0 mg/kg was sacrificed after 24 hours. Hepatocellular necrosis was
observed in the centrilobular and midzonal regions, and in the centrilobular region and surrounding
necrotic areas, single cell death (possibly apoptotic) was reported. All other mice treated with 10 mg/kg
and the two mice treated with 8.0 mg/kg were sacrificed one week after treatment. The authors observed
livers with hypertrophic hepatocytes in the centrilobular region and fibrosis in the centrilobular and
midzonal regions. In addition, a few apoptotic cells were observed in these animals. Kidney effects were
not reported in those animals that survived treatment for at least 24 hours.
A comparison between the acute effects of microcystin-LR on the livers and gastrointestinal tracts of
young and aged mice was done by Ito et al. (1997a). A single dose (500 |ig/kg) was administered to aged
(29 mice age 32 weeks) and young (12 mice age 5 weeks) male ICR mice. The microcystin-LR (purity
not specified) was dissolved in ethanol and diluted in saline and administered via oral gavage. The
controls were 3 aged and 3 young untreated mice. After 2 hours of treatment, 23 aged mice were
sacrificed, five mice at 5 hours, and two mice at 19 hours, and 4 young mice were sacrificed at each time
point. Evaluation of liver damage and gastrointestinal erosion were performed.
The authors observed that the effects in the aged mice were more severe than those in the young mice. No
liver pathology or gastrointestinal changes were reported in young mice. However, 18 of 29 aged mice
treated with the same dose showed pathological changes of the liver, some of them (8) showed liver
injury of the highest severity (severity rating of +4), characterized as bleeding, disappearance of many
hepatocytes in the whole liver and friable tissue. Other aged mice also showed liver injury of different
severity: 5 of 29 mice had severity rating of +3 characterized by bleeding and disappearance of
hepatocytes in centrilobular region; 4 of 29 mice had necrosis in the centrilobular region (severity rating
of +2), and one mouse had eosinophilic changes in the centrilobular region (severity rating of+1).
Other effects in aged mice included gastrointestinal effects characterized by necrosis to one-third depth of
the mucosa and severe duodenal damage including separation of epithelial cells from lamina propria,
decreased villi density, and edema of both the submucosa and villi. Aged mice also showed thinning of
gastrointestinal epithelial cells with consequent exposure of lamina propria and glands in some areas.
Although details of the incidence of these effects were not reported, the authors indicated that the degree
of liver injury was related to the severity of gastrointestinal effects. At 5 to 19 hours after treatment,
regeneration of intestinal tissues was evident in some of the mice sacrificed. No difference was observed
on the enzyme levels (AST and ALT) among untreated aged mice.
The effect of single oral gavage doses of microcystin-LR was studied by Fawell et al. (1999). Doses of
500, 1,580 and 5,000 |ig/kg body weight of microcystin-LR (commercial product; purity not specified) in
aqueous solution were administered to five male and five female CR1:CD-1(ICR)BR(VAF plus) mice
and CR1:CD(SD)BR(VAF plus) rats. No untreated control group was included in the study. After 14
days, animals were sacrificed, necropsy was performed, and microscopic examinations of the lung and
liver were conducted. The LD50 value in mice was estimated at 5,000 j^ig/kg and in rats was >5,000 |_ig/kg.
Signs of hypoactivity and piloerection (involuntary bristling of hairs) were observed in those animals that
died (Fawell et al., 1999). However, no clinical signs were observed in survivors. Those that survived,
showed no signs of body weight changes during the 14-day follow-up. Darkly discolored and distended
livers, as well as pallid kidneys, spleen, and adrenals were observed at necropsy in those animals that
died. The livers had moderate or marked centrilobular hemorrhage. Rats and mice of all dose groups
showed diffuse hemorrhage in the liver, however the incidence was not clearly related to dose. Table 6-1
summarize the incidence and severity of liver lesions observed in the study.
Health Effects Support Document for Microcystins - June, 2015
34
-------
Table 6-1. Incidence of Liver Lesions in Mice and Rats After Exposure to Microcystin-LR (Fawell et
al., 1999)
Mice
Rats
(10 per group)
(10 per group)
500
1580
5000
500
1580
5000
|jg/kg
|jg/kg
|jg/kg
|jg/kg
|jg/kg
|jg/kg
Mortality
0
1
5
0
0
1
Diffuse Hemorrhage
2
1
1
8
7
8
Moderate Centrilobular Hemorrhage
0
2
7
0
0
1
Marked Centrilobular Hemorrhage
0
1
0
0
0
1
Centrilobular Necrosis
0
0
2
0
0
1
Cytoplasmic Vacuolation
0
0
0
0
0
1
A comparison between the acute oral effects of microcystin extracts in young and aged mice was done by
Rao et al. (2005). A single dose of microcystin-LR (3.5 g extract/kg from laboratory cultures ofM
aeruginosa which corresponded to 9.625 mg microcystin-LR/kg) was administered to aged (36 weeks
old) and young (6 weeks old) male Swiss albino mice. After 4 to 5 hours, mortality first occurred with the
mean time to death significantly shorter in the aged mice. In comparison to the control groups, both
groups of mice had an increased relative liver weight and DNA fragmentation. No difference between the
age groups was observed. However, a significantly greater difference in glutathione depletion and lipid
peroxidation was observed in the aged mice when compared with young mice. Although most serum
enzymes were increased over controls in both groups, GGT was increased to a greater extent in aged mice
than in young mice.
6.2.1.2 Inhalation Exposure
No studies of acute inhalation exposures were identified. The microcystins are not volatile; therefore
inhalation exposures are likely to only occur in the form of aerosols. A brief abstract describes a study of
acute microcystin-LR exposure via inhalation (Creasia, 1990). Details of study design and results were
not reported. The LC50 for mice exposed to a microcystin-LR aerosol (nose only) for 10 minutes was
reported to be 18 |ig/L (mg/m3) with a 95% confidence interval of 15.0-22.0 |ig/L (mg/m3). Based on
studies of lung deposition after exposure of mice to the LC50 concentration, an LD50 of 43 |ig/kg body
weight was estimated. The authors reported that histological lesions in mice killed by aerosol exposure
were similar to those in mice dosed intravenously with microcystin-LR.
Fitzgeorge et al. (1994) conducted experiments in newly weaned CBA/BALBc mice (20±1 g) with
microcystin-LR (commercial product; purity not stated) administered either by intranasal instillation or
aerosol inhalation. Few details of study design and findings were given. A single experiment with mice
(number unspecified) inhaling a fine aerosol (particle size 3-5 |im) of 50 (ig microcystin-LR/L for an
unspecified duration of time did not result in any deaths, clinical signs of toxicity or histopathological
changes. The nature of the examinations was not reported. The authors estimated the delivered dose of
microcystin-LR to be very small (about 0.0005 |ig/kg). The LD50 for intranasal instillation of microcystin-
LR was equal to 250 (ig/kg. All deaths occurred within two hours of dosing. Liver and kidney weights
were increased by 41.6 and 7.5%, respectively, in the animals (n = 6; sex not specified) receiving the
LD50 of microcystin-LR intranasally. The estimated LD50 of intranasal instillation, 250 (.ig/kg. is the same
as the LD50 of i.p. exposure, which is much lower than the LD50 of gastric intubation (3000 |ig/kg).
Fitzgeorge et al. (1994) further evaluated the relationship between dose and liver weight increase after
intranasal instillation of microcystin-LR to newly weaned CBA/BALBc mice (20±1 g; assumed n = 6). At
single intranasal doses of 31.3, 62.5, 125, 250 and 500 (.ig/kg. liver weight increased proportionally (0,
Health Effects Support Document for Microcystins - June, 2015
35
-------
1.5, 24.4, 37.4 and 87%). Seven daily intranasal doses of 31.3 (.ig/kg, resulted in a liver weight increase of
75%. The authors reported histopathological findings, but failed to specify which findings resulted from
single doses and which resulted from the multiple-dose experiment reported in the same publication.
Findings included necrosis of respiratory and olfactory epithelium in the nasal mucosa and centrilobular
necrosis with hemorrhage in the liver. Early changes in the liver included vacuolar degeneration and
necrosis of hepatocytes near the central vein. The adrenal glands showed effects as well with vacuolation
and necrosis of the inner cortex and congestion of medullary blood vessels. No histopathological changes
were observed in the trachea, lungs, esophagus, pancreas, spleen, lymph nodes, kidneys or brain.
Several studies demonstrated the potential for uptake from the respiratory system using intratracheal or
intranasal instillation. Ito et al. (2001) evaluated the distribution of purified microcystin-LR after
intratracheal instillation of lethal doses in male ICR mice and included a limited description of toxic
effects. Microcystin-LR in saline solution was instilled at doses of 50, 75, 100, 150 and 200 |ig/kg into 34
mice; three mice were sham-exposed as controls. Mortality was 100% in 12 mice receiving doses of 100
|ig/kg and greater. At 75 (.ig/kg, two of four mice died, while no deaths occurred in 18 mice given 50
|ig/kg intratracheally.
The time course of hepatotoxicity was further evaluated in eight mice given an intratracheal dose of 100
(ig/kg (Ito et al., 2001). One mouse was sacrificed at each of 5, 10, 20, 30, 45, 60, 90 and 120 minutes.
Immunostaining for microcystin-LR showed the toxin in the lungs within 5 minutes and in the liver after
60 minutes. Hemorrhage in the liver was observed after 90 minutes and became severe by 120 minutes.
6.2.1.3 Dermal/Ocular Exposure
No studies evaluating the effects in animals of dermal or ocular exposure to purified microcystins were
identified. Cyanobacteria bloom samples collected from five different lakes or ponds were tested for
allergenic and irritative effects in guinea pigs and rabbits, respectively (Torokne et al., 2001). The
microcystin content (presumed to be total LR, RR, and YR) ranged from 0.1-2.21 mg/g. To determine
sensitization, guinea pigs were initiated with an intradermal injection followed seven days later by topical
application at the injection site. Sensitization was moderate to strong in 30-67% of guinea pigs but did not
correlate with microcystin content. All samples produced only negligible to slight skin and eye irritation
on rabbits.
6.2.1.4 Other Routes
The acute toxicity of microcystins following i.p. injection has been studied in mice and rats. The LD50 for
microcystin-LR in mice ranges between 30 and 60 |ig/kg (Lovell et al., 1989; Slatkin et al., 1983; Gupta
et al., 2003; Rao et al., 2005). The LD50 for microcystin-LR was slightly higher in fed rats (122 |ig/kg)
compared to fasted rats (72 |ig/kg) (Miura et al., 1991) suggesting possible higher uptake by cells in the
absence of competing dietary substrates.
The available studies demonstrate a very steep dose-response curve for microcystin-LR acute toxicity
following i.p. administration. In female mice, the only change observed at 50 |ig microcystin-LR/kg was
Kupffer-cell hyperplasia, while all mice receiving 100 |ig/kg died (Hermansky et al., 1991). A sublethal
dose of about 25 |ig/kg in male mice resulted in a significant increase in liver weight (8.7%) but no
clinical signs or hepatic lesions (Lovell et al., 1989). A single injection of 60 |ig/kg of microcystin-LR in
male mice caused liver injury within 12 hours as indicated by increases in ALT and AST, intrahepatic
hemorrhage and destruction of the liver morphology (Weng et al., 2007). Liver toxicity was also assessed
in male mice administered a single dose of 55 |ig/kg of microcystin-LR (Wei et al., 2008). Animals were
sacrificed at various times from 0.5-12 hours after exposure. Histopathology revealed liver toxicity
beginning at 6 hours including severe, intrahepatic hemorrhage and destruction of hepatic structure.
Health Effects Support Document for Microcystins - June, 2015
36
-------
Gupta et al. (2003) determined mean LD50 values of 43, 235.4 and 110.6 mg/kg for microcystin-LR,
microcystin-RR and microcystin-YR, respectively, using groups of 4 mice per dose (doses not specified).
The microcystins were dissolved in methanol and diluted to the test concentrations with phosphate
buffered saline. The time to death varied considerably with microcystin-RR being the least toxic. A
significant increase in liver body weight index was induced by all of the congeners. Serum levels of AST,
ALT and y -GT increased significantly, compared with controls, as early as 30 minutes post exposure for
all congeners. The acute LD50 determination for these three congeners showed a difference in toxicity
with microcystin-LR being the most toxic followed by microcystin-YR and microcystin-RR. The findings
from this study supports the hypothesis that as the hydrophilic properties of the amino acids increase, the
toxicity decreases.
Both microcystin-YR and microcystin-RR have lower acute toxicity in mice than microcystin-LR with
LD50 estimates of 111 and 171 |ig/kg for microcystin-YR and 235 and 650 j^ig/kg for microcystin-RR
(Gupta et al., 2003; Stotts et al., 1993). The difference in LD50 for microcystin-YR compared to
microcystin-RR is consistent with the higher in vitro cellular toxicity of microcystin-YR using a human
colon carcinoma cell line (Caco-2) (Puerto et al., 2009).
6.2.2 Short-Term Studies
6.2.2.1 Oral Exposure
The effects of microcystin-LR (commercial product; purity not stated) on 11-week-old male hybrid rats
(F1 generation of female WELS/Fohm x male BDIX) after drinking water exposure was evaluated by
Heinze (1999). For 28 days, three groups of 10 rats each received doses of 0, 50 or 150 (.ig/kg body
weight of microcystin-LR in drinking water. Daily measurements of water consumption and rat weights
were done at weekly intervals. Over the 28-day period, 3 to 7% of supplied water was not consumed; the
dose estimates provided by the authors were not adjusted to account for the percentage of incomplete
drinking water consumption. After 28 days of exposure, rats were sacrificed and organ weights (liver,
kidneys, adrenals, thymus and spleen), were recorded. Hematology, serum biochemistry and
histopathology of liver and kidneys were also evaluated.
A 38% increase in the number of leukocytes in rats in the highest dose group was observed after
hematological evaluation (Heinze, 1999). Serum biochemistry showed significantly increased mean levels
of ALP and lactate dehydrogenase (LDH) in both treatment groups; 84% in LDH and 34% in ALP in low
dose group, and 100% increase in LDH and 33% increase in ALP in high dose group. No changes in
mean levels of ALT or AST were observed. A dose-dependent increase in relative liver weights was
observed in both dose groups: 17% at the low dose group and 26% at the high dose. Table 6-2 shows the
mean enzyme levels and the relative and absolute liver weights.
A dose-dependent increase in absolute liver weight in both dose groups was also observed and provided
by the author in a personal communication. The average absolute liver weights were 8.8 grams in the
control group, 9.70 grams in the lower dose and 10.51 grams in the high dose. No statistically significant
changes in other organ weights or body weights were reported and no effects on the kidneys were
observed. The incidence of liver lesions is summarized in Table 6-3. Lesions were spread diffusely
throughout the parenchyma and included increased cell volume, increased mitochondria, cell necrosis,
activation of Kupffer cells and increased amounts of periodic acid-Schiff (PAS)-positive substances,
indicating cell damage. Liver lesions were observed in both treatment groups, but the severity of the
damage was higher in the high dose group (150 |ig/kg). The low dose (50 (ig/kg/day) was the LOAEL for
effects on the liver.
Health Effects Support Document for Microcystins - June, 2015
37
-------
Schaeffer et al. (1999) reported the results of a study in w hich A. flos-aqiicte. a cyanobacterium consumed
as a food supplement, was fed to mice in the diet. The authors used recent analysis of A. flos-aqiicte,
which often coexists w ith Microcystis species, to estimate the microcystin content in the material
consumed by the mice. Analysis of the A. flos-aqiicte samples used in the feeding study showed an
average concentration of 20±5 |ig microcystin-LR per gram of A. flos-aqiicte. The authors estimated the
daily exposure of microcystin-LR in the exposed mice to range from 43.3 (ig/kg-day to 333.3 (ig/kg-day.
No clinical signs of toxicity were reported, and no effects on mortality, body weight, organ weights or
histology were observed in the treated mice. In addition, no effects on reproductive parameters were
reported in five treated mice from the highest dose group allowed to breed and there was no effect on
growth and organ function in fetal and neonatal mice. The 333.3 (ig/kg-day dose was the NOAEL under
the conditions of the study.
Table 6-2. Relative Liver Weights and Serum Enzyme Levels in Rats Ingesting microcystin-LR in
Drinking Water (Heinze, 1999)
Control
(Mean ± SD)
50 |jg/kg
(Mean ± SD)
150 |jg/kg
(Mean ± SD)
Serum Enzymes
Alkaline phosphatase (ALP) (microkatals/L)
9.67 ±2.20
13.00 ± 3.81*
12.86 ± 1.85*
Lactate dehydrogenase (LDH)
(microkatals/L)
16.64 ±4.48
30.64 ± 5.05*
33.58 ± 1.16*
Liver Weight
Relative (g/100 g body weight)
2.75 ± 0.29
3.22 ± 0.34*
3.47 ± 0.49*
Absolute (g)**
8.28 ± 1.37
9.70 ± 1.32
10.51 ± 1.02
* p<0.05 when compared with control
Table 6-3. Histological Evaluation of the Rat Livers After Ingesting Microcystin-LR in drinking
Water (Heinze, 1999)
Activation of
Kupffer Cells
Degenerative and
Necrotic
Hepatocytes with
Hemorrhage
Degenerative and
Necrotic
Hepatocytes
without
Hemorrhage
PAS-positive
Material
Control
Slight
0
0
0
1
Moderate
0
0
0
0
Intensive damage
0
0
0
0
50 |jg/kg
Slight
0
4
0
5
Moderate
10
6
0
5
Intensive damage
0
0
0
0
150 |jg/kg
Slight
0
0
0
0
Moderate
10
6
1
8
Intensive damage
0
3
0
2
The effects of orally administered microcystin-RR on apoptosis in the liver of adult male ICR mice were
evaluated by Huang et al. (2011) (see also section 6.4.5.1.3). For 7 days, doses of 0, 4.6, 23, 46, 93, or
186 |ig/kg body weight of microcystin-RR (commercial product; purity not stated) were administered to
groups of 5 mice viagavage. Animals were sacrificed after 7 days of exposure and DNA fragmentation
Health Effects Support Document for Microcystins - June, 2015
38
-------
was evaluated with the terminal deoxynucleotidyl transferase-mediated dUTP-biotin nick end-labeling
(TUNEL) assay. Analysis of PP2A activity was done with Western blot for B cell lymphoma/leukemia-2
(Bcl-2), Bcl-2 associated x protein (Bax), p53 expression, C/EBP homologous protein (CHOP), and
glucose-related protein 78 (GRP78).
A dose-dependent increase in the percent of apoptotic cells in the liver was observed for all the dose
groups: 10.46% for the 0 dose; 12.6% for the 4.6 dose; 12.7% for the 23 dose; 30.3% for the 46 dose;
28.5% for the 93 dose; and 37.5% forthe 186 |ig/kg dose group. The only doses with statistical
significance were >46 (.ig/kg. A significant increase in Bax protein expression was observed at 46 and 93
|ig/kg and in p53 protein expression was observed at 93 (.ig/kg. Bcl-2 was significantly decreased with
doses >23 |ig/kg but the Bax/Bcl-2 ratio was significantly increased at the same dose. No significant
changes were found in CHOP protein expression. GRP78 protein expression was significantly increased
only at the 93 |ig/kg dose and none of the other doses were different from the control (including the high
dose). No changes in PP2A activity or alterations in PP2A A subunit mRNA expression were seen for any
dose groups.
6.2.2.2 Inhalation Exposure
Groups of six male BALB/c mice were exposed for 30, 60 or 120 minutes each day for seven consecutive
days to monodispersed submicron aerosols of 260-265 (ig/m3 microcystin-LR via nose-only inhalation
(Benson et al., 2005). The dose deposited in the respiratory tract were estimated to be 3, 6 and 12.5 |ig/kg
body weight/day. The control mice were exposed to 20% ethanol in water (aerosolized vehicle). Clinical
signs were recorded daily and sacrifice of mice occurred the day after the last exposure. Blood and serum
were collected and analysis for blood urea nitrogen [BUN], creatinine, total bilirubin, ALP, AST, ALT,
total protein, albumin and globulin. Histopathological examination of the liver, kidney, spleen, thymus,
respiratory tract tissues, adrenals, gastrointestinal tract and testes was conducted and organ weight
(adrenals, lung, liver, kidney, spleen and thymus) were recorded. Histopathological evaluation of the
epithelium lining the bone structure of the nasal passages (turbinates) from different locations was also
done.
Table 6-4. Incidence and Severity of Nasal Cavity Lesions in Mice After Inhalation of
Microcystin-LR
Nasal Cavity Lesions
Severity Grade
Daily Exposure Period (minutes)
Control
30
60
120
Turbinate 1 (immediately caudal to the upper incisors)
Respiratory Epithelial
Necrosis
Not noted
6/6
5/6
0/6
4/6
Minimal
0/6
1/6
0/6
0/6
Mild
0/6
0/6
6/6
0/6
Moderate
0/6
0/6
0/6
2/6
Respiratory Epithelial
Inflammation
Not noted
6/6
5/6
6/6
5/6
Mild
0/6
1/6
0/6
1/6
Olfactory Epithelial
Degeneration, Necrosis
and Atrophy
Not noted
6/6
6/6
6/6
1/6
Mild
0/6
1/6
0/6
4/6
Moderate
0/6
0/6
0/6
1/6
Turbinate 2 (at the dose of the incisive papilla)
Respiratory Epithelial
Necrosis
Not noted
6/6
6/6
0/6
0/6
Mild
0/6
0/6
6/6
3/6
Moderate
0/6
0/6
0/6
3/6
Health Effects Support Document for Microcystins - June, 2015
39
-------
Nasal Cavity Lesions
Severity Grade
Daily Exposure Period (minutes)
Control
30
60
120
Respiratory Epithelial
Inflammation
Mild
6/6
5/6
6/6
6/6
Moderate
0/6
1/6
0/6
0/6
Olfactory Epithelial
Degeneration, Necrosis
and Atrophy
Mild
6/6
6/6
0/6
0/6
Moderate
0/6
0/6
6/6
0/6
Marked
0/6
0/6
0/6
6/6
Turbinate 3 (at the dose of the first upper molar)
Olfactory Epithelial
Degeneration, Necrosis
and Atrophy
Not noted
6/6
6/6
0/6
0/6
Mild
0/6
0/6
6/6
0/6
Moderate
0/6
0/6
0/6
4/6
Marked
0/6
0/6
0/6
2/6
From Benson et al., 2005
No clinical signs or effects on body or organ weights were observed after exposure to microcystin-LR
aerosol (Benson et al., 2005). Histopathological examination revealed treatment-related lesions only in
the nasal cavity. Lesions were not observed in the liver or in any other organs or parts of the respiratory
tract. The authors observed an increase of nasal lesions and severity with length of the daily exposure
period (Table 6-4). The nasal cavity lesions observed included necrosis or inflammation of respiratory
epithelial cells and degeneration, and necrosis and atrophy of olfactory epithelial cells. Necrotic lesions of
olfactory epithelial cells were generally larger patches whereas few cells were involved in respiratory
epithelial cell necrosis.
6.2.2.3 Other Routes
Male BALB/c mice were given 0, 40, or 50 j^ig/kg of microcystin-LR via i.p. injection, once a day for 10
days (Sun et al., 2011). The microcystin-LR was purified in the authors" laboratory, but the purity was not
stated. No deaths were observed at 40 |_ig/kg. while 5/10 animals died after seven days at 50 |_ig/kg.
Pretreatment each day with sulforaphane (an antioxidant found in cruciferous vegetables) prevented
death. Groups of three male Sprague-Dawley rats were administered purified microcystin-LR (purity not
stated) for 28 days via intraperitoneal implantation of osmotic pumps. The pumps were filled with
microcystin-LR diluted in saline that delivered 0, 16, 32, or 48 (ig/kg/day (Guzman and Solter, 1999). No
significant differences were observed between the groups for body weight gain, liver-to-body weight
ratio, and food consumption. Histopathology of the liver revealed necrosis, apoptosis and the presence of
cytoplasmic vacuoles in mid- and high-dose animals and evidence of hepatic inflammation in high-dose
animals. Livers from the mid- and high-dose animals had significantly higher levels of malondialdehyde
(3-4 fold) and tissue slices in culture released greater amounts of ALT compared to controls. Hepatic
ALT activity significantly decreased as its release from the liver tissues increased. There was a dose-
related increase in tissue AST that reached significance for only the high doses. The fact that there were
only 3 animals per dose group is a limitation of this study; the gradual infusion of microcystin-LR
through the use of an osmotic pump is a positive feature of the study design.
6.2.3 Subchronic Studies
6.2.3.1 Oral Exposure
Dried bloom extract with at least seven microcystin congeners with the major peak tentatively identified
as microcystin-YR (no peak could specifically be identified as microcystin-LR), was administered in the
drinking water of pigs (n= 5/group) for 44 days (Falconer et al., 1994). Pigs were administered 0, 80, 227,
or 374 mg of dried algae/kg body weight per day. A decrease in body weight was observed in pigs in the
Health Effects Support Document for Microcystins - June, 2015
40
-------
highest dose group perhaps due to reduced food and/or water consumption at this dose. Dose- and time-
dependent increases in GGT, ALP and total bilirubin, as well as a decrease in plasma albumin were
observed in plasma samples collected over 56 days. Dose-related changes were also observed in the
incidence and severity of histopathological changes of the liver, including Kupffer cell proliferation,
periacinar degeneration, cytoplasmic degeneration, hepatic cord disruption, single cell necrosis, and
congestion. Since exposure was via the dried algae, the study does not identify a NOAEL or LOAEL for
microcystin.
Fawell et al. (1999) reported the results of a subchronic toxicity study of microcystin-LR given via
gavage to Crl :CD-1(ICR)BR (VAF plus) mice (age and body weight not specified). Microcystin-LR was
obtained commercially (purity not stated) and administered in distilled water. The concentration in the
dosing solution was verified by HPLC with UV detection. Groups of 15 male and 15 female mice were
administered daily oral doses of 0, 40, 200 or 1000 jxg/kg body weight for 13 weeks. Eye examinations
were conducted prior to and at the conclusion of treatment, body weight and food consumption were
recorded weekly, and clinical observations were made daily. During the final week of treatment,
hematology and serum biochemistry were evaluated for seven mice of each treatment group. After 13
weeks, the authors performed gross examination of organs and microscopic evaluation of tissues. Lungs,
liver and kidney were examined only in the treated animals. All other tissues were examined in the
control and high dose animals.
At 1,000 (.ig/kg. one female was found dead during week 1 and one male was found moribund and
sacrificed during week 13; a cause of death was not given and both animals appeared to be included in the
histopathology analyses. No treatment-related clinical signs of toxicity were observed throughout the
study. No dose-related trends were evident for body weight gain or body weight in males (data not
reported). The study authors stated that mean body weight gain was decreased approximately 15% in all
treated male groups and was statistically significant at 40 or 200 (ig/kg-day (p<0.05). However, no
quantitative data for these effects were presented in the published paper. Data tables obtained from the
author showed that the mean body weight gain differed from controls by the same amount for all the
exposed dose groups (2 g) and thus lacked a dose-response (Fawell, Personal Communication, 2015).
Mean terminal body weights differed from controls by about 7% in these groups. The only body weight
change observed in females was an increase in body weight gain in the 200 (ig/kg-day group. No dose-
response was observed for body weights or body weight gain. The only body weight change observed in
females was an increase in body weight gain in the 200 (ig/kg-day group.
A slight (10-12%) increase in mean hemoglobin concentration, red blood cell count and packed cell
volume among females receiving 1000 |ig/kg body weight was observed after hematological evaluation.
In the high-dose males, ALP, ALT and AST levels were significantly elevated (2- to 6-fold higher), and
only ALP and ALT were elevated (2- and 6-fold higher, respectively) in high dose females. In the mid-
dose males, ALT and AST were also elevated (2-fold). All treatment groups showed a slightly decreased
GGT. In males of the mid- and high-dose groups, serum albumin and protein were reduced (13%). Table
6-5 shows the clinical chemistry results.
In the males and females of the mid- and high-dose groups, a dose-related increase in incidence and
severity of histopathological changes in the liver were reported (Fawell et al., 1999). The liver lesions
reported were multifocal inflammation with deposits of hemosiderin and hepatocyte degeneration
throughout the liver lobule. The incidence of these liver histopathological changes are summarized in
Table 6-6. Chronic inflammation demonstrated the clearest dose response across all dose groups for both
sexes. Hepatic degeneration showed a steep response to dose for the mid and high dose groups. Sex-
related differences in liver pathology were not apparent. No lesions were found in other tissues. The
NOAEL was 40 (ig/kg/day and the LOAEL 200 (ig/kg/day for liver histopathology and elevated serum
levels of ALT and AST in males.
Health Effects Support Document for Microcystins - June, 2015
41
-------
Table 6-5. Serum Biochemistry Results for Mice Treated with Microcystin-LR for 13 Weeks
microcystin-
LR Dose
((jg/kg-day)
Blood Chemistry Results
(Mean ± Standard Deviation)
Albumin
(g %)
Alkaline
Phosphata
se (ALP)
(U/L)
Alanine
Aminotransfera
se (ALT) (U/L)
Aspartate
Aminotransferas
e (AST) (U/L)
Gamma
Glutamyl
Transaminas
e (GGT) (U/L)
Total
Protein
(g %)
Female
Control
3.1 ±0.14
167±24.6
32±11.3
101 ±38.3
4±1.0
5.1±0.30
40
3.2±0.16
187±76.2
25±7.8
74±13.2
3±0.5
5.2±0.28
200
3.4a±0.1
4
156±33.4
27±9.4
74±22.1
3±0.0
5.3±0.31
1000
3.1 ±0.18
339b±123.7
220b± 149.1
144±71.7
3±0.4
5.1±0.22
Male
Control
3.2±0.19
91±22.2
27±8.0
68±27.7
6±1.0
5.5±0.32
40
3.0±0.13
95±29.2
37±17.2
64±12.2
4±0.7
5.1±0.26
200
2.8C±0.13
94±32.3
59a±28.0
121b±43.7
3C±0.4
4.8b±0.29
1000
2.8C±0.11
232b±103.2
159c±75.0
121b±26.3
4±0.4
4.8C±0.21
From Fawell et al., 1999; Significantly different from controls at: a p<0.05; b p<0.01; c p<0.001
Table 6-6. Liver Histopathology in Male and Female Mice Treated with Microcystin-LR for 13
Weeks
Liver Histopathology
Control
(n=15)
40 [jg/kg-day
(n=15)
200 [jg/kg-day
(n=15)
1000 [jg/kg-day
(n=15)
Female
Chronic inflammation
5
8
8
14
Hepatocyte vacuolation
5
5
11
8
Hemosiderin deposits
0
0
1
14
Hepatocyte degeneration
0
0
1
9
Male
Acute inflammation
0
1
0
0
Chronic inflammation
1
2
4
15
Hepatocyte vacuolation
5
5
6
3
Hemosiderin deposits
0
0
0
15
Hepatocyte degeneration
0
0
1
14
From Fawell et al., 1999
6.2.3.2 Inhalation Exposure
No data from subchronic inhalation exposure of animals were found.
6.2.4 Neurotoxicity
Impaired long-term memory retrieval, as assessed by a step-down inhibitory avoidance task, was reported
in rats after receiving an intrahippocampal injection of 0.01 or 20 |ig/L of a microcystin extract from
Microcystis strain RST 9501 (Maidana et al., 2006). Impaired spatial learning in the radial arm maze was
also observed after exposure to 0.01 |ig/L. but exposure at the higher concentration did not. An increase in
oxidative damage, as measured by lipid peroxides and DNA damage, was observed in tissue homogenates
of the hippocampus from treated animals.
Health Effects Support Document for Microcystins - June, 2015
42
-------
Feurstein et al. (2011) examined the effects of microcystin-LR on isolated murine cerebellar granule
neurons after administration of 5 |iM microcystin-LR. Cell viability was significantly decreased but
apoptosis was not induced by the concentrations given (up to 5 (iM). Capase-3/7 activity was not
increased with concentrations up to 5 |iM but slight impairment of the neurite network was observed in
the cells incubated for 48 hours at concentrations higher than 1 |iM microcystin-LR. A significant dose-
related decrease in neurite length was observed at concentrations ranging from 1 to 10 |iM along with
serine/threonine-specific PP inhibition and sustained hyperphoshorylation of Tau.
After intrahippocampal injection of 1 or 10 |ig/L of microcystin-LR (>98% pure), Li et al. (2012) reported
impaired memory function, assessed by Morris water maze, in male rats. Rats showed an increased
latency to find the platform after injection of both concentrations of microcystin-LR. Only at 10|ig/L.
histology of the brain revealed neuronal damage in the CA1 region of the hippocampus. In the same
region (CA1) only high-dose animals showed a significant decrease in the total number of cells and the
density of cells, but not in the cell volume. At both concentrations, malondialdehyde (MDA) levels and
catalase activity in the hippocampal CA1 region were increased, but superoxide dismutase (SOD) and
glutathione peroxidase activity were only significantly increased at 10 |ig/L.
Li et al. (2014) suggested impairment of spatial learning and memory in groups of male 28 day old
Sprague Dawley rats after oral exposure to microcystin-LR (95% pure) for 8 weeks. The microcystin-LR
was dissolved first in methanol (1 mL/mL) and diluted with 100 mL pure water (0.001% methanol-v/v).
The microcystin-LR was subsequently diluted with pure water to concentrations of 0.2, 1 and 5 (ig/ml. As
a result the methanol concentration of the stock solution increased with the microcystin-LR concentration.
Rats were dosed with 0.2, 1.0, and 5.0 |ig/kg of microcystin-LR every two days and performance in the
Morris water maze test was evaluated. Pure water was used as the control. A weakness in the preparation
of the dosing solution in this study is the fact that as the microcystin-LR dose increased, so did the
methanol dose; thus the animals in each dose group received increasing amounts of methanol as well as
microcystin.
At the conclusion of the dosing period, the animals were trained to find the platform within the water
maze for 6 days. The group that received the 5 j^ig/kg dose took a significantly longer time to find the
platform on day 3 of the training, but was comparable to the animals in the other dose groups by day 6. In
the memory component of the test, after removal of the platform, the treated rats from the highest two
dose groups spent less time in the platform quadrant than the controls. However, these differences were
not statistically significant. At the higher dose (5 |ig/kg), the treated rats showed the activation of
astrocytes in the hippocampus and a dose-related increase of hippocampal nitric oxide synthase (NOS) as
reflected by the number of N-20+ cells and detected by immunostaining of tissues from four animals from
each dose group. Nitric oxide (NO) concentration was measured directly using the supernatant from the
hippocampus tissue homogenate. The increase in NO concentration was also dose-related. Both NO and
NOS were significantly higher (p<0.05) at the high dose.
A subsequent study by Li et al. (2015) examined neurological responses in pups born to dams that had
been exposed to normalized doses of 0, 0.5, 2.5, or 10 |ag microcystin-LR/kg/day for 8-weeks before
mating, but not during gestation or lactation. A description of the developmental portion of this study is
provided in Section 6.2.5. The control and microcystin-LR solutions each contained 0.002 % (v/v)
methanol normalized to 0.001% methanol to account for dosing every other day over the 8 week period.
The liters were culled to 4 males and 4 females per dam where possible. At specific postnatal time periods
the pups were subjects to tests of motor function as follows:
• PND 7: surface righting reflex, negative geotaxis, cliff avoidance;
• PND 28: open field test, Morris water maze learning;
• PND 60 open field test and Morris water maze memory;
Health Effects Support Document for Microcystins - June, 2015
43
-------
Twenty four hours after each behavioral test, one male and one female from each dam was sacrificed and
the brain prepared for histological examination. The hippocampal tissues were analyzed for byproducts of
lipid peroxidation (MDA and Total SOD).
Both the males and females had significantly (p<0.07) lower cliff avoidance performance than the
controls on PND 7 at all doses (Li et al., 2015). There was a dose-related trend for the males but not the
females. There were no significant differences from controls in the negative geotaxis, surface righting
reflex tests or in the open field tests on PND 28 and PND 60. During the water maze training period
(PND-28), there were no differences between groups. However, during the water maze memory tests the
males in all groups scored more poorly than the females. The swimming speed for the females was
significantly decreased for the mid and high dose groups. There was a significant increase in hippocampal
MDA in the normalized 2.5 and 10 |_ig/kg/day dosed males and in both males and females at 10
(ig/kg/day. Both males and females had a significant increase in measures of total hippocampal SOD at 10
Mg/kg/day.
Given the neurotoxic properties of methanol, the presence of methanol in the solution in this study makes
it difficult to evaluate these results as they relate to exposure to microcystin-LR in finished drinking
water. A solution with a normalized 0.001% methanol (v/v) is equivalent to a concentration of 7.9 mg/L2.
The data on intubation volumes for the dams are not provided in the published paper, thus it is not
possible to quantify the methanol dose to the dams. In addition, exposure of the dams to methanol plus
microcystin-LR ceased before mating and conception, making it difficult to quantify the relationship
between the dosing of the dams and the exposures of the pups in the absence of data on half- life for
microcystin-LR and methanol. Based on the postnatal pup responses the NOAEL was 0.2 (ig/kg/day and
the LOAEL was 2.5 (ig/kg/day based on the results of the Morris water maze tests and the analysis of the
hippocample tissues for evidence of ROS. Male pups were more sensitive than the females. The
possibility of synergy between methanol and microcystin-LR cannot be eliminated. The apparent lack of
direct exposure to the dams during conception, gestation and lactation, also confounds the application of
the data to a risk assessment for microcystin-LR in drinking water.
6.2.5 Developmental/Reproductive Toxicity
6.2.5.1 Reproductive Effects
6.2.5.1.1 Oral Exposure
Kirpenko et al. (1981) used extracts fromM aeruginosa from a reservoir during the summer months to
determine reproductive toxicity in rats. Male and female white rats (total of 120 rats) were intubated with
5 x 10-4 or 5 x 10-7 mg/kg of toxin extracts (no details were provided on the content of the extract) or 10
mg/kg ofM aeruginosa biomass for three months (dosing procedure not specified). Histopathology of
the ovaries as well as estrous cyclicity and microscopic studies of the genital appendages and testes in
males were conducted. After 3 months of treatment with 5 x 10"4 mg/kg of toxin extract or 10 mg/kg of
biomass, and absence in the estrous cycle, (specifically an absence of estrus with prolonged diestrus) was
observed. Maturation and growth of the oocytes was also affected. After 1.5 months of treatment with 5 x
10"4 mg/kg of the toxin extract, degeneration of oocytes in Graafian vesicles, decreased follicle
dimensions, and increased number of involuted corpora lutea were observed. In males, there was a
decrease in spermatogonia quality, spermatozoid motility, living spermatozoids (increased dead), and
spermatid quality with a dose of 5 x 10~4 mg/kg of toxin extract. Spermatogonia are stem cells in the walls
of the seminiferous tubules that give rise to spermatocytes, an intermediate step in the formation of
2 0.001% v/v = 0.001 ml/lOOml x 0.79 g/ml (density of methanol) = 0.00079 g/lOOml x 1000 mg/g = 0.79 mg/100
mL = 7.9 mg/L methanol
Health Effects Support Document for Microcystins - June, 2015
44
-------
spermatozoa. Histological evaluation revealed "epithelium shelled out" (not defined) from basal
membranes and greater tubule deformation. Degenerating spermatogonia and morphological
abnormalities in Sertoli cells were also noted.
Falconer et al. (1988) used extract from anM aeruginosa bloom sample to study the reproductive effects
of microcystin in mice. Eight female mice received l/4th dilution of the extract, estimated to contain 14
(ig/mL of unspecified microcystin toxin, as drinking water since weaning. The mice were mated at age 20
weeks with similarly treated males for a premating treatment interval of approximately 17 weeks. The
authors did not observe a difference in number of litters, pups per litter, sex ratio, or litter weight. Seven
of 73 pups from treated parents showed reduced brain size. None of the 67 pups from controls showed
reduced brain size. The authors did not report the litter distribution of the affected pups. After histological
examination of one of the small brains, extensive damage to the hippocampus was observed.
Sperm quality and testicular function were assessed in male specific pathogen free mice (0.015-0.025 kg
at purchase) administered microcystin-LR (commercial product; purity not stated) in the drinking water at
concentrations of 0, 1, 3.2, or 10 (ig/L for 3 or 6 months (Chen et al., 2011). Microcystin-LR was
dissolved in 0.1% methanol and diluted to the required concentration with water; controls received water
only. Although body weight and amount of water consumed were measured, these data were not
presented and doses to the animals were not calculated by the study authors. Based on the subchronic
reference drinking water value of 0.0078 L/day and body weight of 0.0316 kg for the male B6C3F1
mouse (U.S. EPA, 1988), doses to the animals were estimated at 0, 0.25, 0.79, and 2.5 |_ig/kg. Subchronic
reference values were chosen to more accurately reflect status of the animals after 6 months of treatment;
based on growth curves for the B6C3F1 mouse and initial body weights of the SPF mice, the B6C3F1
strain was considered reasonably similar to the strain used in this study.
No clinical signs of toxicity were observed and body weight, testes weight, and water consumption were
not affected by treatment. Results of sperm and hormone analyses are shown in Table 6-7. No significant
changes in any parameter were noted at 1 (.ig/L. At 3.2 and 10 (ig/L, sperm counts were significantly
decreased and sperm motility was reduced at 3 and 6 months with severity increasing with longer duration
of exposure. Animals in the mid- and high-dose groups had a trend towards lower serum testosterone and
higher luteinizing hormone and follicle stimulating hormone after 3 months, which was statistically
significant by 6 months (except for FSH in the mid-dose group). Histopathological evaluation of the
testes showed a slightly loosened appearance of the organization of the epithelium in the seminiferous
tubules at 10 (ig/L after 3 months. After 6 months, slight testicular atrophy associated with sparse
appearance of the seminiferous tubules was found at 3.2 and 10 (ig/L with dose-related increased severity.
The animals given 10 (ig/L also showed loss and derangement of spermatogenic cells, enlargement of the
lumen of the seminiferous tubules, thinning of the spermatogenic epithelium, as well as depopulation of
Leydig cells, Sertoli cells, and mature sperm. The number of apoptotic cells in the testes was increased at
10 (ig/L after 3 months and at 3.2 and 10 (ig/L after 6 months. The NOAEL was 0.25 (ig/kg/day and the
LOAEL was 0.79 (ig/kg/day.
6.2.5.1.2 Other Routes
Ding et al. (2006) studied the effects of microcystin on the reproductive system of male mice
administered 0, 3.33, or 6.67 |ig microcystin/kg i.p. daily from an extract of Microcystis (99.5%
microcystin-LR, 66.476 |ig/mL. and 0.5% microcystin-YR, 0.361 |ig/mL) for 14 days using 0.9% saline
as the vehicle. A significant decrease in body weight gain in both treatment groups was observed during
the course of the study. A dose-dependent decrease in absolute testes weight was observed, but a
significant increase in relative testes weight was observed only in the high-dose group. The high-dose
group had an increase in the percent immobile sperm and a significant decrease in absolute and relative
epididymis weight. A dose-dependent decrease in sperm viability and the proportion of sperm with rapid
progressive motility was observed. No increase in the percent of abnormal sperm was recorded.
Health Effects Support Document for Microcystins - June, 2015
45
-------
Histological evaluation of both treatment groups showed atrophy of the seminiferous tubules with
increased spacing between the seminiferous tubule cells and the effect increased with increasing dose.
The high-dose group also exhibited a decreased number of interstitial cells, Sertoli cells, and mature
sperm in the seminiferous tubules, and a deformation of Leydig and Sertoli cells.
Table 6-7. Serum Hormone Levels and Sperm Analyses From Mice Given Microcystin-LR in the
Drinking Water for 3 or 6 Months
Endpoint
0 (jg/kg/day
0.25 |jg/kg/day
0.79 |jg/kg/day
2.5 |jg/kg/day
3 months
Testosterone (ng/mL)
2.23 ± 1.15
2.77 ± 0.93
2.34 ± 1.11
1.07 ± 0.27
LH (mlU/mL)
7.03 ± 0.41
7.28 ± 0.66
8.05 ± 0.37
7.71 ± 0.27
FSH (mlU/mL)
3.05 ± 0.14
3.12 ± 0.36
3.37 ± 0.32
3.49 ± 0.47
Sperm count (x106/mL)
27.0 ± 1.5
23.5 ± 0.8
17.8 ± 1.5**
13.3 ± 1.3**
Sperm motility (%)
71.7 ± 3.3
57.6 ± 5.5
54.0 ± 6.4*
34.6 ± 3.3**
Abnormal sperm (%)
5.9 ± 1.0
5.9 ± 1.0
6.1 ± 0.9
6.5 ± 1.0
6 months
Testosterone (ng/mL)
3.33 ± 0.98
2.03 ± 0.73
1.08 ± 0.17**
0.89 ± 0.29**
LH (mlU/mL)
4.89 ± 0.25
4.84 ± 0.25
5.88 ± 0.25*
5.66 ± 0.17**
FSH (mlU/mL)
2.36 ± 0.35
2.59 ± 0.37
3.16 ± 0.32
4.27 ± 0.52**
Sperm count (x106/mL)
21.5 ± 0.7
19.7 ± 0.9
13.6 ± 1.1**
6.6 ± 0.9**
Sperm motility (%)
60.6 ± 5.1
46.8 ± 6.7
23.1 ± 3.2**
17.4 ± 5.0**
Abnormal sperm (%)
6.5 ± 1.0
9.0 ± 1.0
13.8 ± 1.8**
14.5 ± 1.1**
From Chen et al., 2011 Data are mean±S.E.; n = 10; Significantly different from control: *p<0.05; **p<0.01.
Li et al. (2008) also observed reproductive effects in male Sprague-Dawley rats after i.p. injection of 0, 5,
10, or 15 |ig microcystin-LR/kg-day by for 28 days. The microcystin-LR was dissolved in a minimal
amount of 0.1% methanol and diluted with saline for injection. In all treatment groups, body weight gain
and sperm motility were decreased. The percent of abnormal sperm was increased in all dose groups. The
high-dose group had decreased absolute and relative testes weights and epididymal sperm concentrations.
In both 10 and 15 (ig/kg-day dose groups, serum testosterone levels were significantly decreased. At 5
(ig/kg-day, both FSH and LH were significantly increased, but significantly decreased at 15 (ig/kg-day.
Histopathological change, including atrophied and obstructed seminiferous tubules in the testes occurred
in all treated groups, but were more pronounced in the high-dose group.
Cellular damage was observed in the testes of male mice administered a single i.p. dose of 55-110 (.ig
microcystin-LR/kg prepared from a crude extract of a lyophilized cyanobacterial bloom (Li et al., 201 lb).
The effects of a single i.p. injection of microcystin extracts from a surface bloom containing 167.7 |ig
microcystin-RR/mL and 47.0 |ig microcystin-LR/mL or 80.5 (.ig microcystin-LR equivalents/mL was
found to have an effect on male rabbit testes. Lesions, including a variety of histological changes to both
spermatogonia and Sertoli cells, were seen in immature male Japanese white rabbits (1.6±0.2 kg) treated
with 12.5 |ig microcystin-LR equivalents/kg; recovery occurred by 48 hours with the tissue resembling
the control (Liu et al., 2010).
In a study by Chen et al. (2013), male rats (10 per group) were i.p. injected with microcystin-LR (purity >
98%) in saline for 50 days at doses of 1 or 10 (ig/kg/day; a control group (n =10) was injected with the
same volume of 0.9% saline solution. Animals were sacrificed twelve hours following the final injection
and the testes removed. The relative testes weight was significantly decreased (p <0.01) at 10 (ig/kg/day,
however, body weight and absolute organ weight data were not given. Light microscopic observations
indicated that the space between the seminiferous tubules and lumen size increased with increasing dose;
blockages in the seminiferous tubules were also reported at 10 (ig/kg/day. Ultrastructural observations in
Health Effects Support Document for Microcystins - June, 2015
46
-------
spermatogonia showed some abnormal histopathological characteristics, including cytoplasmic shrinkage,
cell membrane blebbing, swollen mitochondria and deformed nucleus; these changes became more
pronounced with increasing dose. Using qPCR methods, the transcriptional levels of select cytoskeletal
and mitochondrial genes were determined. Microcystin-LR exposure affected the homeostasis of the
expression of cytoskeletal genes, causing possible dysfunction of cytoskeleton assembly. Transcription of
(3-actin, |3-tubulin, and stathmin were significantly decreased while ezrin and moesin were increased. In
both microcystin-LR treated groups, all 8 mitochondrial genes related to oxidative phosphorylation
(OXPHOS) were significantly increased. The levels of reactive oxygen species (ROS) were significantly
increased (p <0.01) at 10 (ig/kg/day as was mitochondrial swelling and DNA damage. Changes in
testicular hormone levels included increased FSH levels at 10 (ig/kg/day, significantly increased LH
levels in both treated groups (p <0.05 or 0.01), and decreased testosterone levels in both dose groups (p
<0.01) compared to those of the controls. The authors concluded that this study provides evidence that
both cytoskeleton structural disruption and mitochondrial dysfunction interact through induction of
reactive oxygen species and oxidative phosphorylation resulting in testis impairment following exposure
to microcystin-LR.
Wu et al. (2014) studied the effect of i.p. injection of microcystin-LR (commercial product; purity not
specified) on the female reproductive tract of both rats and mice. Female Sprague-Dawley rats (n = 6)
were given 0 or 200 (ig/kg/day for six days and the ovaries removed for Western blot analysis of
microcystin-LR-protein phosphatase 1 and 2A (PP1/2A) adducts. Female BALB/c mice (n = 20) were
given 0, 5, or 20 (ig/kg/day for 28 days. A subset of six mice per group was maintained for 28 days for
estrous cyclicity monitoring. The remaining mice were sacrificed 24 hours after the last injection for
serum hormone analysis and histopathology of the ovaries. In rats, microcystin-LR-PPl/2A adducts were
detected in liver and ovary with the band from the ovarian extract being much weaker than that of liver.
At 20 (ig/kg/day, mice had significantly lower ovarian weight and a significantly decreased number of
primordial follicles compared with those of controls. Estrous cyclicity was not affected by treatment with
microcystin-LR and no differences in serum FSH, LH, and estradiol were seen. Serum progesterone
levels were significantly reduced in both groups of treated mice compared with that of controls.
Bu et al. (2006) evaluated the potential embryotoxicity of microcystis cell extracts from water samples
from the Nanwan reservoir in China in pregnant Kunming mice. HPLC analysis showed that the main
components of these samples were microcystin-LR and -YR, with the majority being microcystin-LR.
The study authors indicated that the LD50 for was much lower compared to previous i.p. studies and
noted that it was possible that other substances that can increase the toxicity of microcystin may have
been present in the extracts. Bu et al. (2006) exposed pregnant mice to 3, 6, or 12 (ig/kg microcystin
(12/dose group) on GD 6-15 via i.p. injection. Control mice were injected with saline on the same GDs.
Mice were sacrificed on GD 18 and number of dead and resorbed fetuses and viable fetuses was recorded.
Additionally, the study authors evaluated the body weight, body length, tail length, skeletal development,
and external anomalies of viable fetuses. The number of viable embryos was statistically significantly
decreased and the number of dead or resorbed embryos was statistically significantly increased at the high
dose. In the fetuses, body weight, body length, and tail length were also statistically significantly
decreased at the high dose. The study authors noted that petechial hemorrhage and hydropic degeneration
were observed in the livers of fetuses at the 6 and 12 |ig/kg doses.
6.2.5.2 Developmental Effects
Fawell et al. (1999) reported the results of a developmental toxicity study of microcystin-LR (commercial
product; purity not stated) given via gavage to Crl:CD-l(ICR)BR (VAF plus) mice. Microcystin-LR (0,
200, 600 or 2,000 (ig/kg/day) was administered to groups of 26 mice on days 6-15 of pregnancy. On day
18, the mice were sacrificed and necropsied. External, visceral and skeletal examinations were performed,
and weight and sex of the fetuses were recorded. Of the 26 dams receiving 2,000 (ig/kg/day, seven died
Health Effects Support Document for Microcystins - June, 2015
47
-------
and 2 others were sacrificed moribund. An altered liver appearance was observed during gross
examination. The surviving dams in this group did not express any clinical signs of toxicity or differences
in body weight or food consumption. According to the authors, fetal body weight was significantly lower
than controls and delayed skeletal ossification was observed at the highest dose. However, data were not
included in the publication. No effects on litter size or resorptions were observed in any treatment group,
nor were there increases in external, visceral or skeletal abnormalities in fetuses. The 600 (ig/kg/day dose
is the apparent NOAEL with a Frank effect level (FEL) of 2,000 (ig/kg/day for decreased skeletal
ossification and lethality low fetal body weight.
Groups of 6-8 timed-pregnant CD-I mice were administered microcystin-LR (commercial product; 95%
purity) in sterile saline by i.p. injection at doses of 0, 32, 64, or 128 (.ig/kg. Animals were treated on
gestation days 7-8, 9-10, or 11-12 followed by sacrifice on day 17. Fetuses were examined for gross and
skeletal malformations (Chernoff et al., 2002). Maternal weight change, pregnancy rate, litter size, fetal
deaths, and fetal body weight were similar between control and treated groups. No treatment-related
malformations were found in fetal examination.
In another part of the Chernoff et al. (2002) study, pregnant CD-I mice were administered microcystin-
LR (commercial product; 95% purity) in sterile saline by i.p. injection at doses of 0, 32, 64, 96, or 128
(ig/kg. Animals were treated on gestation days 7-8, 9-10, or 11-12, and allowed to give birth. The growth
and viability of pups was monitored for 5 days. A different lot of microcystin-LR from the same supplier
was used in this part of the study and was much more toxic than the lot used in the developmental toxicity
study. Maternal deaths were observed at all doses independent of days of dosing. In the control and
treated groups, 0/25, 3/27, 19/35, 33/34, and 33/34 animals died, respectively. For surviving dams,
numbers of pups born, and offspring survival and body weight through postnatal day 5 were not affected
by treatment.
An in vitro study to determine the effect of microcystin-LR in the syncytiotrophoblast using villous
cytotrophoblast isolated from term human placentas was done by Douglas et al. (2014). Cells were
exposed to 0, 0.5, 1.25, 2.5, 5, 10, 20, and 25 (.iM of microcystin-LR and analysis of trophoblast
morphology, detachment, differentiation and apoptosis was performed. Measurement of secretion of
human chorionic gonadotropin (hCG), the pregnancy hormone secreted by the syncytiotrophoblast was
also done. The authors observed round cells and significant cell loss at 25 (.iM microcystin-LR, but no
change in the spreading and general morphology of trophoblasts at concentrations lower than 25 (.iM. No
detachment, apoptosis, or differentiation of cytotrophoblasts to multinucleated syncytiotrophoblast were
observed. However, the secretion of the pregnancy hormone hCG was increased in a dose-dependent
manner. However, the cause of the increase in the hCG secretion remains undetermined.
Li et al. (2015) conducted a developmental neurotoxicity study in female Sprague Dawley Rats. The
neurotoxicity portion of the study is presented in Section 6.2.4 above. Groups of 7 and 28 day old rats
were dosed intragastrically every other day for 8 weeks at doses of 0, 1, 5 or 20 |ag microcystin-
LR/kg/days in a solution that contained 0.002% methanol (v/v). The microcystin-LR was identified as
95% pure. These doses normalize to values of 0, 0.5, 2.5 or 10 (ig/kg/day. At the end of the exposure
period the females were mated with unexposed males. No dosing occurred during the gestation period.
After conception, gestation and delivery, the litters were culled to 4 males and 4 females, where possible,
with subsequent evaluation of the pups for developmental neurotoxicity. At the end of the gestation
period the only significant change observed for the dams was decreased body weight gain (p<0.05). The
number of pregnant dams decreased across dose groups (7, 6, 5, and 5, respectively) while the number of
dead pups increased (0.4 ±0.2, 0.7 ±03, 1.3 ±0.8, and 1.6 ±0.9, respectively) but the differences were not
statistically significant. Other parameters evaluated were live pups/litter, fetal weight and sex-ratio; no
differences across dose groups were noted. Developmental milestones (e.g. incisor eruption, hair
appearance and eye opening) did not differ significantly from those for the controls. The NOAEL based
Health Effects Support Document for Microcystins - June, 2015
48
-------
on maternal gestational weight gain is 2.5 (ig/kg/day and the LOAEL is 10 (ig/kg/day. There was no
direct exposure to the dams during the gestation and lactation periods.
6.2.6 Chronic Toxicity
Falconer et al. (1988) conducted a chronic exposure experiment (up to 1 year) using an extract of a M.
aeruginosa water bloom in Swiss Albino mice. A concentration-dependent increase in mortality, reduced
body weight and a concentration-dependent increase in serum alanine aminotransferase levels were
observed among groups of mice receiving serial dilutions of the extract as their drinking water for a year.
The incidence of bronchopneumonia observed in the treated animals was directly related to the
microcystin concentration. No significant differences in liver histopathology were observed when
compared to the control, although the observed liver changes (neutrophil infiltration, hepatocyte necrosis)
were slightly more prevalent in treated animals. The data showed some indication of sex differences in
susceptibility; male mice showed effects (including mortality and serum enzyme level increases) at lower
concentrations than females.
Thiel (1994) reported the results of a chronic toxicity study of microcystin-LA in velvet monkeys as an
expanded abstract in the proceedings of an international workshop; a published version of this study was
not located. A group of six monkeys was divided into two treatments: three controls and three monkeys
that were given increasing intragastric doses of microcystin-LA for 47 weeks. At the beginning of the
study, the dose was 20 (.ig/kg/dav and increased to 80 (ig/kg/day at study termination. The intervals of the
dosage were not reported. No body weight or clinical signs such as respiration, pulse, or temperature were
observed. No statistically significant changes in hematological parameters (hematocrit, bilirubin,
hemoglobin, erythrocyte and leukocyte, and platelet count) were observed. No changes were observed in
serum biochemistry analyses including albumin, globulins and electrolytes, as well as serum AST, LDH,
ALP, ALT and GGT. Histopathological examination of the liver and other organs, not specified in the
expanded abstract, did not show any differences in treated monkeys when compared with controls.
A chronic study done by Ueno et al. (1999) evaluated the toxicity of microcystin-LR in mice via drinking
water. Two hundred 6-week-old female BALB/c mice were randomly assigned to receive either drinking
water {ad libitum) containing 20 |ig/L of microcystin-LR (95% pure) or no treatment for 7 days/week for
up to 18 months. After 3, 6 and 12 months, 20 animals from each group were sacrificed, and the
remaining 40 animals in each group were retained for chronic toxicity evaluation and sacrificed at 18
months.
The authors recorded daily observations for clinical signs of toxicity, morbidity and mortality, and weekly
estimates of food and water consumption. The body weights were recorded weekly for the first 2 months,
biweekly up until the first year and monthly until 18 months. Blood was obtained from 20 animals from
each group at 3, 6, 12 and 18 months. Hematological evaluations were done in 10 animals per group, and
samples from 10 additional animals were used for serum biochemistry evaluation. Complete necropsy of
10 animals per group was conducted and necropsy was also done in those animals in the chronic study
when moribund or found dead prior to scheduled sacrifice or upon sacrifice at 18 months. Record of
relative and absolute organ weights including liver, kidneys, spleen, thymus, adrenal, ovaries, brain, heart
and uterus were done for 9-10 animals per group at each scheduled sacrifice, and histopathological
evaluation of these and numerous other organs was conducted. Immunohistochemistry of the liver was
also examined upon sacrifice of three to five animals per group to determine the distribution of
microcystin-LR in the liver.
The calculated cumulative intake of microcystin-LR over 18 months was 35.5 (ig/mouse (based on
weekly estimates of water consumption) (Ueno et al., 1999). This is equivalent to an exposure of 2.3
(ig/kg/day based on the reported average adult body weight of 26.68 g/mouse and the reported 567 day
exposure. No clinical signs of toxicity and no statistically significant differences in body weight, food
Health Effects Support Document for Microcystins - June, 2015
49
-------
consumption, water consumption or hematology were observed. However, hematology data were lost due
to sampling errors from the 3-month sacrifice. Survival in the control and chronic treatment groups was
similar. After 12 months, the treated mice had a statistically significant decrease in serum ALP (13%) and
at month 18, a significant increase in cholesterol (22%). None of these effects were considered by the
authors to be toxicologically significant in the absence of other treatment-related effects. However,
according to the authors, the increase in cholesterol could be related to interference of microcystin-LR
with bile acid transport from the liver.
Treated mice showed sporadic changes in absolute and relative thymus weight, but histological and
morphometric evaluation revealed no abnormalities attributable to exposure (Ueno et al., 1999). Treated
mice sacrificed after 12 months showed a decrease in heart weight that was not considered treatment-
related in the absence of histopathological changes. In contrast to other studies, no difference in the
incidence of liver histopathology between treated and control mice was observed. No accumulation of
microcystin-LR was observed after immunohistochemistry of the liver.
Microcystin-LR (commercial product; >95% purity) was administered for 180 days to 8-week old male
C57bl/6 mice (10/treatment group) via drinking water at the following concentrations: 0, 1, 40, or 80 (ig/L
(Zhang et al., 2010). The doses were reported as 0, 0.2, 8.0, and 16 (ig/kg/day, but the method of
calculation was not given by the authors. Body weight was measured at the beginning and at the end of
the study. Livers were removed at sacrifice and processed for routine (hematoxylin-eosin) or
immunohistochemical staining to measure matrix metalloproteinase (MMP3) expression. Measurement of
MMP protein and mRNA levels were measured in other liver portions.
A significant (p <0.01) decrease in body weight, and an increase in relative liver weight, was reported at
8.0 and 16.0 (ig/kg/day (Zhang et al., 2010). Histopathology of mice treated with 8.0 and 16.0 (ig/kg/day,
revealed infiltrating lymphocytes and fatty degeneration in the liver, but incidence and severity data were
not provided. There was a significant increase in the area stained positive for MMP2 at 8.0 and 16.0
(ig/kg/day and for MMP9 in all treatment groups. Only in the high-dose group, the MMP2 protein
concentration was significantly increased. The concentrations of MMP9 protein were increased at all
doses. In the mid- and high-dose groups, messenger RNA expression for both MMPs was significantly
increased. The phosphorylation extracellular signal-regulated protein kinase (ERK) 1/2 and p38 (members
of the mammalian of the mitogen-activated protein kinase (MAPK) family were also increased.
In a subsequent study by Zhang et al. (2012), microcystin-LR (commercial product; >95% purity) at
concentrations of 0, 1, 40, or 80 |ig/L (0, 0.2, 8.0, and 16 (ig/kg/day) was administered to 8-week old male
C57bl/6 mice (10/treatment group) via drinking water for 270 days. . Body weight was measured at the
beginning and at the end of the study, and livers were removed and processed for routine or
immunohistochemical staining at sacrifice to measure MMP expression. MMP protein and mRNA
analyses were done in other liver portions from five randomly selected mice. Body weight results were
not included in the main publication but the data was reported in supplemental information. No
differences in water consumption were observed between the groups. Histopathology showed infiltrating
lymphocytes and fatty degeneration in the livers of mice (doses not specified in main publication). In all
dose groups, MMP expression and protein levels for both MMP2 and MMP9 were significantly
increased. MMP mRNA levels were also increased in all dose groups for MMP2 and in the mid- and
high-dose groups for MMP-9.
3 Matrix metalloproteinases are a family of zinc requiring matrix-degrading enzymes, which include the
collagenases, gelatinases, and the stromelysins, all of which have been implicated in invasive cell behavior (Brooks
etal. 1996).
Health Effects Support Document for Microcystins - June, 2015
50
-------
6.2.7 Immunotoxicity
Shirai et al. (1986) reported that C3H/HeJ mice, immunized i.p. with either sonicated or live cells from a
Microcystis water bloom, developed delayed-type hypersensitivity when challenged 2 weeks later with a
subcutaneous injection of sonicated Microcystis cells. A positive reaction, as assessed by footpad
swelling, was seen in mice immunized with either live cells or sonicated cells. Because this strain of
mouse is unresponsive to LPS, the footpad delayed-type hypersensitivity was not related to LPS, thus, the
antigenic component of the sonicated cells is not known, but might have been microcystin.
Shen et al. (2003) studied the effect of cyanobacterial cell extract on immune function. Mice received 14
daily i.p. injections containing a cell-free extract from a water bloom dominated by M. aeruginosa at 16,
32 and 64 mg lyophilized cells/kg body weight doses or as 4.97, 9.94 and 19.88 |ig/kg of microcystin
equivalents. Analysis by HPLC indicated that the microcystin content of the extract was 79.53%,
although specific congeners in the extract were not reported. Immunotoxicity endpoints examined were:
phagocytosis, lymphocyte proliferation and antibody production in response to sheep red blood cells.
Phagocytic capacity was reduced at the two highest doses, but percentage phagocytosis was not affected.
B-lymphocyte proliferation was significantly reduced (33%), compared to controls (at 32 mg/kg). Body
weight was significantly reduced in all treatment groups. Relative spleen weight was significantly
increased at 9.94 |ig/kg. and significantly decreased at 19.88 (.ig/kg. In the high-dose group, relative
thymus weight was significantly decreased, and relative liver weight was significantly increased in all
treatment groups, although not related to dose. However, changes in T-lymphocyte proliferation were
mild, and deemed biologically insignificant. In the treated mice, humoral immune response, as measured
by antibody-forming plaques, was reduced in a dose-dependent manner.
Shi et al. (2004) reported a study where mice received a single i.p. injection containing a cell-free extract
from a water bloom dominated by M. aeruginosa processed in the same manner as the Shen et al. (2003)
study. Although specific congeners in the extract were not reported, it was stated that microcystin-LR was
the predominant component. Doses were reported as 0, 23, 38, 77 and 115 mg lyophilized cells/kg body
weight or as 0, 7, 12, 24 and 36 |ig/kg of microcystin equivalents. Animals were sacrificed 8 hours after
exposure. Messenger RNA levels of TNF-a, IL-1(3, IL-2, and IL-4 were significantly decreased, IL-6 was
unaffected, and IL-10 was increased at the lowest dose and decreased at higher doses. None of the
changes were dose-related.
Chen et al. (2004b, 2005b) evaluated the role of nitric oxide generation and macrophage related cytokines
on the reduced phagocytic capacity induced by pure microcystin-LR. A dose-dependent inhibition of
nitric oxide production was observed in activated macrophages, and a repressive effect was seen in
cytokine formation at the mRNA level (e.g., IL-1J3, TNF-a, GM-CSF, IFN-y) after either a 24-hour (Chen
et al., 2004b) or a 6-hour treatment (Chen et al., 2005b). Hernandez et al. (2000) showed that microcystin-
LR enhanced the early spontaneous adherence of polymorphonuclear leukocytes (PMNs) to substrate; no
effects were found on late adherence (steady state) or with stimulated PMNs.
Several studies evaluated the effects of microcystin-LR on immune system components in vitro (Lankoff
et al., 2004b; Teneva et al., 2005; Chen et al., 2004b; Kujbida et al., 2006). Lankoff et al. (2004b)
reported that microcystin-LR inhibited B-cell proliferation in human and chicken peripheral blood
lymphocytes at all concentrations tested and decreased T-cell proliferation only at the highest
concentration. Apoptosis was enhanced in both human and chicken lymphocytes (Lankoff et al., 2004b).
Similarly, microcystin-LR was cytotoxic to mouse splenocytes, and caused apoptosis in B-cells but not in
T-cells (Teneva et al., 2005).
Kujbida et al. (2006) assessed the effects of microcystin-LR and [Asp3]-microcystin-LR on human
polymorphonuclear lymphocytes (PMNs) in vitro. Both compounds caused migration of neutrophils in a
chemotaxis chamber, suggesting that PMNs may migrate from the blood stream to the organs such as the
Health Effects Support Document for Microcystins - June, 2015
51
-------
liver that concentrate microcystins. In addition, both caused a dose-related increase in reactive oxygen
species (ROS) production as measured by chemiluminescence of PMN degranulation products that
accompany ROS production. The phagocytosis of Candida albicans by PMNs was increased after
exposure to either compound, but only microcystin-LR increased the intracellular killing of C. albicans.
These findings suggest the possibility that PMNs may mediate some of the toxic effects of microcystins.
Kujbida et al. (2008) found that microcystin-LR, microcystin-LA, and microcystin-YR increased
interleukin-8 levels and extracellular ROS in human neutrophils, and chemoattractant-2a|3 in rat
neutrophils, but had no effect on tumor necrosis factor-a in either rat or human neutrophils. In vitro all
three microcystins caused neutrophil chemotaxis by increased intracellular calcium levels (Kujbida et al.,
2009). In vivo, topical application of microcystin-LR to male rats caused an enhancement of the number
of rolling and adhered leukocytes in the endothelium of postcapillary mesenteric venules, but
microcystin-LA and microcystin-YR had no effect (Kujbida et al., 2009).
Yuan et al. (2012) evaluated the immunotoxicity in rabbits using extracts of microcystins isolated from a
surface bloom in China. The extracts contained 0.84 mg/g dry weight of microcystin-RR, 0.50, mg/g dry
weight of microcystin-LR, and 0.07 mg/g dry weight of microcystin-YR. Four rabbits per treatment group
received single i.p. injections of 0, 12.5, or 50 |ig microcystin-LR equivalents/kg. After administration of
the 50 |ig/kg dose, blood was collected from the heart at 0, 1, and 3 hours, and at 0, 1, 3, 12, 24, 48, and
168 hours after administration for the 12.5 |ig/kg dose. A significant increase in plasma white blood cells
was observed after microcystin-LR treatment with both doses. The peak increase was observed 1 hour
after treatment with 50 (ig/kg and 12 hours after treatment with 12.5 |ig/kg. IFN-y, INF-a, IL-3, IL-4, and
IL-6 production was decreased at all time points measured after the 50 |ig/kg treatment. However, at the
12.5 |ig/kg dose, production of IFN-y, INF-a, IL-3, IL-4, and IL-6 was increased through the first 12
hours after exposure, but decreased or was the same as the controls from 24 to 168 hours.
Bernstein et al. (2011) studied skin sensitization to non-toxic extracts ofM aeruginosa in 259 patients
with chronic rhinitis over 2 years. Patients were evaluated with aeroallergen skin testing and skin-prick
testing (SPT). The authors found that 86% of the clinical subjects had positive skin prick tests to
Microcystis aeruginosa, and that patients with existing allergic rhinitis were more likely to have reactions
and sensitization to cyanobacteria than the controls (non-atopic health subjects). This study indicates that
cyanobacterial allergenicity is associated with the non-toxic portion of the cyanobacteria.
Geh et al., (2015) studied the immunogenicity ofM aeruginosa toxic and non-toxic extracts in patient
sera (18 patients with chronic rhinitis and 3 non-atopic healthy subjects collected from the study done by
Bernstein et al., in 2011). ELISA test was used to test IgE-specific reactivity, and 2D gel electrophoresis,
followed by immunoblot and mass spectrometry (MS), was done to identify the relevant sensitizing
peptides. The authors found an increase in specific IgE in those patients tested with the non-toxic
microcystin extract than the toxic extract. After pre-incubation of the non-toxic extract with various
concentrations of microcystin, the authors found that phycocyanin and the core-membrane linker peptide
were responsible for the release of (3- hexosaminidase in rat basophil leukemia cells. The authors
concluded that non-toxic strains of cyanobacteria are more allergenic than toxic-producing strains in
allergic patients, and that the toxin may have an inhibitory effect on the allergenicity.
6.2.8 Hematological Effects
Several studies have noted thrombocytopenia (platelet deficiency) in laboratory animals treated with
microcystins or bloom extracts purportedly containing microcystins (Slatkin et al., 1983; Takahashi et al.,
1995). Early investigations with parenteral injection into mice of microcystins found thrombocytopenia,
pulmonary thrombi, and hepatic congestion (Slatkin et al., 1983). However, in vitro studies have shown
that microcystin-LR neither induces nor impedes the aggregation of platelets (Adams et al., 1985).
Pulmonary thrombi apparently consist of necrotic hepatocytes circulating in the blood. Subsequent
Health Effects Support Document for Microcystins - June, 2015
52
-------
research supports the hypothesis that hematological effects observed in animals acutely exposed to
microcystins are secondary effects of liver hemorrhage (Takahashi et al., 1995).
Takahashi et al. (1995) reported dose-dependent reductions in erythrocyte count, leukocyte count,
hemoglobin concentration, hematocrit and coagulation parameters one hour after rats were exposed to
microcystin-LR (100 and 200 |ig/kg i.p). None of these parameters changed until after massive liver
hemorrhage commenced. Further, hematological changes such as increased prothrombin time and fibrin
deposition in the renal glomeruli were not observed. The authors concluded that the depletion of blood
components occurred as a result of liver hemorrhage.
Sicinska et al. (2006) evaluated the effects of microcystin-LR on human erythrocytes in vitro.
Microcystin-LR exposure resulted in the formation of echinocytes, hemolysis, conversion of
oxyhemoglobin to methemoglobin, and a decrease in membrane fluidity. In addition, measures of
oxidative stress were affected in treated erythrocytes; glutathione reductase and superoxide dismutase
activities were decreased, while ROS and lipid peroxidation were increased
6.3 Carcinogenicity
6.3.1 Cancer Epidemiology Studies
A survey of microcystin content in drinking water supplies was conducted in Haimen City, China to
determine if microcystins in drinking water supplies could contribute to the higher incidence of liver
cancer (Ueno et al., 1996). Samples were taken in ponds/ditches and river waters as well as shallow and
deep wells and analyzed by ELISA. Microcystin concentrations were higher in pond/ditch water (17%
reported as positive with concentration >50 pg/mL), followed by river water (32% positive), shallow
wells (4% positive), and deep wells (no detections >50 pg/mL). The averages of microcystin
concentrations across the drinking water types differed with an average of 101 in pond/ditch, 160 in river,
and 68 pg/mL in shallow well samples. The authors used the average microcystin concentration in the
ponds/ditches (101 pg/mL) and in river water (160 pg/mL), and the average adult consumption over June
to September (1.5 L) to calculate the exposure levels to microcystin in Haimen city. The authors
determined that over a period of 4 months the levels to which people would have been exposed was 0.19
pg of microcystin per day and the average adult was exposed to these levels over a period of 40 to 50
years. The authors did not collect any data that would support a correlation between consumption of the
different water sources showing seasonal contamination with microcystins, and local cases of
hepatocellular carcinoma. Therefore, this study generates a hypothesis of a possible association to
exposure to microcystins, but does not investigate that relationship. Haimen City, like Haining City
discussed below is on the Yangtze River. It was once largely an agricultural area but is now also noted for
its production of textiles and more recently electronics (http://www.ccpittex.com/eng/tbases/49302.html).
Thus there are likely multiple exposures to possible carcinogens that could account for the high cancer
incidence.
Zhou et al. (2002) conducted a retrospective cohort study to analyze a previously reported association by
Jiao et al. (1985) and Chen et al. (1994) between colorectal cancer and exposure to microcystins in
drinking water in a Chinese province. Between 1977 and 1996, a total of 408 cases of primary colorectal
adenocarcinoma (245 rectums and 163 colons) obtained from the Cancer Registry of Haining Cancer
Research Institute, were diagnosed in eight randomly selected towns within Haining City of Zhejiang
Province. The local cancer registry was used to identify the cases and verified independently by two
pathologists. The drinking water source used during the lifetime was used as a surrogate of oral exposure
to microcystins. Interviews of patients or family members of deceased cases were performed to obtain
information on drinking water source. Ten water sources including 3 rivers, 3 ponds, 2 wells and 2 taps,
were randomly selected and sampled twice per month for microcystins from June through September
Health Effects Support Document for Microcystins - June, 2015
53
-------
(total of eight samples from each source) and analyzed by ELISA. The authors did not provide
information on the congeners tested or a complete description of the "tap" water samples. However, the
study description implies that samples were collected from various treatment plants.
To determine the incidence rate of colorectal cancer, the authors compared the rates among the four
different water sources with well water users serving as the reference population. The authors determined
an average incidence rate of colorectal cancer of 8.37/100,000 per year across all of the study areas. The
colorectal cancer incidence rates among users of the tap, pond, and river water sources were significantly
increased compared with the incidence among well water users. Relative risks (RR) are listed in Table 6-8
and differed by water source; 1.88 for tap water, while river and pond water use both had a RR greater
than 7.0. Very little difference in colorectal cancer incidence between river and pond water users was
observed. The authors suggested that exposure to trihalomethanes in tap water could contribute to the risk
for those users.
Microcystins were detected, only in river and pond water, at concentrations exceeding 50 pg/mL, which
was considered by the authors to be the limit for positive detection (Zhou et al., 2002). Average
concentrations in river and pond water were 30-50 fold higher than those for well or tap water. Since
about 25% of the residents in each of the eight towns used river and pond water for drinking water, a
comparison between the average microcystin concentration in river and pond water in each town with the
incidence rate by town was performed. Their results showed a strong correlation between colorectal
cancer incidence rate and concentration of microcystin (Spearman correlation coefficient = 0.88, p<0.01).
This comparison is limited by failure to test for chemical carcinogens that could have also been present
in the untreated surface water sources. For example, Haining City borders the Yangtze River and is the
site of industries specializing in leather products and textiles and electronics among others
http://en.haining.gov.cn/
The study by Zhou et al. (2002) provides suggestive evidence for an association between colorectal
cancer and exposure to microcystin in drinking water, which is consistent with earlier reports of an
association between drinking water from the river or pond and incidence of colorectal cancer in the
Zhejiang Province of China (Jiao et al., 1985; Chen et al., 1994). Since demographic information was not
provided in the study, it is not clear which factors, including diet, genetics, and lifestyle, and chemical
contaminants associated with colorectal cancer, were adequately controlled in the analysis.
A number of epidemiological studies have been conducted in an area of Southeast China with high rates
of hepatocellular carcinoma. These studies are summarized by the International Agency for Research on
Cancer (IARC, 2010) and Health Canada (2002). Overall a positive association was found between the
risk for hepatocellular carcinoma and surface waters as the drinking water source. In an analysis of pooled
data from six case-control studies, RR was 1.59 (confidence limits not given); estimates of RR from other
individual studies ranged from 1.5-4 (IARC, 2010). Consumption of pond or ditch water was associated
with a higher risk of liver cancer incidence when compared with well water consumption. Confounding
factors such as hepatitis B infection and aflatoxin exposure, were not generally considered in most
studies. The presence of cyanobacteria in the water source was not a component of the study. Thus, the
only relationship between these estimates of risk and cyanotoxins is the fact that cyanobacteria are
primarily surface water contaminants.
Health Effects Support Document for Microcystins - June, 2015
54
-------
Table 6-8. Relative Risk of Colorectal Cancer By Drinking Water Source
Water Source
Well
Tap
River
Pond
Colorectal Cancer Incidence Rate
per 100,000
3.61
6.77
28.50
27.76
Relative Risk of Colorectal Cancer (p<0.01)
-
1.88
7.94
7.7
95% CI
-
1.39-2.54
6.11-10.31
5.75-10.30
Number of Microcystin Samples
>50 pg/mL
0/12
0/17
25/69
6/35
Mean Microcystin Concentration (pg/mL)
3.73
4.85
141.08
106.19
Maximum Microcystin Concentration (pg/mL)
9.13
11.34
1083.43
1937.94
From Zhou et al., 2002
A case-control study was done to evaluate the relationship between liver cancer in Haimen City, China
and microcystin in drinking water (Yu et al., 2002). Participants were selected from a pool of 248 patients
with hepatocellular carcinoma and 248 age-, sex- and residence-matched controls. Of those, 134 paired
cases and controls agreed to blood samples for virus infection and ALDH2 (Aldehyde dehydrogenase 2)
and CYP2E1 gene polymorphism analyses. The authors evaluated a variety of risk factors for liver cancer
including hepatitis B and C virus infection, aflatoxin B1 or microcystin exposure along with genetic
polymorphisms, smoking, drinking, and diet. Questionnaire information on possible lifestyle and dietary
risk factors for liver cancer was also conducted. Exposure to microcystin was assessed based on type of
drinking water supplied (tap, deep or shallow well, river, ditch, or pond water). No association between
consumption of river, pond, or ditch water and hepatocellular carcinoma was determined by either
univariate or multivariate analysis. The authors identified hepatitis B virus infection and history of i.v.
injection as factors strongly associated with primary liver cancer (Yu et al., 2002).
An ecological epidemiological study was conducted to investigate the relationship between drinking
water source and incidence of primary liver cancer in Florida (Fleming et al., 2002). Data on
cyanobacteria and toxins, especially microcystins in surface drinking water sources in Florida were used
to measure the exposure. All cases of primary hepatocellular carcinoma reported to the Florida State
cancer registry between 1981 and 1988 were the study population. The study population was placed in
two groups depending on the residence location at the time of diagnosis: those served by 18 surface
drinking water supplies, and a second group using other sources. Deep groundwater treatment plants and
surface water treatment plants and their service areas were geocoded (Figure 6-1). The following
comparisons were made:
• Comparison between cases residing in the service area of a surface water treatment plant with
those residing in the service area of a deep groundwater treatment plant. Several referent groups
were identified (one randomly sampled from the available groundwater service areas, one
matched on median income and rent, one matched on ethnic makeup and one matched on income,
rent and ethnicity).
• Comparison between cases in the surface water service area with equally-sized buffer areas
surrounding the surface water service area, but not served by the treatment plant (GIS was used to
delineate a buffer area assuming a population living contiguous to, but outside of, surface water
treatment service areas).
• Comparison between cases and the primary liver cancer incidence in the general population of
Florida.
There was no statistically significant difference among the individual incidence rates between the controls
(four sets of 18 groundwater service areas) and the individual rates from the 18 individual surface water
service. A statistically significant age-adjusted cancer rate for hepatocellular carcinoma (1.13) was
Health Effects Support Document for Microcystins - June, 2015
55
-------
associated with residence in a surface water service area for 1981-1998, when the 18 ground water system
areas were pooled. This rate was lower than the age-adjusted rates for the four comparison ground water
areas and the state of Florida in general. The Standard Rate Ratio (SRR) for the surface water areas
compared to the four ground-water area controls was 0.95. 0.84, 0.81, and 0.98. Compared with the state,
the SRR was 0.8 (1.13-1.41). It should be noted that the measure of exposure was residence within a
circular surface water service area derived using a diameter based on the average size of the service area
plus two standard deviations, with the treatment plant theoretically but not physically located at the center
of the circle. The dimensions of the ground water areas were determined in a similar fashion.
V.
Treatment
Control 1
Control 2
Control 3
Control 4
N
-Pi
* *
w
100 0 100 200 Miles
Figure 6-1. Sites of Surface Water Treatment Service Areas and Control Ground Water Treatment
Service Areas.
From: Fleming et al., 2002
A statistically significant increase in the incidence of hepatocellular carcinoma was observed for those
residing within the surface water service area (SRR=1.39, CI=1,38-1.4; average age adjusted cancer rate
1.15 versus 0.83) when compared with residence in the actual (i.e., not estimated as above) surface water
service areas and residence in the buffer areas surrounding the service areas. According to the 1990
census data, ethnic and socioeconomic backgrounds of the service areas and buffer areas were similar
(data not reported by the authors). When compared to the incidence of hepatocellular carcinoma in the
general Florida population, the incidence of hepatocellular carcinoma in the buffer areas was also
significantly lower (SRR=0.59 average age-adjusted state cancer rate = 1.41).
Due to the ecological design of the study by Fleming et al. (2002) establishing an exposure-response
relationship is not possible because of the lack of exposure data on individuals and the strong possibility
of misclassifying the exposure. Given residential mobility and likely latency time for cancer development,
residence in a surface water service area at the time of diagnosis of hepatocellular carcinoma a poor
measure of potential exposure to cyanobacterial toxins. In addition, not using the actual service areas but
instead GIS-generated estimates of surface water service areas with which may be to make the initial
comparisons with groundwater service areas could increase the misclassification of exposure.
Another ecological study by Fleming et al. (2004) evaluated the relationship between incidence of
colorectal cancer and exposure to cvanobactcria using the proximity to a surface drinking water treatment
plant as a surrogate for exposure. The authors used the same methods as those described above for
Health Effects Support Document for Microcystins - June, 2015
56
-------
Fleming et al. (2002). However, the colorectal cancer data was obtained from the Florida Cancer Data
System from 1981 to 1999. The following referent groups were formed:
• A random group of groundwater treatment service areas.
• A group of groundwater treatment service areas matched on median income and rent.
• A group of groundwater treatment service areas matched on ethnic makeup,
• a group of groundwater treatment service areas matched on both median income and ethnicity.
• Groups residing in an equally-sized buffer areas surrounding the surface water service area.
• General Florida population.
No association between colorectal cancer and residence at time of diagnosis in a surface water treatment
area was observed based on results of the Mann Whitney rank sum test, however details of the tests were
not provided.
6.3.2 Animal Studies
6.3.2.1 Oral Exposure
Falconer and Buckley (1989) and Falconer (1991) reported evidence of skin tumor promotion by extracts
of Microcystis spp. The extract was administered at a concentration of 40 (.ig microcystin/mL via drinking
water to mice pretreated topically with an initiating dose of dimethylbenzanthracene (DMBA). Details of
the incidence of tumors in the control mice were not provided by the authors. The total skin tumor weight
in mice drinking Microcystis extract was significantly higher than that of initiated mice receiving only
water after initiation after 52 days. In mice receiving the extract, only the number of tumors per mouse
was slightly increased due to the weight of individual tumors (Falconer and Buckley, 1989). The total
weight of tumors in the mice receiving extract also exceeded that of mice pretreated with DMBA and
subsequently treated with topical croton oil, with or without concurrent consumption of Microcystis
extract.
No evidence of promotion of lymphoid or duodenal adenomas and adenocarcinomas was observed when
Microcystis extract was provided in the drinking water (0, 10, or 40 (ig/mL) of mice pretreated with two
oral doses ofN-methyl-N-nitroso-urea. No primary liver tumors were observed as well. (Falconer and
Humpage, 1996).
No full oral cancer bioassay was found in which animals were administered microcystins or an extract. Ito
et al. (1997b) evaluated the carcinogenicity and liver toxicity of 80 or 100 gavage doses of 80 |ig
microcystin-LR/kg/day (purity not specified) administered to twenty-two ICR mice (13 weeks old; sex
not stated) over the course of 28 weeks (196 days). Microcystin-LR was isolated and dissolved in ethanol
and saline for dosing from a water bloom from Lake Suwa, Japan. After 80 treatments, ten mice were
sacrificed, five were sacrificed after 100 treatments, and seven were withdrawn from treatment and
sacrificed after 2 months of receiving 100 doses. There were three control mice. Although the authors did
not specify the nature of the postmortem examinations, apparently the liver was the only organ examined.
When compared to controls, no change in mean liver weight was observed in the microcystin-LR-treated
animals. The authors reported light injuries to hepatocytes in the vicinity of the central vein in 8 of 15
mice sacrificed immediately after treatment, and in 5 of 7 mice that were withdrawn 2 months after
exposure from treatment. None of the treated animals showed fibrous changes or neoplastic nodules.
Analysis by immunohistochemistry for microcystin-LR and its metabolites failed to detect either the
parent compound or any metabolites in the livers of mice sacrificed immediately after treatment.
Humpage et al. (2000) administered M. aeruginosa extract in drinking water to mice pretreated with
azoxymethane (an extract only control group was not included). The content of microcystins in the
Health Effects Support Document for Microcystins - June, 2015
57
-------
drinking water was determined by mouse bioassay, HPLC, capillary electrophoresis, and protein
phosphatase inhibition. The estimated doses of total microcystins were 0, 382, and 693 (ig/kg/day at the
midpoint of the trial. Mice were sacrificed at intervals up to 31 weeks after commencement of extract
exposure. Enzyme analysis in mice treated with extract showed a concentration-dependent increase in
ALP and decrease in albumin. A concentration-dependent increase in the mean area of aberrant crypt foci
of the colon was observed. However, the number of foci per colon and the number of crypts per focus
were not different among the groups. Two colon tumors were found, one each in a low- and high-dose
animal treated with extract. The authors proposed that the increase in cell proliferation caused the increase
in size of foci. An increase in leukocyte infiltration in animals treated with the highest concentration of
extract was higher after histological examination of the livers of mice treated with extract compared to
those receiving a low concentration.
6.3.2.2 Other Routes of Exposure
Groups of 9-16 male Fischer 344 rats, 7 weeks of age, were given a single i.p. injection of 0 or 200 mg/kg
of iV-nitrosodiethylamine (NDEA) in saline followed 3 weeks later by i.p. injections of 0, 1 or 10 |ig
microcystin-LR/kg twice a week for 5 weeks in a study by Nishiwaki-Matsushima et al. (1992). The
doses of microcystin used did not appear to cause liver damage based on the absence of an increase in
hepatic AST. Phenobarbital (0.05%) in the diet was used as a positive control. At the end of week 8, the
rats were sacrificed, the livers removed and evaluated for GSTP-Foci (both the number of lesion and the
foci area). GST-P foci are considered to be biomarkers for early stage development of potential liver
tumors. All animals receiving DEN had foci; those receiving 10 mg/kg microcystin had significantly (p
<0.01) more foci than the control receiving saline. At the low microcystin dose the differences from
control were not significant. The two groups receiving microcystin alone (1 or 10 j^ig/kg) had no GST-P
foci. The group receiving DEN with phenobarbital as a promoter had the largest number and area of foci.
Accordingly, microcystin showed the properties of a promoter but not an initiator.
In a second part of the study, 4 groups of animals were given NDEA injections as above (Nishiwaki-
Matsushima et al., 1992). One of the groups received no microcystin; the other three received 10 j^ig/kg by
i.p. injection. After 3 weeks the animals received a partial hepatectomy to stimulate tissue repair and
received injections of 0, 10, 25, or 50 (.ig microcystin-LR/kg twice a week for 5 weeks. After the partial
hepatectomy, there was a significant dose-related increase in the number and area of foci compared to the
control not treated with microcystin (p <0.01 or 0.001). The last group of rats received initial 10 j^ig/kg
microcystin injections followed by a 50 j^ig/kg dose after the post partial hepatectomy. Those animals had
a mean number of 0.4 ± 0.3 foci/cm2 and an area of 0.1 ± 0.02 % compared with the NDEA control of
13.4 ± 44.2 foci/cm2 with and area of 2.6 ± 3.1%. The evidence from this part of the study also indicates
that microcystin has little if any initiating potential but can promote the formation of preneoplastistic foci
in the liver of exposed rats (Nishiwaki-Matsushima et al., 1992).
Groups of male Fischer 344 rats (n = 5 to 20), 7 weeks of age, received a single i.p. injection of 0 or 200
mg NDEA/kg in saline followed 2 weeks later by 20 i.p. injections of 0 or 25 |ig microcystin-LR/kg (Ohta
et al., 1994). The study design resembled that of Nishiwaki-Matsushima et al. (1992) discussed above.
Animals treated with NDEA plus microcystin-LR had significant (p<0.005) increases in the number, area,
and volume of GST-P-positive foci per liver compared to NDEA-treated rats. The number of foci from
the animals treated with microcystin-LR alone were six-fold lower than those treated with NDEA alone.
The area and volume of the foci were a tenth of those with NDEA alone. The authors concluded that
microcystin was a tumor promoter rather than a carcinogen.
In a study by Ito et al. (1997b), thirteen male ICR mice, 5 weeks of age, received 100 i.p. injections of 20
(ig/kg-bw of microcystin-LR (five times a week) over 20 weeks and were sacrificed after the end of the
treatment (five mice) or after a 2-month withdrawal period (eight mice). Three non-treated mice were
Health Effects Support Document for Microcystins - June, 2015
58
-------
used as controls. Using the 1980 Guidelines on the Histology Typing of Liver Tumors in Rats by the
National Research Council, neoplastic nodules were found in the liver of all 13 treated mice. These
guidelines have since changed in that some types of nodules once considered as preneoplastic no longer
indicated an increased cancer risk (Wolf and Mann, 2005). Re-examination of the original histopathology
records is required to determine if the original findings can be confirmed.
Sekijima et al. (1999) used a similar approach in evaluating whether microcystin-LR is a tumor initiator,
promoter or both. In their study DEN or Aflatoxin B1 served as initiators. Groups of 5-15 male Fischer
344 rats, 6 weeks of age, received an i.p. injection of 0, 200 mg DEN/kg, or 0.5 mg aflatoxin Bi/kg two
weeks before i.p. injections of 0, 1 or 10 |ig microcystin-LR/kg twice a week for 6 weeks. Other groups
were also treated with aflatoxin Bi plus DEN before microcystin-LR treatment. A subset of each
treatment scenario was given partial hepatectomy one week after initiation of microcystin-LR
administration. There was no statistically significant difference in the number of GST-P positive foci and
their area for the DEN Control and those that received both 1 (.ig/kg. and 10 j^ig/kg microcystin-LR
without the hepatectomy. For those that received 10 j^ig/kg and the hepatectomy, the number and area of
foci increased, but were not significantly higher than the DEN control. Combining Aflatoxin Bl with
DEN resulted in foci numbers and areas significantly greater that the DEN control. With addition of
microcystin-LR at 1 or 10 |_ig/kg. the number and area of foci increased but the increase was not
significant. No foci were observed in livers of animals treated with only microcystin-LR at 10 (.ig/kg. A
combination of Alflatoxin Bl with microcystin-LR |_ig/kg and no hepatectomy resulted in a small number
of foci (0.31 /cm2 and an area of 0.05 mm2/cm2) as compared to the DEN alone control (2.46 /cm and an
areapf 13.6 mm2/cm2).
6.4 Other Key Data
6.4.1 Mutagenicity and Genotoxicity
The available data on mutagenicity and genotoxicity of cyanobacterial toxins, including microcystins, has
been recently reviewed (Zegura et al., 2011). These authors concluded that current evidence indicates that
the microcystins are not bacterial mutagens and that discrepancies in results from cyanobacterial extracts
are likely due to differences in source of the cyanobacteria and composition of the complex extract
mixtures. Both in vitro and in vivo genotoxicity studies have shown positive results with DNA damage
induced by formation of reactive oxygen species as well as inhibition of repair pathways. These studies
are summarized below and listed in Tables 6-9, 6-10, and 6-11.
6.4.1.1 Mutagenicity
Ding et al. (1999), and Huang et al. (2007), did not find that pure microcystin-LR induced mutations in
the Ames assay (strains TA97, TA98, TA100, and TA102), either with or without metabolic activation.
Extracts from Microcystis exhibited mutagenic activity in the absence of activation, which was decreased
slightly in TA98 with activation. A crude toxin extracted fromM aeruginosa did not induce mutations in
the Ames assay (strains TA98 and TA100) with and without activation (Grabow et al., 1982). Wu et al.
(2006) used three assays (ara test in E. coli UC1121, Ames test in S. typhimurium strains TA98 and
TA100, and SOS/umu test in S. typhimurium TA1535/pSK1002) to test the mutagenicity of microcystin-
LR extracted from aM. aeruginosa bloom. All tests were negative with and without metabolic activation.
Repavich et al. (1990) reported that Ames assays (using strains TA98, TA100 and TA102) of a purified
hepatotoxin (supplied by Wright State University and presumed to be microcystin) were negative with
and without metabolic activation, as were Bacillus subtilis multigene sporulation assays.
In contrast, Suzuki et al. (1998) reported increased ouabain resistance mutation frequency in human
embryo fibroblast cells treated with microcystin-LR (purity not specified). Similarly, Zhan et al. (2004)
Health Effects Support Document for Microcystins - June, 2015
59
-------
observed a 5-fold increase in the frequency of thymidine kinase mutations when human lymphoblastoid
TK6 cells were treated with commercially-obtained microcystin-LR over control. More slow-growing
mutants were observed than fast-growing mutants, suggesting that microcystin-LR induced large
deletions, recombinations or rearrangements and that the mutation damage was larger than the TK locus.
The differences in mutagenicity response between bacteria and human cell lines may be related to
differences in the cell uptake of microcystin-LR. For example, the failure of microcystin-LR to induce
mutations in bacterial cells may be related to poor uptake. Zhan et al. (2004) observed that microcystin-
LR is not taken up by many cell types, including bacteria. However, no references to support this
assertion were provided by the authors. While hepatocytes take up microcystin-LR at a significant rate,
other cell types show limited or no uptake unless measures are taken to enhance the penetration of the
cells by microcystin-LR.
Table 6-9. Mutagenicity Assays with Microcystins
Species
(test system)
End-point
With
metabolic
activation
Without
metabolic
activation
Reference
Ames assay
Gene mutation; Pure
microcystin-LR; extracts
containing microcystins
-
-
Ding et al., 1999;
Huang et al., 2007
Ames assay
Gene mutation; Crude extract
-
-
Grabow et al.,
1982
Ames assay; ara
test; SOS/umu test
Gene mutation; microcystin-LR
extract
-
-
Wu et al., 2006
Ames assay
Gene mutation; Purified
hepatotoxin assumed to be
microcystin
-
-
Repavich et al.,
1990
Human embryo
fibroblast cells
Gene mutation; microcystin-LR
(purity not specified)
Not applicable
+
Suzuki et al., 1998
Human
lymphoblastoid TK6
cells
Gene mutation; 5x increased
frequency of thymidine kinase
mutations; induction of
micronuclei
Not applicable
+
Zhan et al., 2004
Shi et al. (2011) showed that microcystin-LR could interact with isolated plasmid DNA (4361 base
pairs) using atomic force microscopy combined with UV and fluorescence quenching in the presence of
ethidium bromide. The results eliminated the potential for intercalation binding and electrostatic
interactions with the DNA phosphate backbone and are most consistent with electrostatic interactions
between the microcystin-LR and exposed bases in the minor groove. In the presence of microcystin-LR,
the plasmid DNA aggregated into rod-like structures. The authors hypothesized that this might be the
result of electrostatic repulsion between the DNA double helix strands because of the interactions with
microcystin-LR.
6.4.1.2 Genotoxicity - in vitro studies
Recent studies suggest that apoptosis may be intimately linked to observations of DNA damage in cells
treated with microcystin-LR. Lankoff et al. (2004a) showed a strong correlation between DNA damage,
as measured by the comet assay, and the induction of apoptosis, as measured by the terminal
deoxynucleotidyl transferase-mediated dUTP-biotin nick end-labeling (TUNEL) assay, in human
lymphocytes. Other evidence has suggested that the comet assay can give a false positive measure of
DNA damage when apoptosis is induced, as DNA fragmentation occurs during the process of apoptosis
(Lankoff et al., 2004a). The authors postulated that earlier reports of DNA damage measured by the
Health Effects Support Document for Microcystins - June, 2015
60
-------
comet assay may have been related to early stages of apoptosis due to cytotoxicity rather than a direct
effect on DNA. The induction of apoptosis appears to be dose-related. Humpage and Falconer (1999)
showed that low (picomolar) concentrations of commercially-obtained microcystin-LR induced
cytokinesis and inhibited apoptosis in primary mouse hepatocytes, while higher (nanomolar)
concentrations resulted in opposite effects. Ding et al. (1999) showed DNA damage in primary rat
hepatocytes by the Comet assay at 1 |_ig microcystin-LR/mL.
Nong et al. (2007) observed a dose-dependent increase in test tail DNA using the Comet assay in HepG2
cells incubated with 1-100 |iM microcystin-LR (purity not reported) for 24 hours; the 30 and 100 |iM
concentrations yielded statistically significant results . Zegura et al. (2006) also found a significant
increase in the proportion of tail DNA (indicating DNA damage in the Comet assay) in HepG2 cells
incubated with microcystin-LR (purity not reported) for up to 16 hours. Buthionine sulfoximine (BSO)
pretreatment increased the susceptibility to microcystin-LR induced DNA damage, while pretreatment
with the glutathione precursor N-acetylcysteine protected against the microcystin-LR induced DNA
damage.
In a study with a similar design using HepG2 cells, Zegura et al. (2008a) observed elevation of p53 and
the down regulated genes p21 and gadd45a, which are responsible for cell cycle arrest and DNA repair, as
well as mdm2, which is a feedback regulator for p53 expression and activity. The study authors concluded
that these findings indicate that microcystin-LR has genotoxic potential. Zegura et al. (2008b) evaluated
the genotoxic effects of microcystin-LR (purity not reported) on different cell types using the Comet
assay. Three human cell lines were used: CaCo-2, which is a human colon adenocarcinoma cell line;
IPDDC-A2, which is a human astrocytoma cell line; and NCNC, which is a human B-lymphoblastoid cell
line. A significant increase in DNA damage was only observed in CaCo-2 cells. Zegura et al. (2011)
observed DNA damage using the Comet assay in human peripheral blood lymphocytes at concentrations
of 0.1 to 10 |ig/m L of microcystin-LR (purity not reported). As was previously observed in HepG2 cells,
DNA damage-responsive gene p53 was upregulated along with its downstream-regulated genes involved
in DNA repair and cell cycle regulation, mdm2, gadd45a, and p21. DNA fragmentation was significantly
increased in rat neutrophils with microcystin-LA and microcystin-YR, but not in human neutrophils
(Kujbida et al., 2008).
Bouai'cha et al. (2005) reported that noncytotoxic concentrations of microcystin-LR slightly decreased the
amount of endogenously formed DNA adducts compared with controls in cultured hepatocytes.
Microcystin-LR was shown to cause a dose- and time-dependent increase in the formation of 8-oxo-7, 8-
dihydro-2'-deoxyguanosine (a measure of oxidative DNA damage) in cultured hepatocytes (Maatouk et
al., 2004; Bouai'cha et al., 2005).
Lankoff et al. (2004a) observed no effect of microcystin-LR on the incidence of chromosomal aberrations
in human peripheral blood lymphocytes. In a separate study by Lankoff et al. (2006a) microcystin-LR
inhibited repair of gamma-induced DNA damage in human lymphocytes and a human glioblastoma cell
line.
Observations of polyploidy in microcystin-LR-treated cells (Humpage and Falconer, 1999; Lankoff et al.,
2003) may be related to its effects on cytokinesis. Lankoff et al. (2003) showed that microcystin-LR,
through its effect on microtubules, damages the mitotic spindle, leading to the formation of polyploid
cells. Repavich et al. (1990) reported a dose-related increase in chromosome breakage in human
lymphocytes exposed to a purified hepatotoxin (presumed to be a microcystin). Microcystin-LR disrupted
chromatin condensation in Chinese hamster ovary cells at the end of interphase and the beginning of
metaphase (Gacsi et al., 2009).
Neither microcystin-LR nor cyanobacterial extracts resulted in an increase in micronucleus formation in
cultured human lymphocytes (Abramsson-Zetterberg et al., 2010).
Health Effects Support Document for Microcystins - June, 2015
61
-------
6.4.1.3 Genotoxicity - in vivo studies
A number of studies have reported DNA damage after microcystin-LR treatment in vivo. microcystin-LR
was shown to cause a dose- and time-dependent increase in the formation of 8-oxo-7,8-dihydro-2'-
deoxyguanosine (a measure of oxidative DNA damage) in rat liver cells after in vivo treatment via i.p.
injection (Maatouk et al., 2004; Bouai'cha et al., 2005).
Gaudin et al. (2008) observed DNA damage in female mice administered microcystin-LR (>95% pure)
via either oral or i.p. injection. Groups of three female Swiss albino mice were administered a single
gavage dose of 0, 2, or 4 mg/kg or a single i.p. dose of 10, 25, 40, or 50 (ig/kg and sacrificed 3 or 24
hours after treatment. DNA damage was assessed in whole blood, bone marrow, liver, kidney, colon, and
intestine using the comet assay. Clinical observations were not reported. After oral administration, a
statistically significant dose-dependent increase in DNA damage was observed in blood from both dose
groups at three hours, but not at 24 hours; no effects were seen in the other tissues assayed. After i.p.
exposure, DNA damage was found at doses >40 j^ig/kg only in bone marrow after 3 hours; after 24 hours
DNA damage was found in kidney, intestine and colon at >25 j^ig/kg with the most pronounced effect
being a dose-related increase in the liver at all doses. In contrast, Gaudin et al. (2009) did not find any
DNA damage as assessed by the Comet assay and unscheduled DNA synthesis in the livers of female rats
administered 12.5-50 |_ig microcystin-LR/kg (commercial product; purity not reported) via intravenous
injection.
Dong et al. (2008) evaluated the genotoxicity of microcystin-LR (source and purity not provided) in
mouse testes. Male KM mice were administered 0, 3, 6, or 12 |ig/kg of microcystin-LR daily for seven
days. Five mice/treatment were sacrificed on day 8 and their testes were removed for analysis. Fourteen
days after injection, five mice per treatment were also sacrificed to evaluate the micronuclei in the sperm
cell early stage. An increase in micronuclei and DNA-protein crosslinks was observed with all doses
(highest dose lower than mid dose, no dose response), but only the 6 and 12 |ig/kg treatments were
statistically different from controls.
Li et al. (201 lb) administered (i.p.) crude extracts from a cyanobacterial bloom containing 244.26 |ig
microcystin-LR per gram of lyophilized algae to male mice and observed a dose-dependent increase in
olive tail moment from the Comet assay in the liver and testes. microcystin-YR has also been found to
induce DNA damage measured by the Comet assay in the blood (lymphocytes), liver, kidney, lung,
spleen, and brain of mice administered 10 |ig/kg of microcystin-YR via i.p. injection every other day for
30 days (Filipic et al., 2007).
Neither microcystin-LR nor cyanobacterial extracts resulted in an increase in micronucleus formation in
erythrocytes from peripheral blood of mice given up to 55 j^ig/kg (Abramsson-Zetterberg et al., 2010).
However, Zhang et al. (201 la) observed a significant increase in the frequency of micronuclei in
polychromatic erythrocytes (PCEs) in the bone marrow of rabbits (6/treatment group) administered 6
(ig/kg-day microcystin from an extract ofM aeruginosa via i.p. injection for 7 or 14 days. The
microcystin extracts contained >80% total microcystin, with 0.84 mg/g dry weight microcystin-RR, 0.50
mg/g dry weight microcystin-LR, and 0.07 mg/g dry weight microcystin-YR. There was also a significant
decrease in PCEs/total erythrocytes. Similarly, dose-related increased micronuclei formation were seen in
bone marrow from male mice given 1-100 mg extract/kg (Ding et al., 1999).
Health Effects Support Document for Microcystins - June, 2015
62
-------
Table 6-10. Genotoxicity of Microcystins In vitro
Species
(test system)
End-point
Results
Reference
Primary rat
hepatocytes
Liver DNA
DNA damage with microcystin extract
containing microcystin-LR.
Ding et al., 1999
Rat hepatocytes
DNA adducts
Noncytotoxic concentrations of
microcystin-LR slightly decreased
endogenously formed DNA adducts
Bouai'cha et al.,
2005
Rat hepatocytes
DNA adducts
microcystin-LR caused oxidative DNA
adducts
Maatouk et al.,
2004; Bouai'cha et
al., 2005
Primary mouse
hepatocytes
DNA damage
Commercial microcystin-LR induced
cytokinesis and inhibited apoptosis at
picomolar concentrations; nanomolar
concentrations resulted in inverse
effects
Humpage and
Falconer, 1999
HepG2 cells
DNA damage
Microcystin-LR increased comet test tail
moment.
Nong et al., 2007;
Zegura et al., 2006
Human hepatoma
cells
Liver DNA and repair
DNA damage with microcystin-LR;
elevated p53 and down regulated p21
and gadd45a
Zegura et al., 2003;
2004; 2008a
CaCo-2, IPDDC-A2,
and NCNC human
cell lines
DNA damage
Microcystin-LR increased DNA damage
only in CaCo-2 cells.
Zegura et al., 2008b
Human and rat
neutrophils
DNA damage
Microcystin-LA and microcystin-YR
increased DNA fragmentation in rat, but
not human, neutrophils.
Kujbida et al., 2008
Human lymphocytes
DNA damage
Microcystin-LR caused DNA damage
and induction of apoptosis but no
chromosome aberrations.
Lankoff et al., 2004a
Human lymphocytes
and glioblastoma
cell line
DNA damage and
repair
No micronuclei formation in
lymphocytes; inhibited repair of gamma-
induced DNA damage.
Lankoff et al., 2006a
Human lymphocytes
DNA damage
Microcystin-LR caused DNA damage
and up regulation of damage-
responsive genes
Zegura et al., 2011
Human lymphocytes
DNA damage
Dose-related chromosome breakage.
Repavich et al.,
1990
Chinese hamster
ovary cells
Cell cycle
Microcystin-LR disrupted chromatin
condensation.
Gacsi et al., 2009
Human lymphocytes
Micronucleus formation
No increase with microcystin-LR or
extract.
Abramsson-
Zetterberg et al.,
2010
Health Effects Support Document for Microcystins - June, 2015
63
-------
Table 6-11. Genotoxicity of Microcystins In vivo
Species
(test system)
End-point
Results
Reference
Mouse
Liver DNA
DNA damage after treatment with
microcystin-LR
Rao and
Bhattacharya, 1996
Mouse
DNA damage
microcystin-LR caused damage in
blood cells after oral; in liver, kidney,
intestine, and colon after i.p.; none in
liver after i.v.
Gaudin et al., 2008;
2009
Mouse
DNA damage
Microcystin-LR extract i.p. caused
dose-dependent olive tail moment in
liver and testes.
Li et al., 2011b
Mouse
DNA damage
Microcystin-YR given i.p. induced
damage in multiple organs.
Filipic et al., 2007
Mouse
DNA damage
Increased DNA-protein crosslinks
and micronuclei in testes with
microcystin-LR.
Dong et al., 2008
Rat
Liver DNA
Oxidative damage after i.p. injection
of microcystin-LR
Maatouk et al.,
2004; Bouai'cha et
al., 2005
Mouse bone
marrow
erythrocytes
DNA damage
Induction of micronuclei with
microcystin extract
Ding et al., 1999
Mouse
erythrocytes;
peripheral blood
DNA damage
No induction of micronuclei with
microcystin-LR or extract.
Abramsson-
Zetterberg et al.,
2010
Rabbit bone
marrow
DNA damage
Extract containing microcystin-RR, -
LR, and -YR increased frequency of
micronuclei in PCEs
Zhang et al., 2011a
6.4.2 Physiological or Mechanistic Studies
6.4.2.1 Noncancer Effects
Mechanistic studies, including in vivo investigations in laboratory animals, in situ studies in isolated
perfused organ systems and in vitro assays in isolated cell preparations have been conducted to
characterize the toxicology of microcystins. These studies have evaluated many aspects of microcystin
toxicity, including: 1) interaction with serine and threonine protein phosphatases (i.e., PP1 and PP2A) as
the molecular target for microcystins, 2) the role of cytoskeletal effects, 3) apoptosis, 4) the importance of
oxidative stress as a mode of toxic action, and 5) the reasons for target organ and cell type specificity of
microcystins. Each of these topics is discussed in further detail below.
6.4.2.1.1 Protein Phosphatase Inhibition
Protein phosphatase enzymes PP1 and PP2A has been identified as the primary molecular target of
microcystins. Protein phosphatases function in the post-translational modification of phosphorylated
cellular polypeptides or proteins. PP1 and PP2A groupings belong to the PPP family of protein
phosphatases, which hydrolyze the ester linkage of serine and threonine phosphate esters. Both enzyme
groupings have a single catalytic unit which is joined to a variety of regulatory and targeting subunits.
There are approximately 1,000 protein phosphatase genes in higher eukaryotes which confer considerable
regulatory diversity to the individual super families (Barford et al., 1998).
Health Effects Support Document for Microcystins - June, 2015
64
-------
The actions of members of the protein kinase family of enzymes precede that of the protein phosphatases
because they esterify phosphates to the hydroxyl functional groups of serine, threonine and tyrosine in
proteins. Together, kinases and phosphatases maintain the balance of phosphorylation and
dephosphorylation for key cellular proteins involved in a variety of activities including transport and
secretory processes, metabolic processes, cell cycle control, gene regulation, the organization of the
cytoskeleton, and cell adhesion (Barford et al., 1998).
Immunoprecipitation, X-ray crystallography, autoradiography, nuclear magnetic resonance (NMR)
solution structures, reverse phase liquid chromatography, and molecular dynamics simulation have been
used to evaluate the molecular interaction between microcystins and protein phosphatases (Runnegar et
al., 1995b; MacKintosh et al., 1995; Goldberg et al., 1995; Craig et al., 1996; Bagu et al., 1997; Mattila et
al., 2000; Mikhailov et al., 2003; Maynes et al., 2004, 2006). Molecular modeling and molecular
dynamics simulations have reported that microcystins bind in a Y-shaped groove containing the catalytic
site on the surface ofPPl (Mattila et al., 2000). Studies withPPl suggest that the C-terminal (312-(313
loop ofPPl (with residues 268-281) is important for microcystin-protein phosphatase interactions as well
as for substrate recognition (Maynes et al., 2004, 2006). Current information indicates that the binding
process primarily involves the amino acids Adda, leucine, Mdha and glutamate of microcystins.
According to Craig et al. (1996), microcystins LR, LA and LL interact with the catalytic subunits ofPPl
and PP2A in two phases: the first phase, a rapid inactivation of the phosphatase occurs within minutes;
the second, a slower phase represented by a covalent interaction that takes place within several hours. The
initial binding and inactivation of protein phosphatases appears to result from several non-covalent
interactions that are still under investigation. Mattila et al. (2000) showed an interaction of the glutamate-
free carboxylate of microcystin-LR with a metal ion, either iron or manganese, in the PP1 catalytic site.
Glutamate appears to be an important component since the esterification of the carboxylate functional
group eliminates toxicity (Namikoshi et al., 1993; Rinehart et al., 1994). A review of the mechanisms of
microcystin toxicity demonstrated that the Adda side chain may be involved in a hydrophobic interaction
between the tryptophan 206 and isoleucine 130 residues in the hydrophobic groove ofPPl (Herfindal and
Selheim, 2006).
Microcystins Adda amino acid residue plays an important role in the inhibition of protein phosphatase
activity (Nishiwaki-Matsushima et al., 1991; Gulledge et al., 2002, 2003a,b). Mattila et al. (2000)
suggested that the long hydrophobic side chain of the Adda residue may guide the toxin into the
hydrophobic groove of the catalytic site. The toxic activity of microcystins is eliminated by the
isomerization of the diene from 4E,6E to 4E,6Z on the Adda chain (Harada et al., 1990; Nishiwaki-
Matsushima et al., 1991; Stotts et al., 1993). Those microcystin analogues with only Adda and one
additional amino acid are capable of substantial inhibition of PP1 and PP2A, while modifications to the
Adda structure abolished the inhibition (Gulledge et al., 2003b). Herfindal and Selheim (2006) indicated
that the L-Leucine of microcystin-LR plays an important role in the hydrophobic interaction with
Tyrosine 272 ofPPl (on the (312-(313 loop).
During the second phase of interaction between microcystins and protein phosphatase, covalent bonding
occurs (Craig et al., 1996). Immunoprecipitation and autoradiography methods indicate that a covalent
bond results from the interaction between the thiol of Cys273 residue of PP 1 and the methylene of the
Mdha residue of microcystins. X-ray crystallography data on the microcystin-LR/PPl complex and NMR
solution structures illustrate the covalent linkage at Cys273 (Goldberg et al., 1995; Bagu et al., 1997).
Site-directed mutagenesis replacing Cys273 in PP1 results in a loss of microcystin binding (MacKintosh
et al., 1995; Maynes et al., 2004). Based on sequence similarity between PP1 and PP2A, Craig et al.
(1996) suggested that Cys-266 is the site of a covalent linkage between PP2A and microcystins.
Microcystin analogues with a reduced Mdha residue are not able to covalently bind to protein
phosphatases. MacKintosh et al. (1995) indicated that a modification of the Mdha residue of microcystin-
Health Effects Support Document for Microcystins - June, 2015
65
-------
YR by reaction with ethanethiol, abolished covalent binding to PP1. Similarly, Craig et al. (1996) showed
that a decrease of the Mdha residue of microcystin-LA with NaBlrh abolished the covalent binding phase
with PP2A. Maynes et al. (2006) confirmed the lack of covalent interaction by showing the crystal
structure of dihydromicrocystin-LA bound to PP1. Their work demonstrated that the (312-|313 loop of PP1
has a different conformation when the covalent bond is absent, and that other interactions (including
hydrogen bonding) are responsible for the bond between dihydromicrocystin-LA and PP1.
The relevance of covalent bonding between microcystins and protein phosphatases to enzyme inhibition
is unknown because other interactions are apparently responsible for the rapid inactivation of the enzymes
(Herfindal and Selheim, 2006). The modifications to either molecule (microcystin or protein phosphatase)
to prevent covalent bonding, usually decrease but do not eliminate the toxic action (Meriluoto et al., 1990;
MacKintosh et al., 1995; Hastie et al., 2005).
Under both in vivo and in vitro conditions, microcystins bind to the phosphatase enzymes, resulting in an
inhibition of enzyme activity and leading to a decrease in protein dephosphorylation. Microcystins have
been shown to directly inhibit the activity of PP 1 and PP2A derived from different species such as fish,
mammals, plants, and different cell types (cultured cell lines as well as isolated tissue cells) (Honkanen et
al., 1990; MacKintosh et al., 1990; Matshushima et al., 1990; Yoshizawa et al., 1990; Sim and Mudge,
1993; Xu et al., 2000; Leiers et al., 2000; Becchetti et al., 2002). Microcystin has also been related to
binding to PP4, another member of the protein phosphatase family (Imanishi and Harada, 2004).
Ito et al. (2002b) observed a similar degree of inhibition of protein phosphatases 1 and 2A in vitro with
microcystin-LR and its glutathione and cysteine conjugates. However, Metcalf et al. (2000) demonstrated
weaker inhibition of PP1 and PP2A in vitro by microcystin glutathione, cysteine-glycine, and cysteine
conjugates than by parent microcystins; these conjugates also are less toxic in the mouse bioassay than
parent microcystin. As noted in Section 5.3, Kondo et al. (1992, 1996) postulated that the Adda and Mdha
moieties could be the sites of CYP oxidation and subsequent conjugation with glutathione or cysteine.
Microcystins have been used as a tool to investigate the importance of serine and threonine
phosphorylation in specific cellular functions. The regulatory effects of phosphorylation on the sodium
channel proteins increases the probability of the channel being open in renal cells (Becchetti et al., 2002).
Phosphorylation appears to inhibit ATP-dependent actin and myosin interaction in smooth and skeletal
muscle contraction (Hayakawa and Kohama, 1995; Knapp et al., 2002) and increase insulin secretion
(Leiers et al., 2000).
Several in vitro studies indicate that low levels of microcystins can upregulate protein phosphatase
mRNA expression such that protein phosphatase activity is increased rather than decreased. Liang et al.
(2011) used the FL amniotic epithelial cell line to test the effects of microcystin-LR and found that
incubation for 6 hours with low concentrations of microcystin-LR (0.5 or 1 (j,M) caused increases in
PP2A activity. However, incubation for 24 hours with higher concentrations (i.e., 5 or 10 |iM) caused a
decrease in PP2A activity. The authors reported that the increases in PP2A activity were due to the up-
regulation of mRNA and protein levels of the C subunit. Fu et al. (2009) and Xing et al. (2008) found
comparable up regulation of PP2A in FL cells at comparable concentrations and incubation times.
However, no change in PP2A activity or in PP2A subunit expression was observed by Huang et al. (2011)
in the livers of male mice after 7 days of oral exposure with doses up to 186 |ig microcystin-RR/kg.
Li et al. (201 Id) tested the effects of microcystin-LR on PP2A in human embryonic kidney (HEK) 293
cells. PP2A activity was inhibited with concentrations of 5-10 (j,M (only significantly inhibited with 7.5
and 10 |iM). and increased at concentrations <2.5 (j,M (only statistically significant with 1 and 2 |iM).
Treatment with microcystin-LR caused a disassociation between PP2A and its a4 regulatory subunit, at
all concentrations tested. The study authors suggested that disassociation of a4, a PP2A subunit that
regulates activity of PP2A leading to an increase in active PP2A catalytic subunit in the cell, could
Health Effects Support Document for Microcystins - June, 2015
66
-------
explain the higher activity at low concentrations. At higher concentrations the increase in the PP2A
catalytic unit is unable to compete with the inhibitory effects of microcystin-LR.
Not all microcystins are equipotent inhibitors of protein phosphatases. Table 6-12 provides comparative
data of the IC50 values for inhibition of protein phosphatases (IC50S) by microcystin-LR, microcystin-YR,
microcystin-RR and microcystin-LA as reported by several different authors. The table demonstrates that
there is not much consistency in the results. Differences across studies are likely due to variations in
methodology of the individual studies. There is also a lack of consistency in the relative potencies of
individual microcystins across the individual studies.
Table 6-12. Protein Phosphatase Inhibition Activity Among Microcystin Congeners
Reference
IC50 (nM)
MC-LR
MC-LA
MC-YR
MC-RR
PP2A Inhibition
Craig et al., 1996
0.15
0.16
Nishiwaki-Matsushima et al., 1991
0.28
0.78
Matsushima et al., 1990
7.6
4.5
5.8
PP1 Inhibition
MacKintosh et al., 1995
0.2
|P
CM
O
Mixture of PPs
Yoshizawa et al., 1990
1.6
1.4
3.4
6.4.2.1.2 Cytoskeletal Disruption
Protein phosphatase inhibition by microcystins relates to changes in cytoskeletal structure and cell
morphology (Eriksson and Golman, 1993). The cytoskeleton is comprised of a variety of polymeric,
proteinaceous filaments that form a flexible framework for the cell. The cytoskeleton provides attachment
points for organelles within cells, and makes possible communication between parts of the cell and
between cells (Sun et al., 2011). The major cytoskeletal proteins can be broadly categorized (Hao et al.,
2010) as microfilament proteins (e.g. actins and myosin; 7 nm diameter), intermediate filaments (e.g.
keratins, desmins; 10 nm diameter), and microtubules (e.g. dyneins, tubulin; 25 nm diameter). In addition,
there are a broad number of individual proteins that are associated with the microtubules and
microfilaments. Serine-threonine proteases are of critical importance in maintaining cytoskeletal integrity
(Eriksson et al., 1992 a,b) because of their dephosphorylating impact on phosphoprotein-cytoskeletal
precursors.
Several studies using light, electron and fluorescent microscopy have demonstrated the cytoskeletal
effects of microcystins in the liver (Runnegar and Falconer, 1986; Eriksson et al., 1989b; Hooser et al.,
1989, 1991b; Falconer and Yeung, 1992). Ultrastructural changes in rats given a lethal dose of
microcystin include:
• breakdown of the endothelium;
• loss of microvilli in the space between the hepatocytes and sinusoids (known as the Space of
Disse);
• progressive cell-cell disassociation followed by rounding, blebbing and invagination of
hepatocytes;
• a widening of intracellular spaces;
• hemorrhage; and
• loss of lobular architecture (Hooser et al., 1989).
Health Effects Support Document for Microcystins - June, 2015
67
-------
No toxicity effects were noted in liver endothelial cells or Kupffer cells. Other studies of isolated
hepatocytes, actin aggregates were seen at the base of the membrane blebs following microcystin
exposure. As membrane blebs grew larger and were drawn toward one pole of the cell, the microfilaments
were organized toward the same pole, resulting in rosette formation with a condensed band of
microfilaments at the center.
Similar histopathological changes in the rat testes have been described by Chen et al. (2013). Repeated
i.p. dosing showed an increased space between the seminiferous tubules, cell membrane blebbing,
cytoplasmic shrinkage, swollen mitochondria, and deformed nuclei. The transcriptional levels of (3-actin
and (3-tubulin were also significantly decreased.
Studies in primary isolated hepatocytes have demonstrated the morphological and histopathological
changes induced by microcystins that relate to loss of sinusoidal architecture and cytotoxicity (Runnegar
et al., 1981; Runnegar and Falconer, 1982; Aune and Berg, 1986; Ding et al., 2000a). Exposure of
microcystin to hepatocytes in suspension or cultured in a monolayer results in membrane blebbing that
becomes more pronounced and localized in one region of the cell surface. Cells are rounded in
appearance and become dissociated from one another. Microfilaments are reorganized as a compact
spherical body in the vicinity of the blebbing, and the rest of cell is depleted of filamentous actin.
Microcystin-LR disrupts hepatocellular morphology within minutes, leading to loss of sinusoidal
architecture and hemorrhage. Morphological changes in hepatocytes (i.e., blebbing, rounding) occurred
prior to any effect on cell viability (generally measured as decreased trypan blue exclusion) or cell
membrane integrity (measured as LDH leakage or release of radiolabeled adenine nucleotides).
Thompson et al. (1988) described the time course of cellular effects of microcystins (type not specified)
on primary cultures of rat hepatocytes. The cells were isolated, attached in a monolayer, treated with
0.001-10 (ig/mL of microcystin, and then monitored for 24 hours. After 15 minutes, disintegration of the
attachment matrix occurred at the highest microcystin concentration. After one hour, cells clustered in
groups with no extracellular material. Between 2 and 4 hours, cells began to release from the plates. After
these visual effects occurred, LDH was released and was concentration-related.
Several studies have demonstrated that the observed reorganization of microfilaments leading to alteration
of hepatocyte morphology does not appear to be due to effects on actin polymerization (Runnegar and
Falconer, 1986; Eriksson et al., 1989b; Falconer and Yeung, 1992). Instead, microcystins caused a
decrease in the dephosphorylation of cytokeratin intermediate filament proteins (Falconer and Yeung,
1992; Ohta et al., 1992; Wickstrom et al., 1995; Blankson et al., 2000). Toivola et al. (1997) studied the
effects of microcystin-LR on hepatic keratin intermediate filaments in primary hepatocyte cultures. The
authors observed a disruption of the desmoplakin, a cytoskeletal linker protein that connects an
intermediate filament to the plasma membrane, followed by a dramatic reorganization of the intermediate
filament and microfilament networks, resulting in intermediate filaments being organized around a
condensed actin core.
The authors observed that the major target proteins for microcystin-induced hyperphosphorylation include
keratins 8 and 18 and desmoplakin I/II. Keratins 8 and 18 are the major proteins of intermediate filaments
in hepatocytes; desmoplakin I and II attach keratin filaments in epithelial cells to desmosomes,
(complexes of adhesion proteins that function in cell to cell adhesion). Hyperphosphorylation of
desmoplankin I/II leads to loosening of cell junction and loss of interactions with cytoplasmic
intermediate filaments. The hyperphosphorylation of keratin proteins prevents subunit polymerization
resulting in the observed morphological changes. A Ca2+/ calmodulin-dependent kinase may be involved
in regulating the serine-specific phosphorylation of keratin proteins 8 and 18. Kinase-induced
phosphorylation in the absence of phosphatase dephosphorylation leads to the disassembly of the
microfilaments, breakdown of the cytoskeleton and it's anchoring to desmoplankin I and II (Toivola et al.,
1998).
Health Effects Support Document for Microcystins - June, 2015
68
-------
An in vitro study investigated the cell-type specificity of the effects caused by microcystin using isolated
rat hepatocytes, rat skin fibroblasts (ATCC 1213) and rat renal epithelial cells (ATCC 1571) (Khan et al.,
1995; Wickstrom et al., 1995). After exposure to microcystin-LR, the time course of light microscopic
and ultrastructural effects was examined (Khan et al., 1995). After 4 minutes, effects were noted in
hepatocytes, in renal cells after 1 hour, and in fibroblasts after 8 hours. Similar lesions observed in all cell
types included cytoplasmic vacuolization, blebbing, clumping and rounding, loss of cell-cell contact, and
redistribution of cellular organelles. Effects that were seen only in hepatocytes include whirling of rough
ER, dense staining, loss of microvilli, and dilated cristae of mitochondria plus pinching off of membrane
blebs.
Meng et al. (2011) demonstrated that microcystin-LR causes reorganization of the cytoskeletal structure
in the neuroendocrine PC12 cell line. Pretreatment with a p38 MAPK inhibitor blocked the cytoskeletal
alterations as well as the hyperphosphorylation of tau and HSP27. According to the study authors, direct
PP2A inhibition by microcystin-LR and indirect p38 MAPK activation may be responsible for the
hyperphosphorylation of tau and HSP27 causing cytoskeletal disorganization.
In addition, Sun et al. (2011) evaluated the effects of microcystin-LR on cultures of a human liver cell
line (HL7702). As was the case for the PC12 cell line, hyperphosphorylation of Heat Shock Protein 27 in
the presence of microcystin as a phosphatase inhibitor was accompanied by increased activity of several
kinases (p38 MAPK, JNK and ERK1/2) leading to cytoskeleton reorganization. Treatment with kinase
inhibitors reduced the cytoskeletal changes (Sun et al., 2011). Taken together these studies implicate
kinase-induced phosphorylation combined with inhibition of phosphatase removal of key phosphate
moieties from serine or threonine esters as the cause of the cytoskeletal changes. When microcystins are
present in vitro, balance can be partially restored by inhibiting the activity kinases.
The effects of exposure intravenously to microcystin-LR (purified from a bloom) on transcription of
cytoskeletal genes of rats were reported by Hao et al. (2010). The authors observed alterations in
transcription of genes for tubulin, actin, an intermediate filament (vimentin), and six associated proteins
(ezrin, radixin, moesin, MAP lb, tau and stathmin) in the liver, kidney, and spleen. Ezrin, moesin, and
stathmin are tumor-associated genes which may contribute to tumor promotion by microcystins.
The direction and degree of the cytoskeletal protein change depended on time of measurement after
exposure and the organ examined. The effects were most pronounced in the liver. Although there were
numerous changes that occurred in the transcription of the nine cytoskeletal genes, only a few of the
changes were directly correlated with the levels of microcystin in the tissue. Alterations in the
transcription such as an increase of actin, ezrin, and radixin, and a decrease of tau in the liver were
correlated with tissue microcystin levels (microcystin-LR, microcystin-RR, and total levels). Other
apparent trends included a steady increase in vimentin and MAP lb in the liver over time, followed by
progressively lower levels. The levels of tubulin and stathmin in the liver were below control levels by the
end of the experiment.
As explained above, the responses in the liver differed from those in other tissues. In the kidney,
increased transcription of stathmin was significantly correlated with levels of microcystin-RR. In the
spleen, a decrease in transcription of radixin was significantly correlated with the levels of microcystin-
RR or total microcystin. The levels of actin at the time of the final measurement were lower than the
control in both the kidney and the spleen.
Health Effects Support Document for Microcystins - June, 2015
69
-------
6.4.2.1.3 Apoptosis
The ultrastructural changes in hepatocytes observed after exposure to microcystin suggest that cell death
is related to apoptosis and not necrosis. Changes include cell shrinkage (decreased volume and increased
density), condensation of chromatin and segregation of organelles separated by apoptotic microbodies. As
discussed in the previous section, the cytoskeletal damage may be related to these changes (Boe et al.,
1991; Fladmark et al., 1998; McDermott et al., 1998; Ding et al., 2000b; Mankiewicz et al., 2001).
Several studies have investigated the effects of microcystins on signaling pathways involved in rapid
apoptosis (Ding et al., 1998a,b, 2000b, 2001, 2002; Ding and Ong, 2003; Huang et al., 2011; Feng et al.,
2011; Jietal., 2011).
In an abstract of a non-English publication by Lei et al. (2006), rates of apoptosis were approximately 22-
29% in L-02 cells (a hepatic cell line) after incubation with different concentrations of microcystin-LR for
36 hours. However, after 60 hours of treatment with 50 |ig/m L of microcystin-LR, the rates of apoptosis
increased to 80%. ROS levels also increased in a time-dependent manner (from 0.5-12 hours) in male
mice after a single i.p. injection of 55 (ig microcystin-LR/kg (purity not reported). After exposure of male
mice orally to microcystin-RR, apoptosis occurred in the liver (Huang et al., 2011). Reported changes in
protein expression including decreased Bcl-2 (an antiapoptotic regulator) and increased Bax (a
proapoptotic regulator), lead to a significant increase in the ratio of Bax/Bcl-2. These changes are
suggestive of altered regulation of the outer mitochondrial membrane apoptosis channel proteins (Campos
and Vasconcelos, 2010).
Botha et al. (2004) indicated that apoptosis and oxidative stress can be induced in nonhepatic cells by
microcystins. Microcystin-RR changed the concentration of several proteins associated with apoptosis in
FL human amniotic epithelial cells (Fu et al., 2009). LDH leakage and increased apoptotic indices were
observed in the human colon carcinoma cell line (CaCo2) and MCF-7 cells (deficient in pro-caspase-3),
accompanied by increased hydrogen peroxide formation and increased calpain activity. Apoptosis was
also observed in testes cells by Chen et al. (2011) in male mice orally administered low doses of
microcystin-LR. Wang et al. (2013) showed apoptosis in testes of mice given >7.5 |_ig microcystin-LR/kg
by i.p. injection; mRNA expression for Bax, capsase 3 and capsase 8 were upregulated and increased
phosphorylation of p53 and Bcl-2 was noted. Zhang et al. (201 lb) also observed apoptosis in isolated rat
Sertoli cells incubated with 10 |ig/m L of microcystin-LR for 24 hours. Accompanying this were increases
in p53, Bax, and caspase-3, and a decrease in Bcl-2. After 48 hours of exposure to microcystin-LR, Gacsi
et al. (2009) observed a dose-dependent increase in apoptosis in Chinese hamster ovary cells. Ji et al.
(2011) also observed apoptosis in vitro with a rat insulinoma cell line exposed to microcystin-LR for 72
hours.
Microcystins have been shown, both in vitro and in vivo studies, to increase the pro-apoptotic Bax and
Bid proteins, and the expression of p53, and to decrease expression of the anti-apoptotic Bcl-2 protein in
rats (Fu et al., 2005; Weng et al., 2007; Xing et al., 2008; Takumi et al., 2010; Huang et al., 2011; Li et
al., 201 lc) as well as change mRNA levels (Lei et al., 2006; Zegura et al., 2008a; Qin et al., 2010; Zegura
et al., 2011; Li et al., 201 lc). The same concentrations of microcystin-LR that induced Fas receptor and
Fas ligand expression (a critical step in inducing apoptosis), were found to induce apoptosis in HepG2
cells, at both the protein and mRNA level (Feng et al., 2011). Also, microcystin-LR induced nuclear
translocation and activation of the p65 subunit of NF-tcB, a signal transduction protein that controls a
number of cellular processes, many linked to inflammation and apoptosis (Feng et al., 2011). The knock-
down of p65 in HepG2 cells resulted in a reduction in microcystin-LR-induced Fas receptor and Fas
ligand expression and reduced apoptosis, suggesting that microcystin-LR-induced apoptosis is a complex
process involving many cellular signaling proteins.
Health Effects Support Document for Microcystins - June, 2015
70
-------
Opening of the mitochondrial permeability transition (MPT) pores, thereby increasing permeability, is
considered to be a critical rate-limiting event in apoptosis. Ding and Ong (2003) observed an early surge
of mitochondrial Ca2+ in cultured hepatocytes prior to MPT and cell death. Prevention of this Ca2+ surge
by either chelation of intracellular Ca2+, blockage of the mitochondrial Ca2+ uniporter or use of a
mitochondrial uncoupler, prevented MPT and cell death. Electron transport chain inhibitors including
rotenone, actinomycin A, oligomycin or carbonyl cyanide m-chlorophenylhydrazone, also inhibited the
onset of MPT. Microcystin-LR caused the release of cytochrome c through MPT, considered as a
universal step in mitochondrial apoptosis. However, caspases-9 and -3, which are also linked to apoptosis,
were not activated. After exposure to microcystins, the increase in intracellular Ca2+may instead facilitate
the activation of calpain, a calcium- dependent protease (Ding and Ong, 2003).
In an English abstract from a Chinese-language publication by Liu et al. (2011), i.p. administration of 50
|ig microcystin-LR/kg to mice caused an increase in ALT, AST, Bcl-2 protein, and liver ROS levels; a
decrease in mitochondria membrane potential; and a significant DNA ladder indicative of apoptosis.
Administration of a MPT inhibitor, cyclosporin A, 1 hour before injection of microcystin-LR blocked the
effects. The study authors concluded that inhibiting MPT inhibited microcystin-LR-induced apoptosis.
Mitochondrial respiration was decreased in primary hepatocytes and isolated kidney mitochondria
incubated with microcystin-LR (Jasionik et al., 2010; La-Salete et al., 2008). An uncoupling effect on the
mitochondria was observed in both studies, as well as an indication of mitochondrial generated ROS.
In a study by Qin et al. (2010), the role for the endoplasmic reticulum stress pathway is also implicated in
microcystin-LR-induced liver apoptosis in male ICR mice treated i.p. with 20 (ig/kg (> 95% pure). After
measuring mRNA and protein levels of endoplasmic reticulum stress-specific molecules in the liver and
kidney, the authors found an increase in mRNA and protein expression of CHOP (an apoptosis linked
protein) and cleaved capase-12 in the liver where apoptotic cells also were noted. In the kidney, only a
slight inhibition of these proteins and no apoptosis was observed. The authors concluded that Bcl-2 was
down-regulated in the liver and slightly up-regulated in the kidney. Xing et al. (2008) also observed
regulation of CHOP in cells incubated for 24 hours with microcystin-LR.
6.4.2.1.4 Reactive Oxygen Generation Cellular Response
Oxidative stress may play a role in the induction of MPT and the onset of apoptosis. In cultured
hepatocytes exposed to microcystins, an increase in the generation of ROS preceded the onset of MPT,
mitochondrial depolarization, and apoptosis. A dose- and time-dependent increase in ROS and lipid
peroxidation, measured as malondialdehyde formation, was shown to precede morphological changes in
hepatocytes and release of LDH. The addition of deferoxamine or cyclosporine A inhibited the formation
of ROS and delayed the onset of MPT and cell death. Addition of superoxide dismutase prevented
collapse of the cytoskeleton and release of LDH from isolated hepatocytes. Ding et al. (2001) showed that
generation of superoxide and hydrogen peroxide radicals preceded microfilament disorganization and
cytotoxicity. Hepatocellular glutathione levels were affected by microcystins, and administration ofN-
acetylcysteine was shown to protect against cytoskeletal alterations (Ding et al., 2000a).
Lipid peroxidation in the liver of male mice was observed after 2 hours of exposure to a single i.p.
injection of 55 |_ig microcystin-LR/kg (purity not stated) (Wei et al., 2008). The effects of microcystin-LR
on ROS and enzyme activities indicated that microcystin-LR-induced liver injury in mice begins with the
production of ROS, which stimulated the sustained activation of c-Jun N-terminal protein kinase (JNK) as
well as AP-1 and Bid, changes that lead to mitochondrial dysfunction followed by apoptosis and
oxidative liver injury.
The role of glutathione homeostasis and lipid peroxidation in microcystin-induced liver toxicity have
been examined in several studies (Runnegar et al., 1987; Eriksson et al., 1989b; Bhattacharya et al., 1996;
Ding et al., 2000a; Towner et al., 2002; Gehringer et al., 2003a,b, 2004; Bouai'cha and Maatouk, 2004).
Health Effects Support Document for Microcystins - June, 2015
71
-------
Ding et al. (2000a) indicated that exposure to microcystin in isolated hepatocytes resulted in an initial
increase in glutathione synthesis followed by a later depletion of glutathione. Gehringer et al. (2004)
suggest that increased lipid peroxidation induced by microcystins is accompanied by an increase in
glutathione peroxidase, transcriptional regulation of glutathione-S-transferase and glutathione peroxidase
and de novo synthesis of glutathione. An intravenous LD50 (87 |ig microcystin-LR equivalents/kg) of a
crude microcystin extract resulted in a general suppression of GSTs (14 GST isoforms were measured) in
both liver and testes of male rats (Li et al., 201 le). Bouai'cha and Maatouk (2004) found that 2 ng/mL of
microcystin-LR in primary rat hepatocytes caused an initial increase in ROS formation and an increase in
glutathione. The antioxidants, vitamin E, selenium, silymarin, and glutathione provided some protection
against liver toxicity and lethality from microcystins in mice (Hermansky et al., 1991; Gehringer et al.,
2003a,b).
Moreno et al. (2005) reported, in both the liver and kidney of rats treated intraperitoneally with single
doses of microcystin-LR, significant reductions in glutathione peroxidase, glutathione reductase,
superoxide dismutase and catalase, along with increases in lipid peroxidation. Glutathione reductase,
SOD, glutathione peroxidase, and catalase were significantly decreased while nitric oxide synthetase
activity was significantly increased in both the liver and kidney of male mice administered i.p. injections
of 25 (ig microcystin-LR/kg (purified from a bloom of M. aeruginosa) every other day for a month
(Sedan et al., 2010). Increases in MDA (a measure of lipid peroxidation) in the livers of mice
administered crude extracts containing microcystin-LR (estimated dose 2.9 (ig microcystin-LR/kg) by i.p.
injection for 21 days were reported by Li et al. (201 lb). The lower doses applied in the study (0.73 and
1.5 |ig/kg) did not significantly increase MDA levels. There was also no change in SOD in these animals,
but there was a significant decrease in catalase.
Some studies report the absence of lipid peroxidation during microcystin-induced hepatotoxicity. In liver
slices exposed to a cell extract (concentration not given), a time-dependent leakage of LDH, ALT and
AST was observed with no change seen in glutathione content or lipid peroxidation (Bhattacharya et al.,
1996). In addition, Runnegar et al. (1987) suggested that glutathione depletion did not occur until after
morphological changes (i.e., blebbing) were observed suggesting that ROS may not be the initiating
factor for the cytoskeletal changes. This suggestion is supported by Eriksson et al. (1989b) who
concluded that rapid deformation of isolated rat hepatocytes by microcystin-LR was not associated with
alterations in glutathione homeostasis.
Liu et al. (2010) demonstrated that lipid peroxidation was induced in the testes of immature male rabbits
with a single i.p. injection of 12.5 (ig microcystin-LR equivalents/kg of a crude extract. Other indicators
of oxidative stress identified were increased hydrogen peroxide, increased catalase, SOD, glutathione
peroxidase, GST, and GSH.
6.4.2.1.5 Target Organ/Cell Type Specificity
Liver
Most oral and injection studies in laboratory animals have demonstrated that the liver is a primary target
organ for microcystin toxicity. Mechanistic studies suggest that the target organ specificity is directly
related to the limited ability of microcystins to cross cell membranes in the absence of an active transport
system (see section 6.2). Liver toxicity produced by in vitro or in vivo exposures to microcystins was
reduced or eliminated by inhibition of hepatocellular uptake using OATp transport inhibitors (e.g.,
antamanide, sulfobromophthalein and rifampicin) and bile salts (i.e., cholate and taurocholate). Lu et al.
(2008) used OATplb2 null mice to demonstrate the importance of the OATp system for transporting
microcystin-LR into the liver.
Toxicological effects of microcystins in the isolated perfused rat liver were similar to those demonstrated
following in vivo exposure (Pace et al., 1991). During a 60-minute exposure, microcystin-LR caused liver
Health Effects Support Document for Microcystins - June, 2015
72
-------
engorgement and cessation of bile flow. Electron microscopy revealed loss of sinusoidal architecture,
dilation of bile canaliculi and the space of Disse and decreased intracellular contact. Mitochondrial
swelling, disruption of endoplasmic reticulum and formation of whorls and loss of desmosomal
intermediate filaments were also observed. Mitochondrial function was impaired, with inhibition of stage
3 respiration and a decrease in the respiratory control index.
Runnegar et al. (1995b) demonstrated cessation of bile flow, increased perfusion pressure, decreased
protein secretion and decreased glucose secretion following exposure to microcystins. Histological
changes included hepatocyte swelling, loss of sinusoidal architecture, pyknotic nuclei and extensive
necrosis. Exposure to high concentrations of toxin extracts in the isolated perfused liver produced loss of
cord architecture due to hepatocyte disassociation, membrane damage, cytolysis and nuclear effects
(pyknosis, karyokinesis, and karyolysis) (Berg et al., 1988). Ultrastructural effects included swollen
mitochondria, vacuoles, necrosis, abnormal nuclei, bile canaliculi lacking microvilli, and whorls of rough
endoplasmic reticulum.
Studies (Runnegar et al., 1981; Runnegar and Falconer, 1982; Aune and Berg, 1986; Ding et al., 2000a)
show that microcystin exposure to hepatocytes in suspension or cultured in a monolayer results in
membrane blebbing that becomes more pronounced and localized in one region of the cell surface.
Morphological changes in hepatocytes (i.e., blebbing, rounding) have been shown to occur prior to any
effect on cell membrane integrity (measured as LDH leakage or release of radiolabeled adenine
nucleotides) or cell viability (generally measured as decreased trypan blue exclusion).
Similar toxicological effects were observed in isolated human hepatocytes (Yea et al., 2001; Batista et al.,
2003; Thompson et al., 1988). Microcystin-LR produced blebbing, fragmentation and hepatocyte
disassociation. Cytotoxicity, as measured by LDH leakage, occurred after morphological changes were
evident. Yea et al. (2001) indicated that cytotoxicity in human hepatocytes was observed at a
concentration (1 (j,M) that did not affect rat hepatocytes. Batista et al. (2003) also reported a slightly
higher susceptibility to microcystin-induced morphological change in human hepatocytes as compared to
rat hepatocytes. Thompson et al. (1988) described the disintegration of the attachment matrix after 15
minutes, followed by cells clustered in groups with no extracellular material at 1 hour and release of cells
from plates between 2 and 4 hours. LDH release did not occur until after these visual effects and was
dose-related when measured.
After incubation with microcystin-LR in the range of 0.1-50 nM, inhibition of mitochondrial respiration
occurred in primary hepatocytes (Jasionek et al., 2010). The authors indicated changes in ATP levels and
mitochondrial uncoupling, suggesting that microcystin-LR may target electron transport chain (ETC)
complex I function. At noncytotoxic concentrations in HepG2 cells, microcystins interfered with the
metabolism of amino acids, lipids, carbohydrates, and nucleic acids (Birungi and Li, 2011).
Kidney
Nobre et al. (1999, 2001) used an isolated perfused kidney model to evaluate the kidney toxicity of 1 |_ig
microcystin-LR/mL. The authors found that microcystin-LR produced vascular, glomerular and tubular
effects in the exposed kidney. An increase in perfusion pressure was followed by an increase in the
glomerular filtration rate (GFR), increased urinary flow rate and a reduction in tubular transport at the
proximal tubules. Protein in the urinary spaces, although not further described, was observed after
histopathological evaluation. Dexamethazone and indomethacin antagonize the effects of microcystin-LR,
possibly by blocking the microcystin-LR-induced activation of phospholipase A2 and cyclooxygenase.
Nobre et al. (2003) used rat peritoneal macrophages exposed to microcystin-LR in the isolated perfused
kidney model to further investigate the role of inflammatory mediators. The authors observed that
macrophage supernatants from exposed rats caused an increase in renal vascular resistance, GFR and
urinary flow and reduced Na+transport. These effects were reduced by cyclohexamide, dexamethasone
Health Effects Support Document for Microcystins - June, 2015
73
-------
and quinacrine, indicating the involvement of phospholipase A2 and other inflammatory mediators in
microcystin-induced kidney toxicity.
A chronic study performed in male Wister rats with low doses of microcystin-LR and microcystin-YR
reported damage to the kidney cortex and medulla (Milutinovik et al., 2002, 2003). For 8 months, the
authors injected 10 mg/kg i.p. of microcystin-LR and microcystin-YR every second day and found
numerous glomeruli collapsed and the renal tubules filled with eosinophilic protein casts. The tubuli of
the outer and inner medulla were dilated and the lumens filled with eosinophilic proteinaceous casts,
which were described and likely composed of congregated actin filaments. The authors concluded that
their results were consistent with microcystin impact on the cytoskeleton as result on PP2 inhibition.
Tubular cells displayed evidence of both apoptosis and necrosis. A TUNEL assay showed DNA damage
in both the kidney cortex and medulla. Microcystin-LR induced more sever pathological changes than
those induced by microcystin-YR. The authors concluded that long-term microcystin exposures presented
a risk for kidney damage with functional consequences.
Alverca et al. (2009) evaluated the effects of microcystin-LR (>85% pure, extracted from M. aeruginosa
isolated from a bloom) on a kidney cell line (Vero-E6). The viability of the cell line decreased in a time
and dose-dependent manner affecting the cell morphology, with enlarged lysosomes, lysosomal leakage,
damage to mitochondrial structure, disassembly of actin filaments, reduction in the number of intact
lysosomes, and shortening or disappearance of stress fibers observed. Swelling of the endoplasmic
reticulum cisterna, Golgi apparatus vacuolization and a dose- and time related increase in apoptotic cells
were also observed.
Testes
The testes are another target organ for microcystin in in vivo studies on male mice or rats (Li et al., 2008;
Liu et al., 2010; Chen et al., 2011; Wang et al., 2012; Ding et al., 2006; Li et al., 201 lb). With the
exception of the Chen et al. (2011) study, dosing was by i.p. administration. The effects of a single i.p.
injection of microcystin extracts from a surface bloom containing 167.7 |ig microcystin-RR/mL and 47.0
|ig microcystin-LR/mL or 80.5 |_ig microcystin-LR equivalents/mL was found to have an effect on male
rabbit testes. Lesions, including a variety of histological changes to both spermatogonia and Sertoli cells,
were seen in animals treated with 12.5 |ig microcystin-LR equivalents/kg; recovery occurred by 48 hours
with the tissue resembling the control (Liu et al., 2010). Apoptosis has been observed in the testes of rats
and mice given microcystin-LR (Chen et al., 2011; Wang et al., 2013) accompanied by changes in
expression of apoptosis-related genes (Zhang et al., 201 lb; Wang et al., 2013).
The in vitro toxicity of microcystins to Leydig cells and Sertoli cells, demonstrated by decreased cell
viability (Li et al., 2008; Li & Han, 2012; Zhang et al., 201 lb), suggests that microcystin uptake by the
testes maybe similar to that by the liver. OATps (OATp 1A4, 1A5, 2A1, 2B1, 3A1, 6A1, 6B1, 6C1, and
6D1) are active in the testes (Klaassen and Aleksunes, 2010; Svoboda et al., 2011) although no studies
have been located addressing their specific contribution to the testicular toxicity of microcystins.
Augustine et al. (2005) found that OAT3 expression in Sertoli cells and the testes was similar to, or
exceeded, that found in the liver.
Zhou et al. (2012) investigated whether the target membrane transporters could deliver microcystin-LR to
the spermatogonia. The authors isolated mRNA from rat spermatogonia using PCR expansion of the
mRNA pool and primer sequences for OATp 1A4, 1A5, 2B1, 3A1, 6B1, 6C1, and 6 Dl. Cultured cells
were exposed to concentrations of 0, 0.5, 5, 50, or 500 nmol microcystin-LR/L and spermatogonia were
isolated from the testes of 9-10 day old male rats and cultured for examination. A significant
concentration-related decline in cell viability at concentrations > 5 nmol/L was observed. Microcystin
entry into the cell was demonstrated using gel electrophoreses to separate the proteins combined with
targeted Western Blot analysis. Five OATps (1A5, 3A1, 6B1, 6C1, and 6D1) were identified in the
spermatogonia. Microcystin-LR affected the testicular and spermatogonia expression of all the identified
Health Effects Support Document for Microcystins - June, 2015
74
-------
OATps, especially that for OATp 131. Cellular apoptosis, as determined using flow cytometry, increased
at concentrations > 50 nmol/L. After a 6 hour exposure to microcystin-LR, a decrease in total antioxidant
capacity as reflected in increased mitochondrial membrane potential, ROS, and free Ca2+ was observed.
The authors hypothesized that the microcystin-LR inhibited PP1 and PP2 causing oxidative stress and
cytotoxicity, thus impacting sperm production. Members of the protein phosphatase PP1 and PP2 families
along with PP-associated proteins have been identified in testes and sperm, some localized to the sperm
head and others to the tail (Mishra et al., 2003; Fardilha et al., 2013).
To analyze the acute effects of microcystin-LR on gene expression and reproductive hormone levels in
male BALB/c mice, Wang et al. (2012) administered by i.p. 0, 3.75, 7.5, 15, or 30 |ig/kg of microcystin-
LR (purity not reported) for 1, 4, 7, or 14 days. Over the 14 days, the animals in the 15 and 30 |ig/kg
groups lost weight resulting in significantly lower body weight by the end of treatment. No effect on the
expression of Kisspeptin-1 (Kiss-1 which stimulates the reproductive system), GPR54 (a Kisspeptin
receptor), gonadotropin releasing hormone receptor (GnRHR), FSH receptor (Fshr), or luteinizing
hormone receptor (Lhr) was observed. However, after 1, 4, 7, and 14 days, a significant decrease of
GnRH expression at all doses was reported. Fsh(3 was upregulated at 7.5 and 15 |ig/kg. but after 14 days
was significantly decreased at 30 |ig/kg. At all doses, Lh(3 expression was significantly decreased.
Through the 7 days of treatment, changes in gene expression corresponded to increases in FSH, LH, and
testosterone levels followed by decreases in LH and testosterone levels at all doses after 14 days of
treatment. At 15 |ig/kg. FSH levels were significantly increased, but significantly decreased at 30 |ig/kg
after 14 days.
Chen et al. (2013) found that repeated i.p. dosing of rats with 10 |_ig microcystin-LR/kg affected
expression of cytoskeletal genes and mitochondrial dysfunction in the testes. Levels of FSH and LH
increased while testosterone levels decreased. Male reproductive effects were consistently observed after
single and repeated parenteral exposures. These studies are described in more detail in section 7.2.5.
Histological damage to the testes was observed in mice, rabbits, and rats administered microcystin-LR or
a cellular extract (Chen et al., 2013; Li et al., 2008; 201 lb; Liu et al., 2010; Ding et al., 2006). Sertoli
cells were shown to be affected in rabbits and mice, testes and epididymal weights were decreased in
mice and rats, and sperm motility and viability were affected in mice and rats.
6.4.2.1.6 Other Tissues
Soares et al. (2007), Carvalho et al. (2010), and Casquilho et al. (2011) all observed lung damage after a
single i.p. administration of microcystin-LR at a sublethal dose (i.e., 40 |ig/kg). None of the studies
detected microcystin-LR in the lungs but damage was evident within 2 hours of exposure. Lung effects
include an increase in the proportion of areas with alveolar collapse accompanied by an increase in the
percentage of PMN cells; increased impedance; increased oxidative stress in the lung as measured by
decreased SOD, and increased catalase, thiobarbituric acid reactive substances, and myeloperoxidase;
elevated pulmonary mechanical parameters; and increases in TNFa, IL-ip, and IL-6.
Milutinovic et al. (2006) demonstrated that 10 |ig/kg of microcystin-LR administered i.p. every other day
for 8 months to male rats caused microscopic lesions to the heart including disarray and short runs of
myocardial fibers interrupted by connective tissue, increased volume density of interstitial tissue with a
few lymphocyte infiltrations, enlarged cardiomyocytes with enlarged and often "bizarre-shaped" nuclei;
some cells also demonstrated loss of cell cross-striations and degenerative muscle fibers with
myocytolysis. A similar study by the same group using microcystin-YR (Suput et al., 2010) also found
similar histopathological results, but less prominent effects on the heart with microcystin-YR compared to
microcystin-LR. Neither microcystin-LR nor microcystin-YR induced apoptosis in the heart.
Zhao et al. (2015) revealed thyroid dysfunction in mice after i.p. injection of microcystin-LR for 4 weeks.
Mice exposed to either 5 or 20 |ig/kg of microcystin-LR showed an increase in the circulating thyroid
Health Effects Support Document for Microcystins - June, 2015
75
-------
hormone (TH) levels and the free triiodothyronine (FT3). The authors also observed a decreased free
thyroxin (FT4), presumably responsible for the changes observed after exposure. An increased expression
of TH receptor (Tra) and mTOR expression in the brain was also observed and related to a consequence
of the increased FT3. In addition, disrupted glucose, triglyceride and cholesterol metabolism with obvious
symptoms of hyperphagia, polydipsia, and weight loss were also observed.
6.4.2.2 Cancer Effects
Mechanistic evidence provides support for the hypothesis that microcystin-LR can act as a promoter at
low doses due to increased cell proliferation and decreased apoptosis, as well as inhibition of repair. Data
related to cancer and cell proliferation indicate that at low doses, microcystin-LR may increase cell
proliferation. microcystin-LR has been shown to increase the expression of the bcl-2 protein (that inhibits
apoptosis) and decrease the expression of the bax protein (that induces apoptosis) (Hu et al., 2002; Lei et
al., 2006; Weng et al., 2007; Li et al., 201 lc). Further, microcystin-LR upregulates the transcription
factors c-fos and c-jun, leading to abnormal proliferation (Zhao and Zhu, 2003). Gehringer (2004), in a
review of the molecular mechanisms leading to promotion by microcystin-LR and the related tumor
promoter okadaic acid, reported that microcystin-LR inhibits protein phosphatase PP2A, which regulates
several MAPKs. The MAPK cascade regulates transcription of genes required for cell proliferation,
including c-jun and c-fos. In addition, activation of the MAPK cascade has been postulated to inhibit
apoptosis and thus increase cell proliferation. In addition, microcystin-LR has been reported to increase
phosphorylation of p53 (Gehringer, 2004; Fu et al., 2005; Li et al., 2009; Hu et al., 2008; Xing et al.,
2008; Zegura et al., 2008a; Li et al., 201 lc), which is involved in regulation of the cell cycle and
apoptosis.
Clark et al. (2007, 2008) found microcystin-LR administered i.p. at a sublethal dose caused changes in
gene transcription related to actin organization, cell cycle, apoptotic, cellular redox status, cell signaling,
albumin metabolism, and glucose homeostasis pathways, as well as the OATp system in the livers of p53
knockout mice. The gene expression analysis found increases in genes related to cell-cycle regulation and
cellular proliferation in microcystin-LR treated mice livers was greater compared to the p5 3-deficient
mice, control livers and that observed in the livers of microcystin-LR treated wild type mice (Clark et al.,
2008). Ki-67 (a marker of cell proliferation) and phospho-histone H3 (a mitotic marker) for
immunoreactivity were also increased in microcystin-LR-treated knockout mice. The study authors
concluded that p53 may play an important role in tumor promotion by microcystin-LR.
Changes in MMP levels have been linked to cancer and tumor promotion. Zhang et al. (2010; 2012)
found increased levels of MMP2 and MMP9 in the livers of male mice orally administered microcystin-
LR for at least 180 days (subchronic and chronic results of these studies were described in Sections 6.2.3
and 6.2.6, respectively). To study further possible effects of microcystin-LR on tumor metastasis, Zhang
et al. (2012) cultured breast cancer cells with different concentrations of microcystin-LR for different
lengths of time. Acceleration of cell migration was found to be dependent on both the concentration and
length of microcystin-LR incubation time. The levels of MMP2 and MMP9 were also increased with
microcystin-LR concentration in breast cancer cells.
Birungi and Li (2011) tested the effects on noncytotoxic concentrations (1-100 ng/mL) of microcystin-
LR, microcystin-YR, and microcystin-RR on HepG2 cells. While higher concentrations (1000 ng/mL) are
known to cause cell death, cells continued to proliferate at the noncytotoxic concentrations used in this
study. The study authors suggested this could lead to uncontrolled growth and possibly tumors.
Microcystin-LR (10 |ig/L) incubated with WRL-68 cells, a human cell line, for 25 passages had an
increased growth rate compared to controls (Xu et al., 2012). Gan et al. (2010) also found that
microcystin-LR enhanced cell proliferation in the liver cancer cell lines HepG2 and Hep3B. Microcystin-
Health Effects Support Document for Microcystins - June, 2015
76
-------
LR was also found to activate nuclear factor erythroid-2 (Nrf2) in a dose-dependent manner. Inhibiting
Nrf2 also inhibited microcystin-LR-induced cell proliferation.
Nong et al. (2007) incubated HepG2 cells with 100 |iM microcystin-LR for 24 or 48 hours. After both
time periods there was an increase in the number of cells in G0/G1 phase of the cell cycle with less in the
S phase of the cell cycle. ROS scavengers (catalase, SOD, or deferoxamine) did not affect the blockage in
the cell cycle induced by microcystin-LR. The opposite was observed in a kidney cell line.
Dias et al. (2010) studied the effects of microcystin-LR on the proliferation of nonhepatic cells using a
kidney epithelial cell line (Vero-E6). Previous studies (Dias et al., 2009; Alverca et al., 2009) had found
microcystin-LR cytotoxic to this cell line, at doses as low as 11 (iM. Therefore, Dias et al. (2010) used
commercial microcystin-LR (purity >95%) or extracted microcystin-LR (purity not reported, but stated to
have been tested) in the range of 5-5000 nM. Even the lowest concentration caused an increase in ERK1/2
activity, suggesting that microcystin-LR stimulates the Gl/S transition and activates the ERK1/2 pathway
(as noted by increases in p38, JNK, and ERK1/2 activity) in kidney cells.
Zhu et al. (2005) reported that microcystin-LR can transform immortalized colorectal crypt cells,
resulting in anchorage-independent growth and enhanced proliferation. Lankoff et al. (2006b) did not find
any DNA damage in CHO-K1 cells incubated with microcystin-LR (10 or 24 |ig/mL). but microcystin-
LR did inhibit the repair of DNA damage induced by ultraviolet light. The study authors suggested that
microcystin-LR inhibited the nucleotide excision repair through inhibition of the inclusion/exclusion
phase as well as the rejoining phase. In a different study, microcystin-LR inhibited DNA repair by gamma
radiation in human lymphocytes and a human glioblastoma cell line (Lankoff et al., 2006a). Microcystin-
LR and microcystin-RR have been shown to increase the expression of the Bcl-2 protein (that inhibits
apoptosis) and decrease the expression of the bax protein (that induces apoptosis) (Hu et al., 2002; Hu et
al., 2010; Huang et al., 2011; Li et al., 201 lc). However, one study found decreased expression of Bax,
Bcl-2, and bad (pro-apoptotic) proteins (Billam et al., 2008). In addition, microcystin-LR upregulates the
transcription factors c-fos and c-jun, leading to abnormal proliferation (Zhao and Zhu, 2003; Li et al.,
2009).
Xing et al. (2008) observed increases in p53 expression and decreased PP2A expression in FL human
amniotic epithelial cells incubated with 10-1000 nM microcystin-LR for 24 hours. Hu et al. (2008)
observed a significant increase in p53 expression in livers of rats exposed to pure microcystin-LR (purity
not reported) via i.p. injection twice a week for 6 weeks, but did not observe a significant increase in p53
expression with cyanobacterial extracts containing microcystin-LR at a concentration of 529.656 ng/L
administered via the drinking water. Neither treatment altered pl6 expression. Takumi et al. (2010)
studied the role of p53 on cell fate in HEK293-OATP1B3 cells exposed to microcystin-LR. The data
suggested that when p53 is inactivated, chronic low exposure to microcystin-LR could lead to cell
proliferation through activation of Akt signaling. Akt is a general mediator of growth factor induced
survival and has been shown to suppress the apoptotic death of a number of cell types induced by a
variety of stimuli, including growth factor withdrawal, cell-cycle discordance, loss of cell adhesion, and
DNA damage. Fu et al. (2009) also found changes in proteins associated with the cell cycle in human
amniotic epithelial cells exposed to microcystin-RR.
6.4.2.3 Structure-Activity Relationships
With a few exceptions, microcystin congeners exhibit i.p. LD50 values between 50 and 300 |ig/kg in mice
(Rinehart et al., 1994; WHO, 1999). Microcystin-LR is one of the most potent congeners (i.p. LD50
approximately 50 |ig/kg). Pharmacokinetic differences among the various microcystin congeners may be
at least partially responsible for observed variations in lethal potency (Ito et al., 2002b). Microcystin
congeners of varying hydrophobicity were shown to interact differently with lipid monolayers
Health Effects Support Document for Microcystins - June, 2015
77
-------
(Vesterkvist and Meriluoto, 2003). Effects on membrane fluidity could alter the cellular uptake of these
toxins.
Wolf and Frank (2002) proposed toxicity equivalency factors (TEFs) for the four major microcystin
congeners based on LD50 values obtained after i.p. administration. The proposed TEFs, using microcystin-
LR as the index compound (TEF=1.0) were 1.0 for microcystin-LA and microcystin-YR and 0.1 for
microcystin-RR. The application of TEFs based on i.p. LD50 values to assessment of risk from oral or
dermal exposure is questionable given that differences in liphophilicity and polarity of the congeners may
lead to variable absorption by non-injection routes of exposure.
Health Effects Support Document for Microcystins - June, 2015
78
-------
7.0 CHARACTERIZATION OF RISK
7.1 Synthesis and Evaluation of Major Noncancer Effects
Although, some studies have shown differences in toxicity and in uptake rates (Zeller et al., 2011;
Zurawell et al., 2005), the preponderance of toxicological data on the effects of microcystins are restricted
to the microcystin-LR congener. As a result, this section largely describes the available information on
the toxic effects of microcystin-LR.
Elevated liver enzymes have been measured in humans served by a public water supply contaminated
with abloom ofM aeruginosa (Falconer et al., 1983) and in children consuming high levels of
microcystin through water and food (Li et al., 201 la). One study of human exposure to drinking water
before, during and after a bloom of M. aeruginosa reported a significant increase in GGT levels during
the bloom compared with levels before the bloom and compared to the levels in patients living in areas
served by other water supplies (Falconer et al., 1983). The study population consisted of all persons
subjected to liver function tests in the area served by the affected drinking water supply; as such, it is not
fully representative of the general population. A study in China evaluated liver damage in children in
relation to the microcystin levels in the drinking water and select aquatic foods (e.g. carp and duck) (Li et
al., 201 la). Microcystin levels were associated with increasing levels of AST and ALP, but not ALT and
GGT. The Odds Ratio for liver damage, as defined by increased serum enzyme levels in exposed
children, was 1.72 (95% CI: 1.05-2.76).
A major noncancer health effect of exposure to microcystin-LR in animal studies is liver damage. Oral
exposure to single 500 (ig/kg gavage doses of microcystin-LR caused diffuse hemorrhage in the liver of
mice and rats; more pronounced liver damage occurred at higher doses (Ito et al., 1997a; Fawell et al.,
1999). Young mice (5 weeks old) did not develop signs of hepatotoxicity at 500 |ig/kg of microcystin-LR,
while aged mice (32 weeks old) developed clear signs (Ito et al., 1997a). This difference may result in
part from differences in the ontology of the intestinal transporters responsible for gastrointestinal
absorption of microcystins, but cannot be entirely explained by absorption differences, because similar
age-dependent effects were reported after i.p. exposure (Adams et al., 1985; Rao et al., 2005). However,
liver transporters may also show age-related differences in expression.
A 28-day study of oral exposure to 50 or 150 |ig/kg of microcystin-LR in drinking water showed
increased liver weight, slight to moderate liver lesions with necrosis (with and without hemorrhage) and
increased ALP and LDH in rats exposed at 50 (ig/kg-day (Heinze, 1999). A subchronic gavage study in
mice using a similar dose range identified a LOAEL of 200 |ig/kg (Fawell et al., 1999). At this dose, mild
liver lesions, including chronic inflammation and hepatocyte vacuolization were observed. Two animals
had hepatocyte degradation and there were hemosiderin deposits in one liver. Mean serum ALT and AST
were significantly increased in male animals. No adverse effects were identified at a dose of 40
|ig/kg. although mean bodyweight gains were uniformly reduced to the same extent for all treated
animals. The authors expressed that these reductions were within the normal range for this strain of mice
and because no dose response was observed, they considered these findings coincidental. Mild hepatocyte
injury in the area of the central vein was reported in mice given 80 or 100 gavage doses of 80 |ig/kg each
over 28 weeks, corresponding to time-weighted average doses of 33-41 (ig/kg-day (Ito et al., 1997b). No
liver or other toxicity was reported after a mean cumulative microcystin-LR drinking water intake of 35.5
|ig per mouse for 18 months (Ueno et al., 1999).
It is important to consider the route of administration in conjunction with the effects observed after oral
exposures to microcystins. It is known that organic anion transporting polypeptides control uptake of
microcystin from serum into the liver and other organs (Fischer et al., 2005). Less is known about uptake
from the gastrointestinal tract. Given the resistance of microcystins to digestion and their molecular
Health Effects Support Document for Microcystins - June, 2015
79
-------
composition, some form of facilitated transport is likely. Two in vitro studies using human Caco-2
intestinal cells demonstrated that microcystin-LR cellular apical uptake with efflux from the cell
apparently required active transport (Zeller et al., 2011; Henri et al., 2014). Henri et al. (2014) concluded
that basolateral efflux and not apical uptake was the limiting factor for transfer to portal circulation.
Under such circumstances, dosing by drinking water is the preferred route for delivering the dosed
material to serum and subsequently to organs. Dosing from drinking water, wherein exposure occurs
relatively consistently across a day, can deliver a larger portion of the dosed compound to circulation than
gavage dosing. The opportunity for absorption with a bolus dose is limited by the dosing medium,
concentration, and the small intestine transit time.
Delivery to target tissues is also transport controlled and impacted by serum concentration. Uptake as a
proportion of dose by an organ such as the liver, is greater when the serum level is low and constant than
when the level in serum is high and of short duration. These factors become important when contrasting
the results from drinking water studies such as the 28 day study by Heinze (1999) with gavage studies
such as the 90 day study done by Fawell et al. (1999). The same factors must also be considered when
comparing these results with those of Guzman and Solter (1999), wherein an osmotic pump was used to
slowly deliver microcystin-LR directly into the intraperitoneal membrane. Conceptually, one would
expect that the risk for hepatic damage would be greatest in the Guzman and Solter study and lowest in
the Fawell et al. study (1999) under the situation where the total daily doses were the same or similar and
there was a difference in dose delivery (gradual versus bolus).
In the Guzman and Solter (1999) study, the 32 and 48 (ig/kg/day dose caused histological damage to the
liver of male Sprague Dawley rats (3 per dose group) as manifested by inflammation, fibrous tissue, cell
death and apoptosis. Infiltrates of macrocytes, lymphocytes, and neutrophils were seen in the centrilobular
area and round lipid staining vacuoles in the pericentral region. Hepatocellular damage was more severe
in the high dose group than the mid-dose group. Changes in liver enzymes and the concentration of
malondialdehyde increased in a dose related manner for the mid and high doses. The 16 (ig/kg/day dose
did not display any histological damage. The malondialdehyde concentrations suggest that oxidative
stress is part of the pathological changes from exposure to microcystin-LR after 28 days. The leakage of
liver enzymes suggests the inability of the hepatocytes to maintain membrane integrity due to toxin
induced injury. The intraperitoneal infusion route of exposure can account for the fact that the rats in this
study, were vulnerable to liver effects at a lower dose than the animals in Heinze (1999) and Fawell et al.
(1999).
The Fawell et al. (1999) study in groups of 15 male and 15 female Crl:CD-l(ICR)BR (VAF plus) mice
used gavage dosing of microcystin-LR in aqueous solution over a 90 day period. Conceptually gavage
would deliver a lower daily dose to the liver than a drinking water dose, given that the time for serum
uptake will be limited by small intestinal transit time and transporter kinetics. Fawell et al. (1999) used
doses of 0, 40, 200, or 1000 (ig/kg/day. There were no signs of liver damage at the lowest dose and mild
evidence at the 200 (ig/kg/day dose. Chronic inflammation was present in animals from all dose groups as
was hepatocyte vacuolization. There was a dose related increase across all dose groups for chronic
inflammation with 1, 2, 8, and 14 males impacted in the control, low, mid and high dose groups
respectively, and 5, 8, 8, and 14, respectively in females. The hepatic vacuolization did not exhibit a clear
dose-response in females. Hepatocyte degeneration was first reported at 200 (ig/kg/day in two animals
and increased to 23/30 animals at the high dose. Serum ALT and AST were significantly increased at
doses > 200 (ig/kg/day in males and females. Accordingly, the results are consistent with the concept that
once per day gavage dosing allows less of the microcystin-LR to be delivered from the intestines to the
liver.
Considering that the kinetics of i.p. infusion differs from that of drinking water and oral gavage routes of
administration, it is difficult to compare the effects across the three studies. In addition, the Fawell et al.
Health Effects Support Document for Microcystins - June, 2015
80
-------
(1999) study was conducted using mice while those by Heinze (1999) and Guzman and Solter (1999)
used male rats, which further limits comparisons.
The Heinze (1999) 28 day drinking water study of groups of 10 male F1 hybrid rats (WELS/Fohm-
BD1X,) with doses of 0, 50 or 150 (ig/kg/day, has a broader dose range than the Guzman and Solter
(1999) study. It uses the most relevant route of exposure and has more animals/dose group (n=10) than
the Guzman and Solter (1999) study where there were only 3 per dose group. Only males were studied in
the Heinz et al (1999) study. Liver histopathology was seen in both dose groups. There was degeneration
and necrosis of hepatocytes and PAS positive staining (indicative of cell membrane damage) that
increased in severity with dose. Serum levels of LDH and ALP were significantly (p=0.05) increased
above controls for both doses. All animals displayed Kupffer cell activation in response to hepatic cell
injury. Each of these biomarkers for liver damage was increased at both doses in all ten animals. The
severity scores increased with dose but not the number of animals affected. The doses used by Heinze
(1999) in the rat study also falls between the exposures used in the Fawell et al. mice study, however
fewer animals (10) were used by Heinze (1999) in comparison to Fawell et al. (1999) (30).
While the liver is the primary target of microcystin toxicity, there have been some reports of effects in
other systems, including hematological, kidney, cardiac, reproductive, and gastrointestinal effects. It has
been suggested that some effects in other organs observed after high doses of microcystin-LR may result
from ischemia or hypoxia caused by hepatic hemorrhage. However, effects outside the liver have been
observed in the absence of hemorrhage. In most cases effect levels are at doses greater than those
impacting the liver.
Gastrointestinal effects (necrosis, duodenal damage) were observed in aged mice exposed orally to single
500 |ig/kg doses of microcystin-LR (Ito et al., 1997a). Female mice exposed subchronically to 1000 (.ig/kg
had slight increases in hemoglobin concentration, erythrocyte count and packed cell volume (Fawell et
al., 1999).
Kidney effects, including eosinophilic materials in the Bowman's spaces, were observed in two mice
exposed to lethal doses of microcystin-LR (Yoshida et al., 1997). Fawell et al. (1999) described the
appearance of the kidneys as pallid. Milutinovic et al. (2002, 2003) exposed Wistar rats to i.p. doses of
microcystin-LR and microcystin-YR (10 |ig/kg for 8 months) finding evidence for necrotic and apoptotic
damage to both the glomerulus and the tubular epithelium. Evidence of action aggregation implicated
damage to the cytoskeleton as a result of PP2A inhibition as a factor contributing to the damage. The
effects associated with microcystin-LR were greater than those caused by microcystin-YR.
Mechanistic studies by Alverca et al. (2009) in cultured Vero cells and Nobre et al. (1999) examining
perfused kidneys observed changes in cellular morphology such as enlarged lysosomes, reduction in the
number of intact lysosomes, and lysosomal leakage in the former study, and urinary flow rate in the latter
study.
Studies by Li et al. (2014, 2015) found an impact of microcystin-LR on learning the Morris water maze
and subsequent retention of the learning. In the memory phase of the trial, the male rats did not do as well
at remembering the quadrant where the platform had been located during the learning component of the
study. The 2014 study evaluated mature male rats directly exposed to microcystin-LR in solutions with
methanol. The 2015 study exposed the tested animals (pups PND7-60) through their dams. The dams
were dosed during an 8-week period prior to mating but not during gestation and lactation. The results of
both studies are confounded by the presence of methanol in the dosed solutions.
A single oral study of developmental toxicity in mice reported maternal toxicity, liver effects and deaths
in some dams treated at the highest dose of microcystin-LR (2000 |ig/kg during GD 6-15), along with
reduced fetal body weight and delayed skeletal ossification in offspring (Fawell et al., 1999). Li et al.,
Health Effects Support Document for Microcystins - June, 2015
81
-------
(2015) identified a NOAEL of 2.5 |_ig/kg/day and a LOAEL of 10 |_ig/kg/day based on maternal
gestational weight gain in Sprague Dawley rats exposed to microcystin-LR for 8 weeks prior to
conception, but not during mating and gestation. The study is confounded by the presence of methanol in
all the tested solutions, resulting in uncertainty regarding whether synergy between the microcystin-LR
and methanol could have influenced the results. There were no significant differences in the
developmental parameters in the pups, although neurological effects were observed in postnatal testing as
described above. In an i.p. study by Chernoff et al., (2002) there were maternal deaths at doses >32 j^ig/kg
but no observed effects on number of pups and pup body weight up to PND 5 for the dams that survived.
Effects observed in the male reproductive system include decreased absolute and relative testes weights;
decreased absolute and relative epididymis weight; decreased epididymal sperm concentration, decreased
sperm viability, decreased sperm motility, increased percent immobile sperm and sperm abnormalities.
Histological examination of the testes revealed atrophy of the seminiferous tubules, obstructed
seminiferous tubules, deformation of androgonial and sperm mother cells; decreased number of interstitial
cells, Sertoli cells, and mature sperm in the seminiferous tubule; lipid peroxidation; and apoptosis (Chen
et al., 2011, 2013; Li et al., 201 lb; Liu et al., 2010; Li et al., 2008; Zhang et al., 201 lb). In vitro studies of
rat spermatogonia, the precursor cells from which spermatocytes arise, demonstrate uptake of
microcystin-LR with resultant cellular apoptosis and oxidative stress (Zhou et al., 2012).
Male mice administered microcystin-LR via their drinking water for 3 or 6 months at low concentrations
(3.2g/L) had decreased sperm counts and sperm motility (Chen et al., 2011). By 6 months there was also
an increase in sperm abnormalities, decreased serum testosterone and increased serum LH levels. Testes
weights, however, were not affected. The LOAEL for these effects was 0.79 j^ig/kg with a NOAEL of
0.25 |ag/kg. The observed effects suggest a need to confirm the reported results.
Data from a number of mechanistic studies involving the male reproductive system support the need for
additional research. In vitro cell viability of Sertoli and Leydig cells was decreased by exposure to
microcystin-LR (Li et al., 2008; Zhang et al., 201 lb; Li and Han, 2012). Changes in morphology were
marked by cell shrinkage and loss of membrane integrity. Wang et al. (2012), found that microcystin-LR
was not able to enter Leydig cells reflected by the lack of Leydig cell cytotoxicity. Testosterone
production was also decreased in vitro in Leydig cells incubated with microcystin-LR (Li et al., 2008).
Male hormone levels were affected by microcystin-LR in both in vitro and in vivo studies. In vivo studies
in male mice found that microcystin-LR induced decreases in serum testosterone and increases in serum
LH and FSH (Chen et al., 2011; Li et al., 2008). Microcystin-LR also affected hormone levels in male
mice by damaging the hypothalamic-pituitary axis as measured by decreased mRNA expression for
GnRH (Wang et al., 2012; Xiong et al., 2014).
7.1.1 Mode of Action of Noncancer Effects
Mechanistic studies of microcystin cellular effects shed light on the mode of action for noncancer effects.
One important feature appears to be the need for membrane transporters for systemic uptake and tissue
distribution of microcystin by all exposure routes (Fischer et al., 2005; Feurstein et al., 2010). Members
of the OATp transporter family regulate uptake and efflux from the intestines, liver, kidney, testes, brain,
lung, heart, and placenta (Augustine et al., 2005; Cheng et al., 2005; Svoboda et al., 2011). The
importance of the transporters to tissue access is demonstrated by the data that indicate a reduction in, or
lack of, liver damage when OATp is inhibited (Hermansky et al., 1990 a,b; Thompson and Pace, 1992).
Uptake of microcystins causes protein phosphatase inhibition and loss of coordination between kinase
phosphorylation and phosphatase dephosphorylation resulting in destabilization of the cytoskeleton; this
event initiates altered cell function followed by cellular apoptosis and necrosis (Barford et al., 1998).
Together cellular kinases and phosphatases maintain the balance between phosphorylation and
Health Effects Support Document for Microcystins - June, 2015
82
-------
dephosphorylation of key cellular proteins that control metabolic processes, gene regulation, cell cycle
control, transport and secretory processes, organization of the cytoskeleton and cell adhesion.
Microcystins LR, LA and LL each interact with catalytic subunits of PP1 and PP2 inhibiting their
functions (Craig et al., 1996).
The consequences of the microcystin induced changes in cytoskeleton appear to be increases in apoptosis
and ROS. Cellular pro-apoptotic Bax and Bid proteins increased and anti-apoptotic Bcl-2 decreased in
both in vitro and in vivo studies (Fu et al., 2005; Weng et al., 2007; Xing et al., 2008; Takumi et al., 2010;
Huang et al., 2011; Li et al., 201 Id). Mitochondrial permeability transition pore, and mitochondrial
membrane potential changes (Ding and Ong, 2003; Zhou et al., 2012) led to membrane loss of
cytochrome c, a biomarker for apoptotic events. Wei et al., (2008) found that microcystin-LR induces a
time-dependent increase in ROS production and lipid peroxidation in mice. The levels of hepatic ROS
increased rapidly within 0.5 hours of receiving a 55 jj.g/kg body weight i.p. injection of microcystin-LR,
and continued to accumulate in a time-dependent manner for up to 12 hrs.
7.1.2 Dose-Response Characterization for Noncancer Effects
7.1.2.1 Human Data
There are no dose response data from the epidemiology case studies of microcystin. Acute intoxication
with microcystin-producing cyanobacteria blooms in recreational water was reported in Argentina in 2007
(Giannuzzi et al., 2011). A single person was immersed in a Microcystis blooms with a concentration of
48.6 (ig/L. After four hours of exposure, the patient showed fever, nausea, and abdominal pain and three
days later, presented dyspnea and respiratory distress and was diagnosed with an atypical pneumonia. A
week after the exposure, the patient developed a hepatotoxicosis with a significant increase of ALT, AST
and yGT. The patient was completely recovered within 20 days.
The scant human data on the oral toxicity of microcystin-LR are limited by the potential co-exposure to
other toxins and microorganisms and by the lack of quantitative information. Symptoms reported after
acute recreational exposure to cyanobacterial blooms (including microcystin-producing genera) included
headache; sore throat; vomiting and nausea; stomach pain; dry cough; diarrhea; blistering around the
mouth; and pneumonia (Turner et al., 1990). Elevated liver enzymes have been measured in humans
served by a public water supply contaminated with a bloom ofM aeruginosa (Falconer et al., 1983) and
in children consuming high levels of microcystin through water and food (Li et al., 201 la). Symptoms
occurring after exposure to cyanobacteria cannot be directly attributed to microcystin toxins or other
endotoxins; some effects may result from exposure to the cyanobacterial cells themselves, or from
exposure to multiple toxins in the bloom.
7.1.2.2 Animal Data
A major noncancer health effect of exposure to microcystin-LR in animal studies is liver damage. Oral
exposure to single 500 (ig/kg doses of microcystin-LR resulted in diffuse hemorrhage in the liver of mice
and rats; more pronounced liver damage occurred at higher doses (Ito et al., 1997a; Fawell et al., 1999).
Oral LD50 values ranged from 3000 j^ig/kg to >5000 j^ig/kg in rats and mice (Fawell et al., 1999; Yoshida
et al., 1997; Fitzgeorge et al., 1994). Studies which utilized parenteral administration of microcystin-LR
show a steep dose-response with rapid onset of liver damage.
The dose-response database for microcystins is almost exclusively limited to data on a single congener,
microcystin-LR. Data on the RR, YR, and LA do not provide useful dose-response information suitable
for quantification. With consideration of the seasonal episodic nature of algal blooms and resultant
Health Effects Support Document for Microcystins - June, 2015
83
-------
potential exposures to microcystins from public water supplies, the following studies summarized in
Table 7-1 were those selected as most suitable for derivation of guideline values.
Table 7-1. Adverse Effects By Route of Exposure to Microcystins
Dose
|jg/kg/day
Severity
Finding
Description of Effect
Study
Intraperitoneal infusion
32
++
Fibrous tissue, cell death, necrosis, lipid vacuoles, Kupffer
cell activation
Guzman and
Solter, 1999
48
+++
Fibrous tissue, cell death, necrosis, lipid vacuoles, Kupffer
cell activation
Gavage
40
10/30
Chronic inflamation1
Fawell (1994)
200
15/30
Chronic inflamation1
hepatocyte degeneration (2/30)
Drinking Water
50
++
Hepatocyte degeneration and necrosis; and PAS positive
staining, Kupffer cell activation
Heinze (1999)
150
+++
Hepatocyte degeneration and necrosis; and PAS positive
staining Kupffer cell activation
1Lesion with the best response to dose; the effect was seen in 6/30 controls, 10 at the low does, 12 at the mid dose and 29 at the
high dose
++ moderate severity
+++ high severity
7.2 Synthesis and Evaluation of Carcinogenic Effects
Several human epidemiological studies have reported a possible association between consumption of
surface waters containing cyanobacteria and microcystins that served as drinking water sources and liver
or colon cancer in certain areas of China (Ueno et al., 1996; Zhou et al., 2002). In these studies, the use of
a surface drinking water supply was used as a surrogate for exposure to microcystins. Individual exposure
to microcystins was not estimated and there was no examination of numerous possible confounding
factors such as hepatitis infection, industrial discharges and/or waste water discharges to the same surface
water sources.
Flemming et al. (2002, 2004) failed to find a significant association for primary liver cancers between
populations living in areas receiving their drinking water from a surface water treatment plant (with the
potential for microcystin exposures) and the Florida general population plus those receiving their water
from ground water sources. The strongest association observed was that between those receiving their
water from a surface water service area and those receiving their water from the surroundings of the
buffer zones with socioeconomic factors assumed to be similar for the residents of the 18 surface water
sources evaluated. The origin of the water supplies for the buffer zones was not identified. The age-
adjusted cancer rate for the surface-water area was 1.15 cases versus 0.83 cases for the buffer zone,
yielding a SRR of 1.39 cases in the surface water zone to 1 case in the buffer zone area. The buffers zone
also had a lower age adjusted cancer rate than the state of Florida (SRR = 0.59). A major weakness of this
study is the fact that it examined the cancer rate based on location of residence at the time of diagnosis
without any data on the duration of residence. Florida is known to be a state with population residence
turnover because of its appeal to retirees and winter-only residents.
Ito et al. (1997b) conducted the only longer-term oral animal study of a purified microcystin. In this
study, chronic gavage doses of 80 |ag microcystin-LR/kg/day for 80 or 100 days over 28 weeks (7
months) failed to induce neoplastic nodules of the liver in mice. Despite the study duration problem, the
Health Effects Support Document for Microcystins - June, 2015
84
-------
lack of hyperplastic nodules and limited liver damage at 7 months suggests that microcystin is not a
mutagenic initiator of tumors. The i.p. studies of microcystin-LR as an initiator found low levels of GST-
P foci when up to 10 j^ig/kg of microcystin-LR was injected in rats with liver weights that were
statistically equivalent to those of the initiated controls (Nishiwaki- Matsushima et al., 1992; Ohta et al.,
1994; Sekijima et al., 1999), but significantly increased after a partial hepatectomy stimulated tissue
repair. The lack of an increase in liver weight was used as a marker for lack of liver damage and protein
phosphatase inhibition. All three studies concluded that microcystin was a promoter of tumorigenesis
rather than an initiator of the process.
7.2.1 Mode of Action and Implications in Cancer Assessment
Protein phosphatase inhibition and its impact on the cytoskeleton increases the risk for DNA replication
errors during cell division. Microcystin-LR can promote tumorigenesis because it perpetuates existing
DNA damage in cases where a cell divides before replication errors can be repaired. Once a cell had been
damaged by microcystin-LR and the repair process has begun, ROS, spindle problems, the presence of
alkylating agents such as DEN, and other factors generate a high risk for uncontrolled cell proliferation
resulting in tumors.
Genotoxicity studies of microcystin-LR provide conflicting results. Two microcystin-containing extracts
gave positive results in the Ames assay (Ding et al., 1999; Huang et al., 2007), while negative results
were observed using M. aeruginosa extracts as well as purified microcystin (Grabow et al., 1982; Wu et
al., 2006; Repavich et al., 1990).
Positive genotoxicity results were observed in mammalian cell lines (Suzuki et al., 1998; Zhan et al.,
2004; Nong et al., 2007; Zegura et al., 2006, 2008a,b, 2011; Li et al., 201 lb) but in vivo animal studies
yielded conflicting results (Gaudin et al., 2008, 2009; Abramsson-Zetterberg et al., 2010; Zhang et al.,
201 la; Dong et al., 2008). Evidence for microcystin-LR-induced DNA damage as measured by the comet
assay has been called into question by the finding that apoptosis can lead to false positive findings in this
assay (Lankoff et al., 2004a). Some evidence exists for a clastogenic effect of microcystin-LR (Ding et
al., 1999; Zhan et al., 2004; Lankoff et al., 2006a; Repavich et al., 1990). Metabolic activation has been
found to decrease microcystin-LR mutagenicity. The inconsistent outcomes from the mutagenicity studies
may be related to differences in the cell uptake of microcystin-LR, the metabolism of microcystin-LR in
the test system, or the amount of damage to the cytoskeleton and its impact on DNA and cell replication.
DNA fragmentation was significantly increased in rat neutrophils with microcystin-LA and microcystin-
YR, but not in human neutrophils (Kujbida et al., 2008). Microcystin-YR has also been found to induce
DNA damage in the blood (lymphocytes), liver, kidney, lung, spleen, and brain of mice administered 10
|ig microcystin-YR/kg via i.p. injection every other day for 30 days (Filipic et al., 2007). Lankoff et al.
(2003) showed that microcystin-LR, through its effect on microtubules, damages the mitotic spindle,
leading to the formation of polyploid cells.
7.2.2 Weight of Evidence Evaluation for Carcinogenicity
Applying the Guidelines for Carcinogen Risk Assessment (U.S. EPA, 2005a), there is inadequate
information to assess carcinogenic potential for microcystins. The epidemiology studies are limited by
their ecological study design, poor measures of exposure, potential co-exposure to microbial and/or
chemical contaminants and, in most cases, failure to control for known liver and colorectal risk factors.
Oral exposure, dose-response data from animal studies of the carcinogen potential of microcystins are not
available. Several studies suggest that microcystin-LR is a promoter of tumors initiated by known
mutagens or during tissue damage repair. However, there are no clear data that demonstrate heritable,
structural changes to genes or chromosomes as a consequence of the direct interaction of microcystin-LR
Health Effects Support Document for Microcystins - June, 2015
85
-------
with the genome to justify classifying it as a direct carcinogen. Given microcystin-LR potential impact on
the cell cytoskeleton, liver necrosis, generation of ROS, and cell biochemistry, it is not surprising that
studies support the concept that microcystin-LR plays a secondary role in the tumorigenic process. The
work by Nishiwaki-Matsushima et al. (1992) and others, which compared liver P-GST foci from 10 (ig/L
microcystin-LR in combination with initiation, indicate that microcystin-LR can be a promoter, especially
when accompanied by tissue damage requiring repair. The results from the second part of the Nishiwaki-
Matsushima et al. (1992) study that compared P-GST foci following initiation with NDEA followed by
microcystin-LR (10 j^ig/kg) treatment, both before and after a partial hepatectomy, support this
conclusion. The number of foci/cm2 with NDEA alone was 13.4 ± 4.2 foci/cm2; with DEN, microcystin-
LR and a partial hepatectomy it was 17.4 ±3.8 foci/cm2. The results for microcystin exposure alone (10
(ig/kg before the hepatectomy and 50 j^ig/kg after) in the absence of the NDEA was 0.4 ±0.3 foci/ cm2
The International Agency for Research on Cancer (IARC) classified microcystin as a Group 2B (possibly
carcinogenic to humans) largely based on its ability to disrupt cellular architecture, along with cell
division and repair as supported by the i.p. tumor promotion data. The EPA 2005 cancer guidelines
support selecting a descriptor for an agent that has not been tested in a cancer bioassay if sufficient other
information is available to make a strong, convincing, and logical case through scientific inference. In the
case of Microcystin-LR, strong information to support classifying microcystin as a carcinogen is not
available, even though mechanistic data support a role for its contribution to the progression of tumors
initiated by other compounds.
7.2.3 Dose Response Characterization for Cancer Effects
Dose-response data regarding the carcinogenicity of microcystins from animal studies are not available.
Some studies suggest that microcystin is a promoter for tumors initiated by known mutagens or through
complication of tissue repair such as that caused by a partial hepatectomy. Given the potential impact on
the cell cytoskeleton, necrotic effects on liver cells, generation of ROS and other biochemical changes,
this is not surprising. Tissue damage requires cell division in the repair process. The cytoskeleton plays an
integral role in reparative cell division and ROS are capable of changing the altered DNA structure. Thus,
there are multiple opportunities for changes that lead to loss of control of the cell division process and
clonal explosion of the impacted cells. The work by Nishiwaki-Matsushima et al., (1992),which compares
P-GST foci from 10 |_ig/L microcystin LRto that from the phenobarbital (0.05% in the diet) as a positive
control, suggests that it is at best a weak promoter because the combination of NDEA with phenobarbital
resulted in more GSTP foci than the combination of NDEA with microcystin. The results from the second
part of the same study, which compared GST-P foci following initiation with NDEA followed by
microcystin (10 j^ig/kg) both before and after a partial hepatectomy, support this conclusion. The number
of foci/cm2 with NDEA alone was 13.4 ± 4.2 foci/cm2 and with microcystin-LR and a partial hepatectomy
the number of foci/cm2 was 17.4 ± 3.8. The impacted foci area was lower for the initiated promoted rats
(0.1 ±0.2 %) than the initiated group in the absence of microcystin-LR promotion (2.7 ±3.1%). Therefore,
although there were slightly more foci in the initiated/promoted rats, the area of the liver impacted was
smaller.
7.3 Potentially Sensitive Populations
Available animal data is not sufficient to determine if there is a difference in the response of males versus
females following oral exposure to microcystin. Fawell et al. (1999) observed a slight difference between
male and female mice in body weight and serum proteins (ALT and AST), but no sex-related differences
in liver pathology.
Studies in laboratory rodents suggest that the acute effects of microcystin-LR may be more pronounced in
adult or aged animals than in juvenile animals (Adams et al., 1985; Ito et al., 1997a; Rao et al., 2005). In
Health Effects Support Document for Microcystins - June, 2015
86
-------
these studies, young animals showed little or no effect with microcystin-LR doses found to be lethal to
adult animals. Age-dependent differences in toxicity were observed after both oral and i.p. exposure,
suggesting that differences in gastrointestinal uptake were not entirely responsible for the effect of age.
The relevance of these age-related differences to acute toxicity in humans is unknown. However, for
cyanotoxins including microcystins, drinking water contributes the highest risk of the total cyanotoxin
intake for infants to one-year old fed exclusively with powdered formula prepared with tap water
containing cyanotoxins. Based on the average drinking water intake rates for infants (<12 months; 0.15
L/kg/day), the exposure of infants is 5 times higher than those of adults (>21 yrs. old) on a body weight
basis.
Based on the available studies in animals, individuals with liver and/or kidney disease might be more
susceptible than the general population since the detoxification mechanisms in the liver are compromised
and excretory mechanisms in the kidney are impaired. Data from an episode in a dialysis clinic in
Caruaru, Brazil, where microcystins were not removed by treatment of dialysis water, identify dialysis
patients as a population of potential concern in cases where the drinking water source for the clinic is
contaminated with cyanotoxins. Other individuals of potentially sensitivity are pregnant woman, nursing
mothers, and the elderly population.
7.4 Characterization of Health Risk
7.4.1 Choice of Key Study
The critical study chosen for determining the guideline value is the short-term study by Heinze (1999) in
which rats were administered microcystin-LR via drinking water for 28 days at concentrations of 0, 50 or
150 |ig/kg body weight (Heinze, 1999). The LOAEL was determined to be 50 (.ig/kg/dav based on
increased liver weight, slight to moderate liver lesions with hemorrhages, and increased enzyme levels.
The selection of the study by Heinze (1999) was based on the appropriateness of the study duration, the
use of multiple doses, dose-related toxicological responses, and histopathological evaluations of toxicity.
After 28 days of exposure, rat organ weights (liver, kidneys, adrenals, thymus and spleen) were measured,
and hematology, serum biochemistry plus histopathology of liver and kidneys were evaluated.
The route of exposure was another important factor for the selection of this study. Although the studies
discussed above used different species and strains of laboratory animal and differed in dose, duration,
route of exposure, and description of liver histopathology, they all reported effects to the liver in the 30-
50 |ig/kg dose range. The results reported are consistent with the hypothesis that the risk for liver damage
is proportional to the exposure route as predicted because of the requirement for intestinal facilitated or
active transport (i.p. infusion > drinking water > gavage). Although the biomarkers for liver damage
differ, the results are consistent with the hypothesis that the route of exposure needs to be considered as
an important variable. The cell necrosis following a 28 day exposure reported by Heinze (1999) is
supported by the Guzman and Solter (1999) findings using slow osmotic infusion to the peritoneum. The
fact that the oral drinking water dose in Heinze (1999) caused similar signs of liver damage as did i.p.
infusion in Guzman and Solter (1999), is consistent with the conclusion that the intestinal barrier limits
flow to serum for distribution to the liver. Although the Fawell et al. (1999) study had a longer duration,
dosing occurred as a bolus to the intestine once per day, thereby limiting most absorption to the period of
small intestinal transit. It may also be that mice of the strain used by Fawell et al. (1999) are less sensitive
to microcystin-LR hepatic injury than rats. The increase in hepatic chronic inflammation in the 40
(ig/kg/day group (10/30) compared to the control group (4/30) is supportive of the conclusion that a
smaller amount of the chemical reached the liver in the Fawell et al. (1999) experiment than it did in the
drinking water and i.p infusion studies based on the manifestation of necrosis in the latter at similar doses.
Health Effects Support Document for Microcystins - June, 2015
87
-------
7.4.2 Endpoint Selection
Upon considering all available studies, liver damage was considered the most appropriate basis for
quantitation as it was a common finding among oral toxicology studies (Falconer et al., 1994; Fawell et
al., 1999; Ito et al., 1997b). While the liver is the usual target of microcystin toxicity, there have been
reports of effects of microcystin-LR on the male reproductive system and sperm development following
oral exposures (Chen et al., 2011).
Oral exposures to low concentrations of microcystin-LR for 3 to 6 months showed reproductive toxicity
including decreased sperm counts and sperm motility, as well as an increase in sperm abnormalities,
decreased serum testosterone and increased serum luteinizing hormone (LH) levels (Chen et al., 2011).
Since these effects were observed at doses lower (0.79 |_ig/kg/day) than those observed for liver effects in
Heinze (1999), EPA evaluated Chen et al. (2011) and the lesions in the testes and effects on sperm
motility as the potential critical study and points of departure for the derivation of the RfD for
microcystins.
The Chen et al. (2011) study has several limitations in the experimental design and reporting. There was
a lack of data reported on testis weights and sperm motility. The authors only reported "no significant
differences in testis weights", but no information was provided on the weights of the testis or whether
there was a trend toward decreasing weights that failed to reach significance. Also, no information was
given on the methodology used for sperm motility evaluation. No information was provided on how
samples were handled and what measurements were made to determine the percentage of sperm motility.
Although body weight and amount of water consumed were measured, these data were not presented and
doses to the animals were not calculated by the study authors. In addition, the purity of microcystin-LR
and the species and age of the mouse used were not reported. Male sperm characteristics such as volume,
motility, and structure of sperm differ developmentally by age; therefore, not knowing the age of the
mouse in the study introduces uncertainty to the quantification of the reproductive effects.
The fixation and staining of the testes used for microscopic examination (paraformaldehyde in phosphate-
buffered saline (PBS) and paraffin) in Chen et al. (2011) could result in the generation of artifacts, such as
disruption of the testicular tubes. Cytoplasmic shrinkage and chromatin aggregations were observed in
both control and experimental groups. In order to preserve the microstructure of the testis, dual fixation
such as Davidson's or Bouin's fixation followed by PAS staining should have been done. In addition, the
histopathology analysis of the testis reported by the authors did not provided sufficient detail to
adequately assess the degree of damage.
The quality of the medium used for the sperm analysis in Chen et al. (2011), and the lack of additional
data from the sperm analysis measurements carried out through the computer-assisted sperm analysis
(CASA) are additional limitations in experimental design for this study. Very few details of the serum
hormone assay protocol and the quantitative parameters of sperm motility from the CASA analysis were
provided. Therefore, the calculation for the motility of the sperm was unclear and couldn't be verified.
Based on the limitations in study design, report and methods used by Chen et al. (2011), it was concluded
that the quantitative data on decreased sperm counts and sperm motility were not appropriate for use as
the effect to determine the point of departure for the derivation of the RfD for microcystin-LR.
7.4.3 RfD Determination
The LOAEL from the Heinze (1999) study was the 50 (ig/kg/day dose based on liver effects (increased
liver weight, slight to moderate liver necrosis lesions, with or without hemorrhages at the low dose and
increased in severity at the high dose, and changes in serum enzymes indicative of liver damage). The
RfD for microcystin-LR is calculated as follows:
Health Effects Support Document for Microcystins - June, 2015
88
-------
RfD = 5°^g/kg/day = 0.05 jug/kg/day
1000
where:
50 pg/kg/day = The LOAEL for liver effects in 11-week-old male hybrid rats exposed to MC-
LR in drinking water for 28 days (Heinze, 1999).
1000 = The composite UF including 10 for intraspecies variability (UFh), 10 for
interspecies differences (UFa), 3 (1005) for LOAEL to NOAEL extrapolation
(UFl), and 3 (100 5) for uncertainties in the database (UFd).
• UFh. A Ten-fold value is applied to account for variability in the human population. No
information was available to characterize interindividual and age -related variability in the
toxicokinetics or toxicodynamics among humans.
• UFA. A Ten-fold value is applied to account for uncertainty in extrapolating from laboratory
animals to humans (i.e., interspecies variability). Information to quantitatively assess
toxicokinetic or toxicodynamic differences between animals and humans is unavailable for
microcystin. Allometric scaling is not applied in the development of the Ten-Day HA values for
microcystin. The allometric scaling approach is derived from the relationship between body
surface area and basal metabolic rate in adults (U.S. EPA, 2011). For infants and children, surface
area and basal metabolic rates are very different than adults and are not appropriate for infants
and children.
• UFl. An uncertainty factor of 3 (10°5 = 3.16, rounded to 3) to adjust the LOAEL to a NOAEL
was applied. The threefold factor is justified based on the evidence that suggests that the uptake
of microcystins by tissues requires membrane transporters. Uptake from the intestines involves
both apical and basolateral transporters, uptake by the microvilli capillaries and portal transport to
the liver. Transporters are again necessary for hepatic uptake. When there is slow infusion into
the peritoneum and into the portal intraperitoneal capillaries, uptake is described as rapid because
of the rich blood supply and large surface area of the peritoneal cavity (Klassen, 1996). Delivery
of the microcystin to the intraperitoneum increases the amount of the dose that reaches the liver
for three additional reasons: 1) the apical and basolateral intestinal barriers to uptake are
eliminated with the i.p. infusion; 2) there is no dilution of dose by the gastric plus intestinal fluids
as well as with food residues in the gastrointestinal track; and 3) there is no delay in reaching the
site of absorption because of gastric emptying time (Klassen, 1996). In addition, facilitated
transporter kinetics are similar to Michaelis Menton enzyme kinetics in that there are Km and
Vmax components that are defined by the affinity of the transported substance for the transporter.
In the Guzman and Solter (1999) intraperitoneal infusion study in rats, the NOAEL was 16
(ig/kg/day and the LOAEL was 32 |_ig/kg/day. Given the 2-fold difference between the NOAEL
and LOAEL in this study, there is no reason to believe that the less direct delivery from the
intestines to the liver expected following oral exposures through drinking water (as was used in
Heinze) would have a more than 3-fold separation between a NOAEL and LOAEL had there
been one.
• UFd. An uncertainty factor of 3 (10°5 = 3.16 rounded to 3) is selected to account for deficiencies
in the database for microcystin. The database includes limited human data, including studies
evaluating the association between microcystin exposure and cancers in liver and colon, and
systemic effects including liver endpoints such as elevated liver enzymes. Oral and i.p. acute and
short-term studies on mice and rats, and subchronic studies done in mice are available. Chronic
data are also available for microcystin, however, they are limited by the lack of quantitative data
provided in the study. Additionally, there are limited neurotoxicity studies (including a recent
publication on developmental neurotoxicity) and several i.p. reproductive and developmental
toxicity studies. The database lacks a multi-generation reproductive toxicity study.
Health Effects Support Document for Microcystins - June, 2015
89
-------
It should be noted that, the default factors typically used cover a single order of magnitude (i.e., 101). By
convention, in the Agency, a value of 3 is used in place of one-half power (i.e., 100.5) when appropriate
(U.S. EPA, 2002).
Health Effects Support Document for Microcystins - June, 2015
90
-------
8.0 RESEARCH GAPS
Microcystin-LR has the most comprehensive database among the cyanotoxins produced by cyanobacteria
and among the microcystin congeners yet much remains to be done. As anthropogenic activities and
climate change continue to stress lakes, rivers, ponds, and streams that serve as sources of drinking water,
irrigation water and sites for recreation, research to fill existing data gaps on health effects in humans,
wildlife and domestic animals becomes increasingly important. This chapter provides a summary of gaps
in knowledge identified during the development of this document. The key research gaps listed below are
not intended to be an exhaustive list. Additional research efforts are needed on:
• The absorption, distribution, and elimination of microcystins in humans and animals following
oral, inhalation or dermal exposure that can be used to support extrapolation of the oral exposure
data across species and to other exposure routes.
• The toxicity of microcystin-LR to the male reproductive system after sub-acute to chronic oral
exposure. Special attention should be given to the potential clinical significance of the decreased
sperm count and motility; reduced testosterone levels; and microscopic lesions in the testes
observed in mice by Chen et al. (2011).
• The toxicity of microcystin-LR to the female reproductive tissues and those of offspring
following oral exposure.
• The relative potencies of other microcystin congeners when compared to microcystin-LR.
• Health risks posed by repeated, low-level exposures of microcystins.
• The adverse effects of chronic exposures to microcystins.
• The immunotoxic, neurotoxic and developmental/reproductive toxicity of microcystins following
oral exposure.
• The carcinogenic potential of microcystin-LR.
• Potential health risks from exposure to mixtures of microcystins with other cyanotoxins and
chemical stressors present in ambient and or drinking water supplies.
• Populations that might be sensitive to microcystins exposure via the oral, dermal and/or
inhalation routes.
• Bioconcentration and bioaccumulation of microcystins in aquatic vertebrates and invertebrates
and the transfer in the food web.
• Bioavailability of microcystins in seafood and crops to humans consuming fish, shellfish and
edible plants that have been exposed to microcystins contaminated water.
Health Effects Support Document for Microcystins - June, 2015
91
-------
9.0 REFERENCES
Abramsson-Zetterberg, L., Sundh, U. B., and Mattsson, R. 2010. Cyanobacterial extracts and
microcystin-LR are inactive in the micronucleus assay in vivo and in vitro. Mutation Research,
699(1-2): 5-10.
Adams, W. H., Stoner, R. D., Adams, G. et al. 1985. Pathophysiologic effects of atoxic peptide from
Microcystis aeruginosa. Toxicon, 23(3): 441-447.
Adamovsky, O., Kopp, R., Hilscherova, K., et al. 2007. Microcystin kinetics (bioaccumulation and
elimination) and biochemical responses in common carp (Cyprinus carpio) and silver carp
(Hypophthalmichthys molitrix) exposed to toxic cyanobacterial blooms. Environmental
Toxicology and Chemistry, 26(12): 2687-2693.
Adhikary, S. 1996. Ecology of Freshwater and Terrestrial Cyanobacteria. Journal of Scientific &
Industrial Research, 55: 753-762.
Alverca, E., Andrade, M., Dias, E., et al. 2009. Morphological and ultrastructural effects of microcystin-
LR from Microcystis aeruginosa extract on a kidney cell line. Toxicon, 54(3): 283-294.
Augustine, L. M., Markelewicz, R. J., Boekelheide, K., et al. 2005. Xenobiotic and endobiotic transporter
mRNA expression on the blood testes barrier. Drug Metabolism and Disposition, 33(1): 182-189.
Aune, T. and Berg, K. 1986. Use of freshly prepared rat hepatocytes to study toxicity of blooms of the
blue-green algae Microcystis aeruginosa and cyanotoxin Oscillatoria agardhii. Journal of
Toxicology and Environmental Health, 19(3): 325-336.
AWWA Research Foundation. (2001). Assessment of Blue-Green Algal Toxins in Raw and Finished
Drinking Water. Final report #256. Prepared by Carmichael, W. W. AWWA Research
Foundation and American Waterworks Association. Denver, CO.
Azevedo, S. M., Carmichael, W. W., Jochimsen, E. M., et al. 2002. Human intoxication by microcystins
during renal dialysis treatment in Caruaru-Brazil. Toxicology, 181-182: 441-446.
Backer, L. C., Carmichael, W., Kirkpatrick, B., et al. 2008. Recreational exposure to low concentrations
of microcystins during an algal bloom in a small lake. Marine Drugs, 6: 389-406.
Backer, L. C., Landsberg, J. H., Miller, M., et al. 2013. Canine cyanotoxin poisonings in the United States
(1920s-2012): Review of suspected and confirmed cases from three data sources. Toxins, 5(9):
1597-1628.
Bagu, J. R., Sykes, B. D., Craig, M. M., and Holmes, C. F. B. 1997. A molecular basis for different
interactions of marine toxins with protein phosphatase-1. Molecular models for bound motuporin,
microcystins, okadaic acid, and calyculin A. Journal of Biological Chemistry, 272(8): 5087-5097.
Barford, D., Das, A., K., and Egloff, M. 1998. The structure and mechanism of protein phosphatases:
Insights into catalysis and regulation. Annual Review of Biophysics and Biomolecular Structure,
27:133-164.
Batista, T., de Sousa, G., Suput, J. S., et al. 2003. Microcystin-LR causes the collapse of actin filaments in
primary human hepatocytes. Aquatic Toxicology, 65(1): 85-91.
Health Effects Support Document for Microcystins - June, 2015
92
-------
Beattie, K. A, Kaya, K., Sano, T., and Codd, G. A. 1998. Three dehydrobutyrine (Dhb)-containing
microcystins from the cyanobacterium Nostoc sp. Phytochemistry, 47(7): 1289-1292. (Cited in
WHO 1999)
Becchetti, A., Malik, B., Yue, G., et al. 2002. Phosphatase inhibitors increase the open probability of
ENaC in A6 cells. American Journal of Physiology — Renal Physiology, 283(5): F1030-F1045.
Benson, J. M., Hutt, J. A.,Rein, K. et al. 2005. The toxicity of microcystin LR in mice following 7 days of
inhalation exposure. Toxicon, 45(6): 691-698.
Berg, K., Wyman, J, Carmichael, W. W., and Dabholkar, A. S. 1988. Isolated rat liver perfusion studies
with cyclic heptapeptide toxins of Microcystis and Oscillatoria (freshwater cyanobacteria).
Toxicon, 26(9): 827-837.
Bernstein, J.A., Ghosh, D., Levin, L.S., Zheng, S., Carmichael, W., Lummus, Z., Bernstein, I.L. 2011.
Cyanobacteria: An unrecognized ubiquitous sensitizing allergen? Allergy and Asthma
Proceedings 32, 106-110.
Berry, J., Lee, E., Walton, K., et al. 2011. Bioaccumulation of microcystins by fish associated with a
persistent Cyanobacterial bloom in Lago de Patzcuaro (Michoacan, Mexico). Environmental
Toxicology and Chemistry, 30(7): 1621-1628.
Berry, J., Jaja-Chimedza, A.,Davalos-Lind, L., and Lind, O. 2012. Apparent bioaccumulation of
Cylindrospermopsin and paralytic shellfish toxins by finfish in Lake Catemaco (Veracruz,
Mexico). Food Additives and Contaminants, 29(2): 314-321.
Beussink, A. M., and Graham, J. L. 2011. Relations Between Hydrology, Water Quality, and Taste-and-
Odor Causing Organisms and Compounds in Lake Houston, Texas, April 2006-September 2008:
U.S. Geological Survey Scientific Investigations Report 2011-5121, pp. 27.
Bhattacharya, R., Rao, P.V.L., Bhaskar, A. S. B., et al. 1996. Liver slice culture for assessing hepatoxicity
of freshwater cyanobacteria. Human & Experimental Toxicology, 15(2): 105-110.
Billam, M., Mukhi, S., Tang, L., et al. 2008. Toxic response indicators of microcystin-LR in F344 rats
following a single-dose treatment. Toxicon, 51(6): 1068-1080.
Birungi, G. and Li, S. F. Y. 2011. Investigation of the effect of exposure to non-cytotoxic amounts of
microcystins. Metabolomics, 7(4): 485-499.
Blankson, H., Grotterod, E. M., and Seglen, P. O. 2000. Prevention of toxin-induced cytoskeletal
disruption and apoptotic liver cell death by the grapefruit flavonoid, narigin. Cell Death &
Differentiation, 7(8): 739-746.
Boe, R., Gjersten, B. T., Vintermyr, O. K., et al. 1991. The protein phosphatase inhibitor okadaic acid
induces morphological changes typical of apoptosis in mammalian cells. Experimental Cell
Research, 195(1): 237-246.
Botha, N, van de Venter, M., Downing, T. G., et al. 2004. The effect of i.p.ly administered microcystin-
LR on the gastrointestinal tract of Balb/c mice. Toxicon, 43(3): 251-254.
Bouai'cha, N. and Maatouk, I. 2004. Microcystin-LR and nodularin induce intracellular glutathione
alteration, reactive oxygen species production and lipid peroxidation in primary cultured rat
hepatocytes. Toxicology Letters, 148(1-2): 53-63.
Health Effects Support Document for Microcystins - June, 2015
93
-------
Bouai'cha, N., Maatouk, I., Plessis, M. J., and Perin, F. 2005. Genotoxic potential of microcystin-LR and
nodularin in vitro in primary cultured rat hepatocytes and in vivo in rat liver. Environmental
Toxicology, 20(3): 341-347.
Boyer, G. L. 2007. Cyanobacterial toxins in New York and the Lower Great Lakes ecosystems. In: H. K.
(Ed.), "Proceedings of the Interagency, International Symposium on Cyanobacterial Harmful
Algal Blooms", Advances in Experimental Medicine and Biology, pp. 151-163.
Brooks, W. P. and Codd, G. A. 1987. Distribution of Microcystis aeruginosa peptide toxin and
interactions with hepatic microsomes in mice. Pharmacology & Toxicology, 60(3): 187-191.
(Cited in WHO 1999)
Brooks, P.C., Stromblad, S., Sanders, L. C., et al. 1996. Localization of matrix metalloproteinase MMP-2
to the surface of invasive cells by interaction in integrin av(33. Cell, 85:683-693.
Bu, Y. Z., Li, X. Y., Zhang, B. J., Chung, I. K., & Lee, J. A. (2006). Microcystins cause embryonic
toxicity in mice. Toxicon, 48(8), 966-972.
Buratti, F. M., Scardala, S., Funari, E., and Testai, E. 2011. Human glutathione transferases catalyzing the
conjugation of the hepatoxin microcystin-LR. Chemical Research in Toxicology, 24(6): 926-933.
Burns, J. 2000. Cyanobacterial Blooms in Florida's Drinking Water Supplies. 20th Annual Meeting of
the Florida Chapter of the American Fisheries Society, March 28-30, 2000, Brooksville, FL.
Abstract
Burns, J. 2008. Toxic cyanobacteria in Florida waters. In: Hudnell, H.K. (Ed.), Cyanobacterial Harmful
Algal Blooms: State of the Science and Research Needs. Advances in Experimental Medicine and
Biology, 619, Chapter 5. Springer Press, New York, pp. 139-152.
Campos, A. and Vascondelos, V. 2010. Molecular mechanisms of microcystin toxicity in animal cells.
International Journal of Molecular Sciences, 11: 268-287.
Caraco, N. F. and Miller, R. 1998. Effects of CO2 on competition between a cyanobacterium and
eukaryotic phytoplankton. Canadian Journal of Fisheries and Aquatic Sciences, 55: 54-62.
Carbis, C. R., Rawlin, G. T., Grant, P., et al. 1997. A study of feral carp Cyprinus carpio L., exposed to
Microcystis aeruginosa at Lake Mokoan, Australia, and possible implication on fish health.
Journal of Fish Diseases, 20: 81-91 (Cited in WHO 1999).
Carey, C. C., Ibelings, B. W., Hoffmann, E. P. et al. 2012. Eco-physiological adaptations that favour
freshwater cyanobacteria in a changing climate. Water Research, 46: 1394-1407.
Carmichael, W. W. 1992. A Status Report on Planktonic Cyanobacteria (Blue Green Algae) and their
Toxins. EPA/600/R-92/079, Environmental Monitoring Systems Laboratory, Office of Research
and Development, U.S. Environmental Protection Agency, Cincinnati, OH. (Cited in WHO 1999)
Carmichael, W. W., Beasley, V., Bunner, D.L. et al. 1988. Naming of cyclic hepatapeptide toxins of
cyanobacteria (blue-green algae). Toxicon, 26: 971-973.
Carmichael, W. W., Drapeau, C., and Anderson, D. M. 2000. Harvesting and Quality Control of
Aphanizomenon flos-aquae from Klamath Lake for Human Dietary Use. Journal of Applied
Phycology, 12: 585-595.
Health Effects Support Document for Microcystins - June, 2015
94
-------
Carmichael, W. W., Azevedo, S. M. F. O., An, J. S., et al. 2001. Human fatalities from cyanobacteria:
Chemical and biological evidence for cyanotoxins. Environmental Health Perspectives, 109(7):
663-668.
Carmichael, W. W. and Stukenberg, M. C. 2006. Blue-green algae (Cyanobacteria). In: Coates, P.M.,
Blackman, M.R., Cragg, G.M., et al. (Eds), Encyclopedia of Dietary Supplements, 2nd Edition.
Marcel Dekker, Inc. (a div. of) Taylor and Francis Books, New York, NY. ISBN# 0-8247-5504-9
Carriere, A., Prevost, M., Zamyadi, A., et al. 2010. Vulnerability of Quebec drinking-water treatment
plants to cyanotoxins in a climate change context. Journal of Water and Health, 8(3): 455-465.
Carvalho, G. M., Oliveira, V. R., Soares, R. M., et al. 2010. Can LASSBio 596 and dexamethasone treat
acute lung and liver inflammation induced by microcystin-LR? Toxicon, 56(4): 604-612.
Casquilho, N. V., Carvalho, G. M., Alves, J. L., et al. 2011. LASSBio 596 per os avoids pulmonary and
hepatic inflammation induced by microcystin-LR. Toxicon, 58(2): 195-201.
Castenholz, R. W. 1973. Ecology of blue-green algae in hot springs. In: N.G. Carr and B.A. Whitton
(Eds.), The Biology of Blue-Green Algae. Blackwell Scientific Publications, Oxford, pp. 379-414.
(Cited in WHO 1999)
Castenholz, R. W. and Waterbury, J. B. 1989. Cyanobacteria. In: J.T. Staley, M.P. Bryant, N. Pfennig
and J.G. Holt Eds. Bergey's Manual of Systematic Bacteriology. Vol. 3, Williams & Wilkins,
Baltimore, 1710-1727. (Cited in WHO 1999)
Cazenave, J., Wunderlin, D.A., Bistoni, M. A., et al. 2005. Uptake, tissue distribution and accumulation
of microcystin-RR in Corydoras paleatus, Jenynsia multidentata and Odontesthes bonariensis in
a field and laboratory study. Aquatic Toxicology, 75:178-190.
Chemical Book. 2012. CAS Index. Retrieved September 25, 2012 from the World Wide Web:
http: //www .chemicalbook. com/Search_EN. aspx?key word=
Chen, K., Shen, Y. Z. and Shen, G. F. 1994. Study on incidence rate of some cancer in areas with
difference in drinking water sources. Chinese Journal of Public Health, 12(3): 146-148 (As cited
in Zhou et al., 2002). (Chinese)
Chen, J., Song, L., Dai, J., et al. 2004a. Effects of microcystins on the growth and the activity of
superoxide dismutase and peroxidase of rape (Brassica rapus L.) and rice (Oryza sativa L.).
Toxicon, 43: 393-400.
Chen, T., Zhao, X., Liu, Y., et al. 2004b. Analysis of immunomodulating nitric oxide, iNOS and
cytokines MRNA in mouse macrophages induced by microcystin-LR. Toxicology, 197(1): 67-77.
Chen, T., Shen, P., Zhang, J., et al. 2005. Effects of microcystin-LR on patterns of iNOS and cytokine
mRNA expression in macrophages in vitro. Environmental Toxicology, 20(1): 85-91.
Chen, J., Zhang, D., Xie, P., et al. 2009. Simultaneous determination of microcystin contamination in
various vertebrates (fish, turtle, duck and water bird) from a large eutrophic Chinese lake, Lake
Taihu, with toxic Microcystis blooms. Science of the Total Environment, 407: 3317-3322.
Chen, Y., Xu, J., Li, Y., and Han, X. 2011. Decline of sperm quality and testicular function in male mice
during chronic low-dose exposure to microcystin-LR. Reproductive Toxicology, 31: 551-557.
Health Effects Support Document for Microcystins - June, 2015
95
-------
Chen, L., Zhang, X., Zhou, W., et al. 2013. The interactive effects of cytoskeleton disruption and
mitochondria dysfunction lead to reproductive toxicity induced by microcystin-LR. PLoS One, 8:
e53949.
Cheng, X., Maher, J., Dieter, M. Z., and Klassen, C. D. 2005. Regulation of mouse organic anion-
transporting polypeptides (OATPs) in liver by prototypical microsomal enzyme inducers that
activate transcription factor pathways. Drug Metabolism and Disposition, 33: 1276-1282.
Chernoff, N., Hunter, E.S. Ill, Hall, et al. 2002. Lack of teratogenicity of microcystin-LR in the mouse
and toad. Journal of Applied Toxicology, 22(1): 13-17.
Chong, M. W. K., Gu, K. D., Lam, P. K. S., et al. 2000. Study of the cytotoxicity of microcystin-LR on
cultured cells. Chemosphere, 41(1-2): 143-147.
Chorus, I., Falconer, I., Salas, H.J., and Bartram, J. 2000. Health risks caused by freshwater cyanobacteria
in recreational waters. Journal of Toxicology and Environmental Health, Part B, Critical
Reviews. 4: 323-347.
Chow, C., Drikas, M., and Ho, J. 1999. The impact of conventional water treatment processes on cells of
the cyanobacterium Microcystis aeruginosa. Water Research, 33(15): 3253-3262.
Christensen, V .G., Maki R. P., and Kiesling, R. L. 2011. Relation of Nutrient Concentrations, Nutrient
Loading, and Algal Production to Changes in Water Levels in Kabetogama Lake, Voyageurs
National Park, Northern Minnesota, 2008-09: U.S. Geological Survey Scientific Investigations
Report 2011-5096, pp. 50.
Christoffersen, K., Lyck, S., and Winding, A. 2002. Microbial activity and bacterial community structure
during degradation of microcystins, Aquatic Microbial Ecology, v27(2): 125-136
Clark, S. P., Davis, M. A., Ryan, T. P., et al. 2007. Hepatic gene expression changes in mice associated
with prolonged sublethal microcystin exposure. Toxicologic Pathology, 35(4): 594-605.
Clark, S. P., Ryan, T. P., Searfoss, G. H., et al. 2008. Chronic microcystin exposure induces hepatocyte
proliferation with increased expression of mitotic and cyclin-associated genes in P53-deficient
mice. Toxicologic Pathology, 36(2): 190-203.
Codd, G. 1995. Cyanobacterial Toxins: Occurrence, Properties and Biological Significance. Water
Science and Technology, 32(4): 149-156.
Codd G. A. and Poon, G. K. 1988. Cyanobacterial toxins. Proceedings of the Phytochemical Society of
Europe, 28: 283-296.
Codd, G. A., Metcalf, J. S., and Beattie, K. A. 1999. Retention of Microcystis aeruginosa and microcystin
by salad lettuce (Lactuca sativa) after spray irrigation with water containing cyanobacteria.
Toxicon,31\ 1181-1185.
Codd, G. A., Morrison, L. F., and Metcalf, J. S. 2005. Cyanobacterial toxins: risk management for health
protection. Toxicology and Applied Pharmacology, 203:264-272.
Corbel, S., Mougin, C., and Bouai'cha, N. 2014. Cyanobacterial toxins: Modes of actions, fate in aquatic
and soil ecosystems, phytotoxicity and bioaccumulation in agricultural crops. Chemosphere, 96:
1-15.
Health Effects Support Document for Microcystins - June, 2015
96
-------
Cote, L-M., Lovell, R. A., Jeffrey, E. H., et al. 1986. Failure of blue-green algae (Microcystis aeruginosa)
hepatotoxin to alter in vitro mouse liver enzymatic activity. Journal of Toxicology - Toxin
Reviews, 52(2): 256.
Cousins, I. T., Bealing, D. J., James, H. A., and Sutton, A. 1996. Biodegradation of microcystin-LR by
indigenous mixed bacterial populations. Water Research, 30: 481-485. (Cited in WHO 1999)
Craig, M., Luu, H. A., McCready, T. L., et al. 1996. Molecular mechanisms underlying the interaction of
motuporin and microcystins with type-1 and type-2A protein phosphatases. Biochemistry and
Cell Biology, 74(4): 569-578.
Creasia, D. A. 1990. Acute inhalation toxicity of microcystin-LR with mice. Toxicon, 28(6): 605.
Dahlem, A. M., Hassan, A. S., Swanson, S. P., et al. 1989. A model system for studying the
bioavailability of intestinally administered microcystin-LR, a hepatotoxic peptide from the
cyanobacterium Microcystis aeruginosa. Pharmacology & Toxicology, 64(2): 177-181.
de la Cruz, A., Antoniou, M., Hiskia, A., et al. 2011. Can We Effectively Degrade Microcystins? -
Implications on Human Health. Anti-Cancer Agents in Medicinal Chemistry, 11: 19-37.
De Senerpont Domis, L., Mooij, W. M., and Huisman, J. 2007. Climate-induced shifts in an experimental
phytoplankton community: a mechanistic approach. Hydrobiologia, 584: 403-413.
Deblois, C. P., Giani, A., and Bird, D. F. 2011. Experimental model of microcystin accumulation in the
liver of Oreochromis niloticus exposed subchronically to a toxic bloom of Microcystis sp.
Aquatic Toxicology, 103: 63-70.
Dias, E., Andrade, M., Alverca, E., et al. 2009. Comparative study of the cytotoxic effect of microcystin-
LR and purified extracts from Microcystis aeruginosa on a kidney cell line. Toxicon, 53(5): 487-
495.
Dias, E., Matos, P., Pereira, P., et al. 2010. Microcystin-LR activates the ERK1/2 kinases and stimulates
the proliferation of the monkey kidney-derived cell line Vero-E6. Toxicology in Vitro, 24(6):
1689-1695.
Dietrich, D. and Hoeger, S. 2005. Guidance values for microcystins in water and cyanobacterial
supplement products (blue green algal supplements): a reasonable or misguided approach?
Toxicology and Applied Pharmacology, 203: 273-289.
Ding, W. X., Shen, H. M., Shen, Y., et al. 1998a. Microcystic cyanobacteria causes mitochondrial-
membrane potential alteration and reactive oxygen species formation in primary cultured rat
hepatocytes. Environmental Health Perspectives, 106(7): 409-413.
Ding, W. X., Shen, H. M., Zhu, H. G., and Ong, C. N. Ong. 1998b. Studies on oxidative damage induced
by cyanobacteria extract in primary cultured rat hepatocytes. Environmental Research, 78(1): 12-
18.
Ding, W. X., Shen, H. M., Zhu, B. L., et al. 1999. Genotoxicity of microcystic cyanobacterial extract of a
water source in China. Mutation Research, 442(2): 69-77.
Ding, W. X., Shen, H. M., and Ong, C. N. 2000a. Microcystic cyanobacteria extract induces cytoskeletal
disruption and intracellular glutathione alteration in hepatocytes. Environmental Health
Perspectives, 108(7): 605-609.
Health Effects Support Document for Microcystins - June, 2015
97
-------
Ding, W. X., Shen, H. M., and Ong, C. N. 2000b. Critical role of reactive oxygen species and
mitochondrial permeability transition in microcystin-induced rapid apoptosis in rat hepatocytes.
Hepatology, 32(3): 547-555.
Ding, W. X., Shen, H. M. and Ong, C. N. 2001. Critical role of reactive oxygen species formation in
microcystin-induced cytoskeleton disruption in primary cultured hepatocytes. Journal of
Toxicology and Environmental Health, Part A, 64(6): 507-519.
Ding, W. X., Shen, H. M. and Ong, C. N. 2002. Calpain activation after mitochondrial permeability
transition in microcystin-induced cell death in rat hepatocytes. Biochemical and Biophysical
Research Communications, 291(2): 321-331.
Ding, W. X. and Ong, C. N. 2003. Role of oxidative stress and mitochondrial changes in cyanobacteria-
induced apoptosis and hepatotoxicity. FEMS Microbiology Letters, 220(1): 1-7.
Ding, X., Li, X., Duan, H. Y., et al. 2006 Toxic effects of Microcystis cell extracts on the reproductive
system of male mice. Toxicon, 48(8): 973-979.
Dittman, E., Neilan, B. A., Erhard, M., et al. 1997. Insertional mutagenesis of a peptide synthetase gene
which is responsible for hepatotoxin production in the cyanobacterium. Microcystis aeruginosa
PCC 7806. Molecular Microbiology, 26: 779-787. (Cited in WHO 1999)
Dong, L., Zhang, H., Duan, L., et al. 2008. Genotoxicity of testicle cell of mice induced by microcystin-
LR. Life Science Journal, 5(1): 43-45.
Dor, I. and Danin, A. 1996. Cyanobacterial desert crusts in the Dead Sea Valley. Algological Studies, 83:
197-206. (Cited in WHO 1999).
Douglas, G. C., Thirkill, T. L., Kumar, P., et al. 2014. Effect of microcystin-LR on human placental
villous trophoblast differentiation in vitro. Environmental Toxicology.
Downing, J. A., Watson, S. B., and McCauley, E. 2001. Predicting Cyanobacteria dominance in lakes.
Canadian Journal of Fisheries and Aquatic Sciences, 58(10): 1905-1908.
Drake, J. L., Carpenter, E. J., Cousins, M., et al. 2010. Effects of light and nutrients on seasonal
phytoplankton succession in a temperate eutrophic coastal lagoon. Hydrobiologia, 654: 177-192.
Duy, T. N., Lam, P. K. S., Shaw, G. R., and Connell, D. W. 2000. Toxicology and risk assessment of
freshwater cyanobacterial (blue-green algal) toxins in water. Reviews of Environmental
Contamination and Toxicology, 163: 113-186.
Dyble, J., Gossiaux, D., Landrum, P., et al. 2011. A kinetic study of accumulation and elimination of
microcystin-LR in yellow perch (Perca flavescens) tissue and implications for human fish
consumption. Marine drugs, 9(12): 2553-2571.
Dziallas, C. and Grossart, H. 2011. Increasing oxygen radicals and water temperature select for toxic
Microcystis sp. PLoS One, 6 (9): 255-69.
Edwards C., Graham, D., Fowler, N., and Lawton, L. A. 2008. Biodegradation of microcystins and
nodularin in freshwaters. Chemosphere, 73(8): 1315-1321
Elliott, J. A. 2010. The seasonal sensitivity of cyanobacteria and other phytoplankton to changes in
flushing rate and water temperature. Global Change Biology, 16: 864-876.
Health Effects Support Document for Microcystins - June, 2015
98
-------
Elser, J. J., Bracken, M. E. S., and Cleland, E. E. 2007. Global analysis of nitrogen and phosphorus
limitation of primary producers in freshwater, marine and terrestrial ecosystems. Ecology Letters,
10: 1124-1134.
Eriksson, J. E. and Golman, R. D. 1993. Protein phosphatase inhibitors alter cytoskeletal structure and
cellular morphology. Advances in Protein Phosphatases, 7: 335-357.
Eriksson, J. E., Paater, G. I. L., Meriluoto, J. A. O., et al. 1989. Rapid microfilament reorganization
induced in isolated rat hepatocytes by microcystin-LR, a cyclic peptide toxin. Experimental Cell
Research, 185(1): 86-100.
Eriksson, J. E., Gronberg, L., Nygard, S., et al. 1990. Hepatocellular uptake of 3H-dihydromicrocystin-
LR, a cyclic peptide toxin. Biochimica et Biophysica Acta, 1025(1): 60-66.
Eriksson, J. E., Brautigan, D. L., Vallee, R. D., et al. 1992a. Cytoskeletal integrity in interphase cells
requires protein phosphatase activity. Proceedings of the National Academy of Sciences, 89:
11093-11097.
Eriksson, J. E., Opal, P., and Goldman, R. D. 1992b. Intermediate filament dynamics. Current Opinion in
Cell Biology, 4: 99-104.
Ettoumi, A., Khalloufi, F. E., Ghazali, I. E., et al. 2011. Bioaccumulation of cyanobacterial toxins in
aquatic organisms and its consequences for public health. In: G. Kattel (Ed.), Zooplankton and
Phytoplankton: Types, Characteristics and Ecology. Nova Science Publishers, New York, NY,
pp. 1-34.
Falconer, I. R. 2005. Cyanobacterial Toxins of Drinking Water Supplies: Cylindrospermopsins and
Microcystins. CRC Press Boca Raton, FL, pp. 263.
Falconer, I. R. 1998. Algal toxins and human health. In: J. Hubec (Ed.), Handbook of Environmental
Chemistry, Vol. 5, Part C, Quality and Treatment of Drinking Water, pp. 53-82.
Falconer, I. R. and Buckley, T. H. 1989. Tumour promotion by Microcystis sp., a blue-green alga
occurring in water supplies. Medical Journal of Australia, 150(6): 351.
Falconer, I. R. and Humpage, A. R. 1996. Tumour promotion by cyanobacterial toxins. Phycologia, 35:
74-79.
Falconer, I. R. and Yeung, S. K. 1992. Cytoskeletal changes in hepatocytes induced by Microcystis toxins
and their relation to hyperphosphorylation of cell proteins. Chemico-Biological Interactions,
81(1-2): 181-196.
Falconer, I. R., Beresford, A. M., and Runnegar, M. T. C. 1983. Evidence of liver damage by toxin from a
bloom of the blue-green alga, Microcystis aeruginosa. Medical Journal of Australia, 1(11): 511-
514.
Falconer, I. R., Buckley, T., and Runnegar, M. T. C. 1986. Biological half-life organ distribution and
excretion of 125I-labeled toxic peptide from the blue-green digs.Microcystis aeruginosa.
Australian Journal of Biological Sciences, 39(1): 17-21.
Falconer, I. R., Smith, J. V., Jackson, A. R. B., et al. 1988. Oral toxicity of a bloom of the cyanobacterium
Microcystis aeruginosa administered to mice over periods of up to one year. Journal of
Toxicology and Environmental Health, 24(3): 291-305.
Health Effects Support Document for Microcystins - June, 2015
99
-------
Falconer, I. R., Choice, A., and Hosja, W. 1992. Toxicity of edible mussels (Mytilus edulis) growing
naturally in an estuary during a water bloom of the blue-green alga Nodularia spumigena.
Environmental Toxicology and Water Quality, 7:119-124.
Fardilha, M., Ferreira, M., Pelech, S., et al. 2013. "Omics" of human sperm: profiling protein
phosphatases. Omics, 17: 460-472.
Fawell, J. K., Mitchell, R. E., Everett, D. J., and Hill, R. E. 1999. The toxicity of cyanobacterial toxins in
the mouse. 1. Microcystin-LR. Human & Experimental Toxicology, 18(3): 162-167.
Fay, P. 1965. Heterotrophy and nitrogen fixation in Chlorogloea fritschii. Journal of General
Microbiology, 39:11-20. (Cited in WHO 1999)
Feitz, A. J., Lukondeh, T., Moffitt, M. C., et al. 2002. Absence of detectable levels of cyanobacterial toxin
(microcystin-LR) carry-over into milk. Toxicon, 40: 1173-1180.
Feng, G., Abdalla, M, li, Y., and Bai, Y. 2011. NF-kappaB mediates the induction of Fas receptor and Fas
ligand by microcystin-LR in HepG2 cells. Molecular and Cellular Biochemistry, 352(1-2): 209-
219.
Ferrao-Filho, A. S. and Kozlowsky-Suzuki, B. 2011. Cyanotoxins: bioaccumulation and effects on
aquatic animals. Marine Drugs, 9: 2729-2772.
Feurstein, D., Kleinteich, J., Heussner, A. H., et al. 2010. Investigation of Microcystin Congener-
Dependent Uptake into Primary Murine Neurons. Environmental Health Perspectives, 118(10):
1370-1375.
Feurstein, D., Stemmer, K., Kleinteich, J., et al. 2011. Microcystin Congener- and Concentration-
Dependent Induction of Murine Neuron Apoptosis and Neurite Degeneration. Toxicological
Sciences, 124(2): 424-431.
Filipic, M., Zegura, B., Sedmak, B., et al. 2007. Subchronic exposure of rats to sublethal dose of
microcystin-YR induces DNA damage in multiple organs. Radiology and Oncology, 41(1): 15-
22.
Fischer, W. J., Altheimer, S., Cattori, V., et al. 2005. Organic anion transporting polypeptides expressed
in liver and brain mediate uptake of microcystin. Toxicology and Applied Pharmacology, 203:
257-263.
Fischer, A., Hoeger, S. A., Stemmer, K., et al. 2010. The role of organic anion transporting polypeptides
(OATPs/.S'Z/ YA ) in the toxicity of different microcystin congeners in vitro: A comparison of
primary human hepatocytes and OATP-transfected HEK293 cells. Toxicology and Applied
Pharmacology, 245(1): 9-20.
Fitzgeorge, N.L.M., S.A. Clark and C.W. Kelvin. 1994. Routes of intoxication. In: G. A. Codd, T. M.
Jeffreies, C. W. Kelvin and E. Potter, (Eds.), Detection Methods for Cyanobacterial (Blue-Green
Algal) Toxins and First International Symposium on Detection Methods for Cyanobacterial
(Blue-Green Algal) Toxins. Royal Society of Chemistry, Cambridge, U.K. pp. 69-74. (As cited in
Kuiper-Goodman et al., 1999 and WHO 1999)
Fladmark, K. E., Serres, M. H., Larsen, N. L., et al. 1998. Sensitive detection of apoptogenic toxins in
suspension cultures of rat and salmon hepatocytes. Toxicon, 36(8): 1101-1114.
Health Effects Support Document for Microcystins - June, 2015
100
-------
Fleming, L. E., Rivero, C., Burns, J., et al. 2002. Blue green algal (cyanobacterial) toxins, surface
drinking water and liver cancer in Florida. Harmful Algae. 1(2): 157-168.
Fleming, L. E., Rivero, C, Burns, J., et al. 2004. Cyanobacteria exposure, drinking water and colorectal
cancer. In: K. A. Steidinger, J. H. Landsberg, C. R. Tomas and G. A. Vargo, (Eds.), Harmful
Algae 2002. Proceedings of the Xth International Conference on Harmful Algae. Florida Fish and
Wildlife Conservation Commission and Intergovernmental Oceanographic Commission of
UNESCO, Tallahassee, FL. pp. 470-472.
Fu, W. Y., Chen, J. P., Wang, X. M., and Xu, L. H. 2005. Altered expression of p53, Bcl-2 and Bax
induced by microcystin-LR in vivo and in vitro. Toxicon, 46(2): 171-177.
Fu, W., Yu, Y., and Xu., L. 2009. Identification of temporal differentially expressed protein responses to
microcystin in human amniotic epithelial cells. Chemical Research in Toxicology, 22(1): 41-51.
Funari, E. and Testai, E. 2008. Human health risk assessment related to cyanotoxins exposure. Critical
Reviews in Toxicology, 38: 97-125
Gacsi, M., Antal, O., Vasas, G., et al. 2009. Comparative study of cyanotoxins affecting cytoskeletal and
chromatin structures in CHO-K1 cells. Toxicology in Vitro, 23(4): 710-718.
Gan, N., Sun, X., Song, L. 2010. Activation ofNrf2 by microcystin-LR provides advantages for liver
cancer cell growth. Chemical Research in Toxicology, 23(9): 1477-1484.
Gaudin, J., Huet, S., Jarry, G., and Fessard, V. 2008. In vivo DNA damage induced by the cyanotoxin
microcystin-LR: comparison of intra-paritoneal and oral administrations by use of the comet
assay. Mutation Research, 652: 65-71.
Gaudin, J., Le Hegarat, L., Nesslay, F., et al. 2009. In vivo genotoxic potential of microcystin-LR: a
cyanobacterial toxin, investigated both by the unscheduled DNA synthesis (UDS) and the comet
assays after intravenous administration. Environmental Toxicology, 24: 200-209.
Geh, E. N., Ghosh, D., McKell, M., de la Cruz, A. A., Stelma, G., & Bernstein, J. A. 2015. Identification
of Microcystis aeruginosa Peptides Responsible for Allergic Sensitization and Characterization of
Functional Interactions between Cyanobacterial Toxins and Immunogenic Peptides.
Environmental Health Perspectives. DOI: 10.1289/ehp. 1409065
Gehringer, M. M. 2004. Microcystin-LR and okadaic acid-induced cellular effects: A dualistic response.
FEBSLetters, 557(1-3): 1-8.
Gehringer, M. M., Govender, S., Shaw, M. and Downing, T. G. 2003a. An investigation of the role of
vitamin E in the protection of mice against microcystin toxicity. Environmental Toxicology,
18(2): 142-148.
Gehringer, M. M., Downs, K. S., Downing, T. G., et al. 2003b. An investigation into the effects of
selenium supplementation on microcystin hepatotoxicity. Toxicon, 41(4): 451-458.
Gehringer, M. M., Shephard, E. G., Downing, T. G., et al. 2004. An investigation into the detoxification
of microcystin-LR by the glutathione pathway in Balb/c mice. The International Journal of
Biochemistry & Cell Biology, 36(5): 931-941.
Gewolb, J. 2002. Working Outside the Protein-Synthesis Rules. Science, 295: 2205-2206.
Health Effects Support Document for Microcystins - June, 2015
101
-------
Giannuzzi, L., Sedan, D., Echenique R., and Andrinolo, D. 2011. An Acute Case of Intoxication with
Cyanobacteria and Cyanotoxins in Recreational Water in Salto Grande Dam, Argentina. Marine
Drugs, 9: 2164-2175
Gobler, C., Davis, T., Coyne, K., and Boyer, G. 2007. Interactive influences of nutrient loading,
zooplankton grazing, and microcystin synthetase gene expression on cyanobacterial bloom
dynamics in a eutrophic New York lake. Harmful Algae, 6: 119-133
Goldberg, J., Huang, H. B., Kwon, Y. G., et al. 1995. Three-dimensional structure of the catalytic subunit
of protein serine/threonine phosphatase-1. Nature, 376(6543): 745-753.
Grabow, W. O. K., Du Randt, W. C., Prozensky, O. W., and Scott, W. E. 1982.Microcystis aeruginosa
toxin: Cell culture toxicity, hemolysis, and mutagenicity assays. Applied and Environmental
Microbiology, 43(6): 1425-1433.
Graham, J., Loftin, K., Meyer, M., and Ziegler, A. 2010. Cyanotoxin mixtures and taste-and-odor-
compounds in cyanobacterial blooms from the midwestern United States. Environmental Science
and Technology, 44: 7361-7368.
Graham, J. Ziegler, L., Loving, A. C., et al. 2012. Fate and Transport of Cyanobacteria and Associated
Toxins and Taste-and-Odor Compounds from Upstream Reservoir Releases in the Kansas River,
Kansas, September and October 2011. U.S. Geological Survey Scientific Investigations Report
2012-5129, pp. 65.
Gudasz, C., Bastviken, D., Steger, K., et al. 2010. Temperature controlled organic carbon mineralization
in lake sediments. Nature, 466: 478-481.
Gulledge, B. M., Aggen, J .B., Huang, H. B., et al. 2002. The microcystins and nodularins: cyclic
polypeptide inhibitors of PP1 and PP2A. Current Medicinal Chemistry, 9(22): 1991-2003.
Gulledge, B. M., Aggen, J. B., and Chamberlin, A. R. 2003a. Linearized and truncated microcystin
analogues as inhibitors of protein phosphatases 1 and 2A. Bioorganic & Medicinal Chemistry
Letters, 13(17): 2903-2906.
Gulledge, B. M., Aggen, J. B., Eng, H., et al. 2003b. Microcystin analogues comprised only of Adda and
a single additional amino acid retain moderate activity as PP1/PP2A inhibitors. Bioorganic &
Medicinal Chemistry Letters, 13(17): 2907-2911.
Gupta, Nidhi, et al. 2003. Comparative toxicity evaluation of cyanobacterial cyclic peptide toxin
microcystin congeners (LR, RR, YR) in mice. Toxicology, 188(2): 285-296.
Gutierrez-Praena, D., Jos, A., Pichardo, S., et al. 2013. Presence and bioaccumulation of microcystins and
cylindrospermopsin in food and the effectiveness of some cooking techniques at decreasing their
concentrations: A review. Food and Chemical Toxicology, 53: 139-152.
Guzman, R. E., Solter, P. F., 1999. Hepatic oxidative stress following prolonged sublethal Microcystin
LR exposure. Toxicologic Pathology, 27: 582-588.
Haddix, P. L., Hughley, C. J., and Lechevallier, M. W. 2007. Occurrence of microcystins in 33 US water
supplies. Journal American Water Works Association, 99(9): 118-125.
Health Effects Support Document for Microcystins - June, 2015
102
-------
Hamel, K. 2009. Freshwater Algae Control Program, Report to the Washington State Legislature (2008-
2009) and (2010-2011). Publication No. 09-10-082 and No. 12-10-016. Water Quality Program,
Washington State Department of Ecology, Olympia, WA. Retrieved form the World Wide Web
https ://fortress .wa.gov/ecy/publications/publications/1210016 .pdf
Hamel, K. 2012. Aquatic Algae Control Program: Report to the Washington State Legislature (2010-
2011). Publication No. 12-10-016.
https ://fortress .wa.gov/ecy/publications/summarypages/1210016 .html
Han, J., Jeon, B., and Park, H. 2012. Cyanobacteria cell damage and cyanotoxin release in response to
alum treatment Water Science & Technology: Water Supply, 12(5): 549-555.
Haney, J. and M. Ikawa. 2000. A Survey of 50 NH Lakes for Microcystin (MCs). Final Report Prepared
for N.H. Department of Environmental Services by University of New Hampshire. Retrieved
September 25, 2012 from the World Wide Web:
http://water.usgs.gov/wrri/AnnualReports/2000/NHfy2000_annual_report.pdf
Hao, L., Xie, P., Li., H., et al. 2010. Transcriptional alteration of cytoskeletal genes induced by
microcystins in three organs of rats. Toxicon, 55(7): 1378-1386.
Harada, K., Ogawa, K. Matsuura, K., et al. 1990. Structural determination of geometrical isomers of
microcystins LR and RR from cyanobacteria by two-dimensional NMR spectroscopic techniques.
Chemical Research in Toxicology, 3(5): 473-481.
Harada K., Ogawa, K., Kimura, Y., et al. 1991. Microcystins from Anabaena flos-aquae NRC 525-17.
Chemical Research in Toxicology, 4: 535-540.
Hastie, C.J., Borthwick, E. B., Morrison, L. F., et al. 2005. Inhibition of several protein phosphatases by a
non-covalently interacting microcystin and a novel cyanobacterial peptide, nostocyclin.
Biochimica etBiophysica Acta, 1726: 187-193.
Hayakawa, K. and Kohama, K. 1995. Reversible effects of okadaic acid and microcystin-LR on the ATP-
dependent interaction between actin and myosin. Journal of Biochemistry, 117(3): 509-514.
Health Canada. 2002. Guidelines for Canadian Drinking Water Quality: Supporting Documentation -
Cyanobacterial Toxins-Microcystin-LR. Water Quality and Health Bureau, Healthy
Environments and Consumer Safety Branch, Health Canada, Ottawa, Ontario. Available at
http://www.hc-sc.gc.ca/ewh-semt/pubs/water-eau/doc_sup-appui/index_e.html.
Health Canada. 2012. Toxicity Profile for Cyanobacterial Toxins. Prepared for Water Quality and Science
Division of Health Canada by MTE GlobalTox. MTE File No.: 36348-100. pp. 48.
Heinze, R. 1999. Toxicity of the cyanobacterial toxin microcystin-LR to rats after 28 days intake with the
drinking water. Environmental Toxicology, 14(1): 57-60.
Heinze, R., Fastner, J., Neumann, U. and, Chorus, I. 2001. Testing cyanobacterial toxicity with primary
rat hepatocyte and cell-line assays. In: I. Chorus, (Ed.) Cyanotoxins: Occurrence, Causes,
Consequences. Springer-Verlag, New York, NY. pp. 317-324.
Henri, J., Huguet, A., Delmas, J. M., et al. 2014. Low in vitro permeability of the cyanotoxin microcystin-
LR across a Caco-2 monolayer: With identification of the limiting factors using modelling.
Toxicon, 91: 5-14
Health Effects Support Document for Microcystins - June, 2015
103
-------
Herfindal, L. and Selheim, F. 2006. Microcystin produces disparate effects on liver cells in a dose
dependent manner. Mini Reviews in Medicinal Chemistry, 6(3): 279-285.
Hermansky, S. J., Casey, P.J., and Stohs, S.J. 1990a. Cyclosporin A - a chemoprotectant against
microcystin-LR toxicity. Toxicology Letters, 54(2-3): 279-285.
Hermansky S. J., Wolff, S. N., and Stohs, S. J. 1990b. Use of rifampin as an effective chemoprotectant
and antidote against microcystin-LR toxicity. Pharmacology, 41(4): 231-236.
Hermansky, S. J., Stohs, S. J., Eldeen, Z. M., et al. 1991. Evaluation of potential chemoprotectants against
microcystin-LR hepatotoxicity in mice. Journal of Applied Toxicology, 11(1): 65-73.
Hernandez, M., Macia, M., Padilla, C., and Del Campo, F. F. 2000. Modulation of human
polymorphonuclear leukocyte adherence by cyanopeptide toxins. Environmental Research, 84(1):
64-68.
Heussner, A. H., Mazija, L., Fastner, J., and Dietrich, D. R. 2012. Toxin content and cytotoxicity of algal
dietary supplements. Toxicology and Applied Pharmacology, 265: 263-271.
Hilborn E., Carmichael, W., Yuan M., and Azevedo, S. 2005. A simple colorimetric method to detect
biological evidence of human exposure to microcystins. Toxicon, 46(2): 218-221.
Hilborn E., Carmichael, W., Yuan M., et al. 2007. Serologic evaluation of human microcystin
exposure. Environmental Toxicology. 22(5): 459-463.
Hilborn E. D., Soares R. M., Servaites J. C., et al. 2013. Sublethal Microcystin Exposure and Biochemical
Outcomes among Hemodialysis Patients. PLoS ONE, 8(7): e69518
Hitzfeld, B. Hoeger, S. J., and Dietrich, D. R. 2000. Cyanobacterial Toxins: Removal during Drinking
Water Treatment, and Human Risk Assessment. Environmental Health Perspectives, 108: 113-
122.
Hoeger, S. J. and Dietrich, D. R. 2004. Possible health risks arising from consumption of blue-green algae
food supplements. Sixth International Conference on Toxic Cyanobacteria, Bergen, Norway, pp.
30. Abstract
Honjo M., Matsui, K., Ueki, M., et al. 2006. Diversity of virus-like agents killing Microcystis aeruginosa
in a hyper-eutrophic pond. Journal of Plankton Research, 28(4): 407-412.
Honkanen, R. E., Zwiller, J., Moore, R. E., et al. 1990. Characterization of microcystin-LR, a potent
inhibitor of type 1 and type 2A protein phosphatases. Journal of Biological Chemistry, 265(32):
19401-19404.
Hooser, S.B., Waite, L. L., Beasley, V. R., et al. 1989. Microcystin-A induces morphologic and
cytoskeletal hepatocyte changes in vitro. Toxicon, 27(1): 50-51.
Hooser, S. B., Kuhlenschmidt, M. S., Dahlem, A. M., et al. 1991a. Uptake and subcellular localization of
tritiated dihydro-microcystin-LR in rat liver. Toxicon, 29(6): 589-601.
Hooser, S. B., Beasley, V. R., Waite, L. L., et al. 1991b. Actin filament alterations in rat hepatocytes
induced in vivo and in vitro by microcystin-LR, a hepatoxin from the blue-green alga, Microcystis
aeruginosa. Veterinary Pathology, 28(4): 259-266.
Health Effects Support Document for Microcystins - June, 2015
104
-------
Hu, Z., Chen, H., Li, Y., et al. 2002. The expression of bcl-2 and bax genes during microcystin induced
liver tumorigenesis. Zhonghua Yu Fang Yi Xue Za Zhi, 36(4): 239-242. (Chinese)
Hu, Z., Chen, H., Pang, C., et al. 2008. The expression of p53 and pl6 in the course of microcystin-LR
inducing of liver tumor. The Chinese-German Journal of Clinical Oncology, 7(12): 690-693.
Hu, Z., Chen, H., Xue, J., et al. 2010. The expression of Bcl-2 and Bax produced by sub-chronic
intoxication with the cyanotoxin Microcystin-LR. The Chinese-German Journal of Clinical
Oncology. 9(2): 68-72.
Huang, W. J., Lai, C. H., and Cheng, Y.-L. 2007. Evaluation of extracellular products and mutagenicity in
cyanobacteria cultures separated from a eutrophic reservoir. Science of the Total Environment,
377(2-3): 214-223.
Huang, P., Zheng, Q., and Xu, L.-H. 2011. The apoptotic effect of oral administration of microcystin-RR
on mice liver. Environmental Toxicology, 26: 443-452.
Hudnell, H. K. (ed.). 2008. Cyanobacterial Harmful Algae Blooms, State of the Science and Research
Needs. Proceedings of the Interagency, International Symposium on Cyanobacterial Harmful
Algal Blooms. RTP North Carolina, Sept. 2005. Advances in Experimental Medicine & Biology.
Springer Science. Vol. 619, pp. 948.
Hudnell, H. K. 2010. The state of U.S. freshwater harmful algal blooms assessments policy and
legislation. Toxicon, 55: 1024-1034.
Huisman, J., Matthijs, H. C. P., and Visser, P. M. 2005. Harmful Cyanobacteria. Springer, Dordrecht.
Humpage, A. R. and Falconer, I. R. 1999. Microcystin-LR and liver tumor promotion: Effects on
cytokinesis, ploidy, and apoptosis in cultured hepatocytes. Environmental Toxicology, 14(1): 61-
75.
Humpage, A. R., Hardy, S. J., Moore. E. J., et al. 2000. Microcystins (cyanobacterial toxins) in drinking
water enhance the growth of aberrant crypt foci in the mouse colon. Journal of Toxicology and
Environmental Health, Part A, 61(3): 155-165.
Ibelings, B. W., Bruning, K., de Jonge, J., et al. 2005. Distribution of microcystins in a lake foodweb: no
evidence for biomagnification: Microbial Ecology, 49(4): 487-500.
Ibelings, B.W. and Chorus, I. 2007. Accumulation of cyanobacterial toxins in freshwater "seafood" its
consequences for public health: a review. Environmental Pollution, 150:177-192.
Ibelings, B. W., and Havens, K. E. 2008. Cyanobacterial toxins: a qualitative meta-analysis of
concentrations, dosage and effects in freshwater, estuarine and marine biota. Advances in
Experimental Medicine and Biology, 619: 675-732.
ILS (Integrated Laboratory Systems). 2000. Cylindrospermopsin: Review of ToxicologicalLiterature.
Prepared by Integrated Laboratory Systems, for National Toxicology Program, NIEHS, EPA. pp.
37.
Imanishi, S. and Harada, K.-I. 2004. Proteomics approach on microcystin binding proteins in mouse liver
for investigation of microcystin toxicity. Toxicon, 43: 651-659.
(IARC) International Agency for Research on Cancer. 2010. IARC Monographs on the Evaluation of
Carcinogenic Risks to Humans: Ingested Nitrate and Nitrite and Cyanobacterial Peptide Toxins.
Health Effects Support Document for Microcystins - June, 2015
105
-------
Ito, E. and Nagai, H. 2000. Bleeding from the small intestine caused by aplysiatoxin, the causative agent
of the red alga Gracilaria coronopifolia poisoning. Toxicon, 38: 123-132.
Ito, E., Kondo, F., and Harada, K.-I. 1997a. Hepatic necrosis in aged mice by oral administration of
microcystin-LR. Toxicon, 35(2): 231-239.
Ito, E., Kondo, F., Terao, K., and. Harada, K.-I. 1997b. Neoplastic nodular formation in mouse liver
induced by repeated i.p. injections of microcystin-LR. Toxicon, 35(9): 1453-1457.
Ito, E., Kondo, F., and Harada, K.-I. 2001. Intratracheal administration of microcystin-LR, and its
distribution. Toxicon, 39(2-3): 265-271.
Ito, E., Satake, M., and Yasumoto, T. 2002a. Pathological effects of lyngbyatoxin A upon mice. Toxicon,
40: 551-556.
Ito, E., Takai, A., Kondo, F., et al. 2002b. Comparison of protein phosphatase inhibitory activity and
apparent toxicity of microcystins and related compounds. Toxicon, 40(7): 1017-1025.
Jaag, O. 1945. Untersuchungen fiber die Vegetation and Biologie der Algan des nackten Gesteins in den
Alpen, im Jura and im schweizerischen Mittelland. Kryptogamenflora der Schweiz, Band IX,
Heft 3. Kommissionsverlag Buchdruckerei Btichler and Co., Bern. (Cited in WHO 1999)
Jarvenpaa, S., Lundberg-Niinisto, C., Spoof, L., et al. 2007. Effects of microcystins on broccoli and
mustard, and analysis of accumulated toxin by liquid chromatography-mass spectrometry.
Toxicon, 49: 865-874.
Jasionek, G., Zhdanov, A., Davenport, J., et al. 2010. Mitochondrial toxicity of microcystin-LR on
cultured cells: application to the analysis of contaminated water samples. Environmental Science
& Technology, 44(7): 2535-2541.
Jensen, H.S. and Andersen, F. O. 1992. Importance of temperature, nitrate, and pH for phosphate release
from aerobic sediments of 4 shallow, eutrophic lakes. Limnology and Oceanography, 37: 577-
589.
Jensen, G. S., Ginsberg, D. I., and Drapeau, C. 2001. Blue-green algae as an immuno-enhancer and
biomodulator. Journal of the American Medical Association, 3: 24-30.
Jeppesen, E., Sondergaard. M., Meerhoff, M., et al. 2007. Shallow lake restoration by nutrient loading
reduction-some recent finding and challenges ahead. Hydrobiologia. 584: 239-252.
Jeppesen, E., Kronvang, B., Meerhoff, M., et al. 2009. Climate change effects on runoff, catchment
phosphorus loading and lake ecological state, and potential adaptations. Journal of Environmental
Quality, 38: 1930-1941.
Jeppesen, E., Meerhoff, M., Holmgren, K., et al. 2010. Impacts of climate warming on lake fish
community structure and dynamics, and potential ecosystem effects. Hydrobiologia, 646: 73-90.
Ji, Y., Lu, G., Chen, G., et al. 2011. Microcystin-LR Induces Apoptosis via NF-kappaB/iNOS Pathway in
INS-1 Cells. International Journal of Molecular Sciences, 12(7): 4722-4734.
Jiao, D. A., Shen, G. F, Shen, Y. Z., and Zheng, G. M. 1985. The case-control study of colorectal cancer.
Chinese Journal of Epidemiology, 6:285-288 (As cited in Zhou et al., 2002). (Chinese)
Health Effects Support Document for Microcystins - June, 2015
106
-------
Jochimsen, E. M., Carmichael W. W., An J. S., et al. 1998. Liver failure and death after exposure to
microcystins at a hemodialysis center in Brazil. The New England Journal of Medicine, 338(13):
873-8.
Jones, G. J., Blackburn, S. I., and Parker, N. S. 1994. A toxic bloom of Nodularia spumigenaMertens in
Orielton Lagoon, Tasmania. Australian Journal of Marine & Freshwater Research, 45: 787-800.
(Cited in WHO 1999)
Kaebernick, M., Neilan, B. A., Borner, T., and Dittman, E. 2000. Light and the transcriptional response of
the microcystin biosynthesis gene cluster. Applied and Environmental Microbiology, 8: 3387-
3392.
Kann, E. 1988. Zur Autokologie benthischer Cyanophyten in reinen europaischen Seen and
Fliessgewassem. Algological Studies, 50-53: 473-495. (Cited in WHO 1999)
Khan, S. A., Ghosh, S., Wickstrom, M. L., et al. 1995. Comparative pathology of microcystin-LR in
cultured hepatocytes, fibroblasts and renal epithelial cells. Natural Toxins, 3(3): 119-128.
Kirpenko, Y. A., Sirenko, L. A., and Kirpenko, N. I. 1981. Some aspects concerning remote after effects
of Blue-green Algae toxin impact on animals. In: W. W. Carmichael (Ed.), The Water
Environment: Algal Toxins and Health. Plenum Press, pp.257-270.
Klassen, C. D. and Aleksunes, L. M. 2006. Xenobiotic bile acid and cholesterol transporters: function and
regulation. Pharmacological Reviews, 62: 1-96. (As cited in Zhou et al., 2012).
Knapp J., Aleth, S., Balzer, F., et al. 2002. Calcium-independent activation of the contractile apparatus in
smooth muscle of mouse aorta by protein phosphatase inhibition. Naunyn-Schmiedeberg's
Archives of Pharmacology, 366(6): 562-569.
Komatsu, M., Furukawa, T., Ikeda, R., et al. 2007. Involvement of mitogen-activated protein kinase
signaling pathways in microcystin-LR-induced apoptosis after its selective uptake mediated by
OATP1B1 and OATP1B3. Toxicological Sciences, 97(2): 407-416.
Kondo, F., Ikai, Y., Oka, H., et al. 1992. Formation, characterization, and toxicity of the glutathione and
cysteine conjugates of toxic heptapeptide microcystins. Chemical Research in Toxicology, 5(5):
591-596.
Kondo, F., Matsumoto, H., Yamada, S., et al. 1996. Detection and identification of metabolites of
microcystins in mouse and rat liver. Chemical Research in Toxicology, 9: 1355-1359.
Kosakowska, A., Nedzi, M., and Pempkowiak, J. 2007. Responses of the toxic cyanobacterium
Microcystis aeruginosa to iron and humic substances. Plant Physiology and Biochemistry, 45:
365-370.
Kosten, S, Huszar, V. L. M., Carees, E. B., et al. 2011. Warmer climates boost cyanobacterial dominance
in shallow lakes. Global Change Biology, 18: 118-126.
Kotak, B. G., Zurawell, R. W., Prepas, E. E., and Holmes, C. F. B. 1996. Microcystin-LR concentration
in aquatic food web compartments from lakes of varying trophic status. Canadian Journal of
Fisheries and Aquatic Sciences, 53: 1974-1985.
Kujbida, P., Hatanaka, E., Campa, A., et al. 2006. Effects of microcystins on human polymorphonuclear
leukocytes. Biochemical and Biophysical Research Communications, 341(1): 273-277.
Health Effects Support Document for Microcystins - June, 2015
107
-------
Kujbida, P., Hatanaka, E., Campa, A., et al. 2008. Analysis of chemokines and reactive oxygen species
formation by rat and human neutrophils induced by microcystin-LA, -YR and -LR. Toxicon,
51(7): 1274-1280.
Kujbida, P., Hatanaka, E., Ramirez Vinolo, M. A., et al. 2009. Microcystins -LA, -YR, and -LR action on
neutrophil migration. Biochemical and Biophysical Research Communications, 382(1): 9-14.
Laamanen, M. 1996. Cyanoprokaryotes in the Baltic Sea ice and winter plankton. Algological Studies, 83:
423-433. (Cited in WHO 1999)
Lahti, K., Niemi, M. R., Rapala, J., and Sivonen, K. 1997a. Bi ode gradation ofcyanobacterial
hepatotoxins - characterization of toxin degrading bacteria. Proceedings of the VII International
Conference on Harmful Algae. (Cited in WHO 1999).
Lahti, K., Rapala, J., Farding, M., et al. 1997b. Persistence of cyanobacterial hepatotoxin, microcystin-
LR, in particulate material and dissolved in lake water. Water Research, 31(5): 1005-1012. (Cited
in WHO 1999)
Lam, A. K.-Y., Fedorak, P. M., and Prepas, E. E. 1995. Biotransformation of the cyanobacterial
hepatotoxin microcystin-LR, as determined by HPLC and protein phosphatase bioassay.
Environmental Science & Technology, 29: 242-246.
Lankoff, A., Banasik, A., Obe, G., et al. 2003. Effect of microcystin-LR and cyanobacterial extract from
Polish reservoir drinking water on cell cycle progression, mitotic spindle, and apoptosis in CHO-
K1 cells. Toxicology and Applied Pharmacology, 189(3): 204-213.
Lankoff, A., Krzowski, L., Glab, J., et al. 2004a. DNA damage and repair in human peripheral blood
lymphocytes following treatment with microcystin-LR. Mutation Research, 559(1-2): 131-142.
Lankoff, A., Carmichael, W. W., Grasman, K. A., and Yuan, M. 2004b. The uptake kinetics and
immunotoxic effects of microcystin-LR in human and chicken peripheral blood lymphocytes in
vitro. Toxicology, 204: 23-40.
Lankoff, A., Bialczyk, J., Dziga, D., et al. 2006a. The repair of gamma-radiation-induced DNA damage is
inhibited by microcystin-LR, the PP1 and PP2A phosphatase inhibitor. Mutagenesis, 21(1): 83-
90.
Lankoff, A., Bialczyk, J., Dziga, D., et al. 2006b. Inhibition of nucleotide excision repair (NER) by
microcystin-LR in CHO-K1 cells. Toxicon, 48(8): 957-965.
La-Salete, R., Oliveira, M. M., Palmeira, C. A., et al. 2008. Mitochondria a key role in microcystin-LR
kidney intoxication. Journal of Applied Toxicology, 28(1): 55-62.
Lei, L. M., Song, L. R., and Han, B. P. 2006. Microcystin-LR induces apoptosis in L-02 cell line. Nan
Fang Yi Ke DaXue Xue Bao, 26(4): 386-389.
Leiers, T., Bihlmayer, A., Ammon, H. P. T., and Wahl, M. A. 2000. [Ca2+](i)- and insulin-stimulating
effect of the non-membranepermeable phosphatase-inhibitor microcystin-LR in intact insulin-
secreting cells (RINm5F). British Journal of Pharmacology, 130(6): 1406-1410.
Lewis, W. M., Wurtsbaugh, W. A., and Paerl, H. W. 2011. Rationale for control of anthropogenic
nitrogen and phosphorus in inland waters. Environmental Science & Technology, 45: 10030-
10035.
Health Effects Support Document for Microcystins - June, 2015
108
-------
Li, Y. and Han, X. 2012. Microcystin-LR causes cytotoxicity effects in rat testicular Sertoli cells.
Environmental Toxicology and Pharmacology, 33(2): 318-326.
Li, Y., Sheng, J., Sha, J., and Han, X. 2008. The toxic effects of microcystin-LR on the reproductive
system of male rats in vivo and in vitro. Reproductive Toxicology, 26(3-4): 239-245.
Li, H., Xie, P., Li, G., et al. 2009. In vivo study on the effects of microcystin extracts on the expression
profiles of proto-oncogenes (c-fos, c-jun and c-myc) in liver, kidney and testis of male Wistar rats
injected i.v. with toxins. Toxicon, 53(1): 169-175.
Li, Y., Chen, J.-A., Zhao, Q., et al. 201 la. A Cross-Sectional Investigation of Chronic Exposure to
Microcystin in Relationship to Childhood Liver Damage in the Three Gorges Reservoir Region,
China. Environmental Health Perspectives, 119(10): 1483-1488.
Li, D., Liu, Z., Cui, Y., et al. 201 lb. Toxicity of cyanobacterial bloom extracts from Taihu Lake on
mouse, Mus musculus. Ecotoxicology, 20(5): 1018-1025.
Li, G., Xie, P., Li., H.-Y., et al. 201 lc. Involvement of p53, Bax, and Bcl-2 pathway in microcystins-
induced apoptosis in rat testis. Environmental Toxicology, 26(2): 111-117.
Li, T., Huang, P., Liang, J., et al. 201 Id. Microcystin-LR (MCLR) induces a compensation of PP2A
activity mediated by alpha4 protein in HEK293 cells. International Journal of Biological Science,
7(6):740-52.
Li, G., Xie, P., Li, H., et al. 201 le. Acute effects of microcystins on the transcription of 14 glutathione S-
transferase isoforms in Wistar rat. Environmental Toxicology, 26(2): 187-194.
Li, G., Yan, W., Cai, F.,et al. 2012. Spatial learning and memory impairment and pathological change in
rats induced by acute exposure to microcystin-LR. Environmental Toxicology, 29(3): 261-268.
Li, X., Zhang, X., Ju, J., Li, Y., Yin, L., and Pu, Y. 2014. Alterations in neurobehaviors and inflammation
in hippocampus of rats by oral administration of microcystin-LR. Environmental Science and
Pollution Research. 21:12419-12424.
Li, X., Zhang, X., Ju, J., Li, Y., Yin, L., and Pu, Y. 2015. Maternal repeated oral exposure to microcystin-
LR affects neurobehaviors in developing rats. Environmental Toxicology and Chemistry.
34(l):64-69.
Liang, J., Li, T., Zhang, Y.-L., et al. 2011. Effect of microcystin-LR on protein phosphatase 2A and its
function in human amniotic epithelial cells. Journal ofZhejiang University SCIENCE B, 12(12):
951-960.
Lin, J. R. and Chu, F. S. 1994. Kinetics of distribution of microcystin-LR in serum and liver cytosol of
mice: an immunochemical analysis. Journal of Agricultural and Food Chemistry, 42(4): 1035-
1040.
Liu, Y., Xie, P., Qiu, T., et al. 2010. Microcystin extracts induce ultrastructural damage and biochemical
disturbance in male rabbit testis. Environmental Toxicology, 25(1): 9-17.
Liu, J., Wei, Y., and Shen, P. 2011. Effect of membrane permeability transition on hepatocyte apoptosis
of the microcystin-LR-induced mice. Wei Sheng Yan Jiu, 40(1): 53-56.
Health Effects Support Document for Microcystins - June, 2015
109
-------
Long B.M., Jones, G. J., and Orr, P.T. 2001. Cellular microcystin content in N-limitedMicrocystis
aeruginosa can be predicted from growth rate. Applied and Environmental Microbiology, 67(1):
278-83.
Lovell, R. A., Schaeffer, D. J., Hooser, S. B., et al. 1989. Toxicity of intraperitoneal doses of microcystin-
LR in two strains of male mice. Journal of Environmental Pathology, Toxicology and Oncology,
9(3): 221-237.
Lu, H., Choudhuri, S., Ogura, K., et al. 2008. Characterization of organic anion transporting polypeptide
lb2-null mice: essential role in hepatic uptake/toxicity of phalloidin and microcystin-LR.
Toxicological Sciences, 103(1): 35-45.
Maatouk, I., Bouai'cha, N., Plessis, M. J., and Perin, F. 2004. Detection by 32P-postlabelling of 8-oxo-7,8-
dihydro-2'-deoxyguanosine in DNA as biomarker of microcystin-LR- and nodularin-induced
DNA damage in vitro in primary cultured rat hepatocytes and in vivo in rat liver. Mutation
Research, 564(1): 9-20.
MacKintosh, C., Beattie, K. A., Klumpp, S., et al. 1990. Cyanobacterial microcystin-LR is a potent and
specific inhibitor of protein phosphatases 1 and 2A from both mammals and higher plants. FEBS
Letters, 264(2): 187-192.
MacKintosh, R.W., Dalby, K. N., Campbell, D. G., et al. 1995. The cyanobacterial toxin microcystin
binds covalently to cysteine-273 on protein phosphatase 1. FEBS Letters, 371(3): 236-240.
Maidana, M., Carlis, V., Galhardi, F.G., et al. 2006. Effects of microcystins over short- and long-term
memory and oxidative stress generation in hippocampus of rats. Chemico-Biological Interactions,
159(3): 223-234.
Magalhaes, V. F., Soares, R. M., and Azevedo, S. M. F. O. 2001. Microcystins contamination in fish from
the Jacarepagu'a Lagoon (RJ, Brazil): Ecological implication and human health risk. Toxicon, 39:
1077-1085.
Magalhaes, V. F., Marinho, M. M., Domingos, P., et al. 2003. Microcystins (cyanobacteria hepatotoxins)
bioaccumulation in fish and crustaceans from Sepetiba Bay (Brasil, RJ). Toxicon, 42: 289-295.
Makarewicz, J., Boyer, G., Guenther, W., et al. 2006. The Occurrence of Cyanotoxins in the Nearshore
and Coastal Embayments of Lake Ontario. Great Lakes Research Review, 7: 25-29.
Makarewicz, J., Boyer, G., Lewis, T., et al. 2009. Spatial and temporal distribution of the cyanotoxin
microcystin-LR in the Lake Ontario ecosystem: Coastal embayments, rivers, nearshore and
offshore, and upland lakes. Journal of Great Lakes Research, 35: 83-89.
Mankiewicz, J., M. Tarczynska, K.E. Fladmark et al. 2001. Apoptotic effect of cyanobacterial extract on
rat hepatocytes and human lymphocytes. Environmental Toxicology, 16(3): 225-233.
Matsushima, R., S. Yoshizawa, M.F. Watanabe et al. 1990. In vitro and in vivo effects of protein
phosphatase inhibitors, microcystins and nodularin on mouse skin and fibroblasts. Biochem.
Biochemical and Biophysical Research Communications, 171(2): 867-874.
Mattila K., Annila, A., and Rantala. T. T. 2000. Metal Ions Mediate the Binding of Cyanobacterial Toxins
to Human Protein Phosphatase I: A Computational Study. Oulu University Library, Oulun
Yliopisto, Oulu.
Health Effects Support Document for Microcystins - June, 2015
110
-------
Maynes J.T., Perreault, K. R., Cherney, M. M., et al. 2004. Crystal structure and mutagenesis of a protein
phosphatase-l:calcineurin hybrid elucidate the role of the (312-J313 loop in inhibitor binding.
Journal of Biological Chemistry, 279(41): 43198-43206.
Maynes, J.T., Luu, H. A., Cherney, M. M., et al. 2006. Crystal structures of protein phosphatase-1 bound
to motuporin and dihydromicrocystin-LA: Elucidation of the mechanism of enzyme inhibition by
cyanobacterial toxins. Journal of Molecular Biology, 356(1): 111-120.
McDermott, C.M., Nho, C. W., Howard, W., and Holton, B. 1998. The cyanobacterial toxin, microcystin-
LR can induce apoptosis in a variety of cell types. Toxicon, 36(12): 1981-1996.
McElhiney, J. and Lawton, L. A. 2005. Detection of the cyanobacterial hepatotoxins microcystins.
Toxicology and Applied Pharmacology, 203: 219-230.
Meng, G., Sun, Y., Fu, W., et al. 2011. Microcystin-LR induces cytoskeleton system reorganization
through hyperphosphorylation of tau and HSP27 via PP2A inhibition and subsequent activation
of the p38 MAPK signaling pathway in neuroendocrine (PC12) cells. Toxicology, 290: 218-229.
Meriluoto, J. A., Nygard, S. E., Dahlem, A. M., and Eriksson, J. E. 1990. Synthesis, organotropism and
hepatocellular uptake of two tritium-labeled epimers of dihydromicrocystin-LR, a cyanobacterial
peptide toxin analog. Toxicon, 28(12): 1439-1446.
Metcalf J. S., Beattie, K. A., Pflugmacher, S., and Codd, G. A. 2000. Immuno-crossreactivity and toxicity
assessment of conjugation products of the cyanobacterial toxin, microcystin-LR. FEMS
Microbiology Letters, 189(2): 155-158.
Metcalf, J., Richer, R., Cox, P., and Codd, G. 2012. Cyanotoxins in desert environments may present a
risk to human health. Science of the Total Environment, 421-422: 118-123.
Mikhailov, A., Harmala-Brasken, A. S., Hellman, J., et al. 2003. Identification of ATP-synthase as a
novel intracellular target for microcystin-LR. Chemico-Biological Interactions, 142(3): 223-237.
Miller, W. A., Toy-Choutka, S., Dominik, C., et al. 2010. Evidence for novel marine harmful algal
bloom: Cyanotoxin (microcystin) transfer from land to sea otters. PLoS One, 5: 1-11.
Milutinovic, A., Sedmak, B., Horvat-Znidarsic, I., and Suput, D. 2002. Renal injuries induced by chronic
intoxication with microcystins. Cellular and Molecular Biology Letters, 7(1): 139-141.
Milutinovic A, Zivin, M., Zore-Pleskovic, R., Sedmak, B., and Suput, D. 2003. Nephrotoxic effects of
chronic administration of microcystins-LR and -YR. Toxicon, 42(3): 281-288.
Milutinovic, A., Zorc-Pleskovic, R., Petrovic, D., et al. 2006. Microcystin-LR induces alterations in heart
muscle. Folia Biologica (Praha), 52(4): 116-118.
Mishra, S., Payaningal, R., Huang, Z., and Vijayaraghavan, S. 2003. Binding and inactivation of the germ
cell-specific protein phosphatase PPly2 by sds22 during epididymal sperm maturation. Biology of
Reproduction, 69: 1572-1579.
Miura, G. A., Robinson, N. A., Lawrence, W. B., and Page, J. G. 1991. Hepatotoxicity of microcystin-LR
in fed and fasted rats. Toxicon, 29(3): 337-346.
Mohamed, Z. 2008. Toxic cyanobacteria and cyanotoxins in public hot springs in Saudi Arabia. Toxicon,
51: 17-27.
Health Effects Support Document for Microcystins - June, 2015
111
-------
Mohamed, Z. A. and A1 Shehri, A. M. 2009. Microcystins in groundwater wells and their accumulation in
vegetable plants irrigated with contaminated waters in Saudi Arabia. Journal of Hazardous
Materials, 172: 310-315.
Mohamed, Z. A., and Hussein, A. A. 2006. Depuration of microcystins in tilapia fish exposed to natural
populations of toxic cyanobacteria: A laboratory study. Ecotoxicology and Environmental Safety,
63(3): 424-429.
Z.A. Mohamed, Z.A., W.W. Carmichael, A.A. Hussein. 2003 Estimation of microcystins in the fresh
water fish Oreochromis niloticus in an Egyptian fish farm containing a. Microcystis bloom
Environmental Toxicology, 18:137-141
Moreno, I., Pichardo, S., Jos, A., et al. 2005. Antioxidant enzyme activity and lipid peroxidation in liver
and kidney of rats exposed to microcystin-LR administered intraperitoneally. Toxicon, 45(4):
395-402.
Mulvenna, V., Dale, K., Priestly, B., et al. 2012. Health Risk Assessment for Cyanobacterial Toxins in
Seafood. International Journal of Environmental Research and Public Health, 9(3): 807-820.
Namikoshi M., Choi, B. W., Sun, F., et al. 1993. Chemical characterization and toxicity of dihydro
derivatives of nodularin and microcystin-LR, potent cyanobacterial cyclic peptide hepatotoxins.
Chemical Research in Toxicology, 6(2): 151-158.
NDEQ (Nebraska Department of Environmental Quality). 2011. Microcystin Toxin Migration,
Bioaccumulation, and Treatment Fremont Lake #20 Dodge County, Nebraska. Prepared by Water
Quality Assessment Section, Water Division, Nebraska Department of Environmental Quality,
pp. 48.
Neilan, B. A., Pearson, L. A., Moffitt, M. C., et al. 2007. Chapter 17: The genetics and genomics of
cyanobacterial toxicity. In: H. K. Hudnell (Ed.), Proceedings of the Interagency, International
Symposium on Cyanobacterial Harmful Algal Blooms Advances in Experimental Medicine &
Biology, 423-458.
Nishiwaki, R., Ohta, T., Sueoka, E., et al. 1994. Two significant aspects of microcystin-LR: Specific
binding and liver specificity. Cancer Letters, 83: 283-289.
Nishiwaki-Matsushima, R., Nishiwaki, S., Ohta, T., et al. 1991. Structure-function relationships of
microcystins, liver tumor promoters, in interaction with protein phosphatase. Japanese Journal of
Cancer Research, 82(9): 993-996.
Nishiwaki-Matsushima, R., Ohta, T., Nishiwaki, S., et al. 1992. Liver tumor promotion by the
cyanobacterial cyclic peptide toxin microcystin-LR. Journal of Cancer Research and Clinical
Oncology, 118(6): 420-424.
Nobre, A. C. L., Jorge, M. C. M., Menezes, D.B., et al. 1999. Effects of microcystin-LR in isolated
perfused rat kidney. Brazilian Journal Of Medical and Biological Research, 32(8): 985-988.
Nobre, A. C. L., Coelho, G. R., Coutinho, M. C. M., et al. 2001. The role of phospholipase A(2) and
cyclooxygenase in renal toxicity induced by microcystin-LR. Toxicon, 39(5): 721-724.
Nobre, A. C. L., Martins, A. M. C., Havt, A., et al. 2003. Renal effects of supernatant from rat peritoneal
macrophages activated by microcystin-LR: Role protein mediators. Toxicon, 41(3):377-381.
Health Effects Support Document for Microcystins - June, 2015
112
-------
Nong, Q., Komatsu, M., Izumo, K., et al. 2007. Involvement of reactive oxygen species in Microcystin-
LR-induced cytogenotoxicity. Free Radical Research, 41(12): 1326-1337.
NRC (National Research Council). 1983. Risk Assessment in the Federal Government: Managing the
Process. National Academy Press, Washington, DC.
Ohio EPA (OHEPA) 2012. 2011 Grand Lake St. Marys Algal Toxin Sampling Data. Retrieved September
25, 2012 from the World Wide Web: http://www.epa.state.oh.us/dsw/HAB.aspx
Ohta, T., Nishiwaki, R., Yatsunami, J., et al. 1992. Hypersphosphorylation of cytokeratins 8 and 18 by
microcystin-LR, a new liver tumor promoter, in primary cultured rat hepatocytes.
Carcinogenesis, 13(12): 2443-2447.
Ohta, T., Sueoka, E., Iida, N., et al. 1994. Nodularin, a potent inhibitor of protein phosphatases 1 and 2A,
is a new environmental carcinogen in male F344 rat liver. Cancer Research, 54(24): 6402-6406.
O'Neil, J., Davis, T., Burford, M., and Gobler, C. 2012. The rise of harmful cyanobacteria blooms: The
potential roles of eutrophication and climate change. Harmful Algae, 14: 313-334.
O'Reilly, A., Wanielista, M., Loftin, K., and Chang, N. 2011. Laboratory simulated transport of
microcystin-LR and cylindrospermopsin in groundwater under the influence of storm water ponds:
implications for harvesting of infiltrated stormwater. GQ10: Groundwater Quality Management
in a Rapidly Changing World (Proc. 7th International Groundwater Quality Conference held in
Zurich, Switzerland, 13-18 June 2010). IAHS Publ 342, 2011, 107-111.
Orihel, D. M., Bird, D. F., Brylinsky, M., et al. 2012. High microcystin concentrations occur only at low
nitrogen-to-phosphorus ratios in nutrient-rich Canadian lakes. Canadian Journal of Fisheries and
Aquatic Sciences, 69: 1457-1462.
Orr, P. T., Jones, G. J., Hunter, R. A., et al. 2001. Ingestion of toxic Microcystis aeruginosa by dairy
cattle and the implications for microcystin contamination of milk. Toxicon, 39: 1847-1854.
Orr, P. T., Jones, G. J., Hunter, R. A., and Berger, K. 2003. Exposure of beef cattle to sub-clinical doses
of Microcystis aeruginosa: toxin bioaccumulation, physiological effects and human health risk
assessment. Toxicon, 41: 613-620.
Pace, J. G., Robinson, N. A., Miura, G. A., et al. 1991. Toxicity and kinetics of 3H microcystin-LR in
isolated perfused rat livers. Toxicology and Applied Pharmacology, 107(3): 391-401.
Paerl H. W. and Huisman, J. 2008. Blooms like it hot. Science, 320: 57-58.
Paerl, H. and Scott, J. 2010. Throwing Fuel on the Fire: Synergistic Effects of Excessive Nitrogen Inputs
and Global Warming on Harmful Algal Blooms. Environmental Science & Technology, 44: 7756-
7758.
Paerl, H., Xu, H., McCarthy, M., et al. 2011. Controlling harmful cyanobacterial blooms in ahyper-
eutrophic lake (Lake Taihu, China): The need for a dual nutrient (N & P) management strategy.
Water Research, 45(5): 1973-1983.
Paerl, H.W. and Otten, T. G. 2013a. Blooms Bite the Hand That Feeds Them. Science, 342(25): 433-434.
Paerl, H.W. and Otten, T. G. 2013b. Harmful Cyanobacterial Blooms: Causes, Consequences, and
Controls. Microbial Ecology, 65: 995-1010.
Health Effects Support Document for Microcystins - June, 2015
113
-------
Papadimitriou, T., Kagalou, I., Bacopoulos, V., and Leonardos, I. D. 2010. Accumulation of microcystins
in water and fish tissues: an estimation of risks associated with microcystins in most of the Greek
lakes. Environmental Toxicology, 25: 418-427.
Papadimitriou, T., Kagalou, I., Stalikas, C., et al. 2012. Assessment of microcystin distribution and
biomagnification in tissues of aquatic food web compartments from a shallow lake and evaluation
of potential risks to public health. Ecotoxicology, 21: 1155-1166.
Poste, A. E., Hecky, R. E., and Guildford, S. J. 2011. Evaluating microcystin exposure risk through fish
consumption. Environmental Science & Technology, 45: 5806-5811.
Prepas, E. E., Kotak, B. G., Campbell, L. M., et al. 1997. Accumulation and elimination of cyanobacterial
hepatotoxins by the freshwater clam Anodonta grandis simpsoniana. Canadian Journal of
Fisheries and Aquatic Sciences, 54: 41-46. (Cited in WHO 1999)
Puerto, M., S. Pichardo, et al. 2009 Comparison of the toxicity induced by microcystin-RR and
microcystin-YR in differentiated and undifferentiated Caco-2 cells. Toxicon. 54(2): 161-169.
Puddick, J., Prinsep, M., Wood, S., et al. 2015. Further Characterization of Glycine-Containing
Microcystins from the McMurdo Dry Valleys of Antarctica. Toxins, 7: 493-515
Qin, W., Xu, L., Zhang, X., et al. 2010. Endoplasmic reticulum stress in murine liver and kidney exposed
to microcystin-LR. Toxicon, 56(8): 1334-1341.
Rai, A. N. 1990. CRC Handbook of Symbiotic Cyanobacteria. CRC Press, Boca Raton, FL. pp. 253.
(Cited in WHO 1999)
Rao, P. V. L. and Bhattacharya, R. 1996. The cyanobacterial toxin microcystin-LR induced DNA damage
in mouse liver in vivo. Toxicology, 114(1): 29-36.
Rao, P. V. L., Gupta, N., Bhaskar, A. S. B., and Jayaraj, R. 2002. Toxins and bioactive compounds from
cyanobacteria and their implications on human health. Journal of Environmental Biology, 3: 215-
224.
Rao, P.V.L., Gupta, N., Jayaraj, R., et al. 2005. Age-dependent effects on biochemical variables and
toxicity induced by cyclic peptide toxin microcystin-LR in mice. Comparative Biochemistry and
Physiology Part C: Toxicology & Pharmacology, 140(1): 11-19.
Rapala, J., Lahti, K., Sivonen, K., and Niemeld, S. 1994. Biodegradability and adsorption on lake
sediments of cyanobacterial hepatotoxins and anatoxin-a. Letters in Applied Microbiology, 19:
423-428. (Cited in WHO 1999)
Rapala, J., Niemela, M., Berg, K., et al. 2006. Removal of cyanobacteria, cyanotoxins, heterotropohic
bacteria and endotoxins at an operating surface water treatment plant. Water Science and
Technology, 54: 3:23.
Repavich, W. M., Sonzogni, W. C., Standridge, J. H., et al.1990. Cyanobacteria (blue-green algae) in
Wisconsin waters: Acute and chronic toxicity. Water Research, 24(2): 225-231.
Reynolds, C. S. 2006. The Ecology ofPhytoplankton. Cambridge University Press, Cambridge.
Rinehart K. L., Namikoshi, M., and Choi, B. W. 1994. Structure and biosynthesis of toxins from blue-
green algae (cyanobacteria). Journal of Applied Phycology, 6: 159-176.
Health Effects Support Document for Microcystins - June, 2015
114
-------
Robarts, R. D. and Zohary, T. 1987. Temperature effects on photosynthetic capacity, respiration, and
growth rates of bloom-forming cyanobacteria. New Zealand Journal of Marine and Freshwater
Research, 21: 391-399.
Robinson, N. A., Miura, G. A., Matson, C. F., et al. 1989. Characterization of chemically tritiated
microcystin-LR and its distribution in mice. Toxicon, 27: 1035-1042.
Robinson, N.A., Pace, J. G., Matson, C. F., et al. 1991. Tissue distribution, excretion and hepatic
biotransformation of microcystin-LR in mice. Journal of Pharmacology and Experimental
Therapeutics, 256(1): 176-182.
Runnegar, M. T. C. and Falconer, I. R. 1982. The in vivo and in vitro biological effects of the peptide
hepatotoxin from the blue-green alga Microcystis aeruginosa. South African Journal of Science,
78: 363-366.
Runnegar, M. T. C. and Falconer, I. R. 1986. Effect of toxin from the cyanobacterium Microcystis
aeruginosa on ultrastructural morphology and actin polymerization in isolated hepatocytes.
Toxicon, 24(2): 109-115.
Runnegar, M. T. C., Falconer, I. R., and Silver, J. 1981. Deformation of isolated rat hepatocytes by a
peptide hepatoxin from the blue-green alga Microcystis aeruginosa. Naunyn-Schmiedeberg's
Archives of Pharmacology, 317(3): 268-272.
Runnegar, M. T. C., Falconer, I. R., Buckley, T., and Jackson, A. R. B. 1986. Lethal potency and tissue
distribution of 125I-labelled toxic peptides from the blue-green alga Microcystis aeruginosa.
Toxicon, 24(5): 506-509.
Runnegar, M. T. C., Andrews, J., Gerdes, R. G., and Falconer, I. R. 1987. Injury to hepatocytes induced
by a peptide toxin from the cyanobacterium Microcystis aeruginosa. Toxicon, 25(11): 1235-1239.
Runnegar, M. T. C., Berndt, N., and Kaplowitz, N. 1991. Identification of hepatic protein phosphatases as
novel critical targets for the hepatotoxicity of microcystin in vivo. Hepatology, 14(4): A159.
Runnegar, M. T. C., Berndt, N., Kong, S. M., et al. 1995a. In vivo and in vitro binding of microcystin to
protein phosphatases 1 and 2A. Biochemical and Biophysical Research Communications, 216(1):
162-169.
Runnegar, M. T. C., Berndt, N., and Kaplowitz, N. 1995b. Microcystin uptake and inhibition of protein
phosphatases: effects of chemoprotectants and self-inhibition in relation to known hepatic
transporters. Toxicology and Applied Pharmacology, 134(2): 264-272.
Saker, M. L., Jungblut, A.-D. Neilan, B. A., et al. 2005. Detection of microcystin synthase genes in health
food supplements containing the freshwater cyanobacterium Aphanizomenon flos-aquae. Toxicon,
46: 555-562.
Sarma, T. A. 2013. Cyanobacterial Toxins. In: Handbook of Cyanobacteria. CRC Press, Taylor and
Francis Group, pp. 487-606.
Schaeffer, D. J., Malpas, P. B., and Barton, L. L. 1999. Risk assessment of microcystin in dietary
Aphanizomenon flos-aquae. Ecotoxicology and Environmental Safety, 44(1): 73-80.
Scheffer, M., Rinaldi, S., Gragnani, A., et al. 1997. On the dominance of filamentous cyanobacteria in
shallow turbid lakes. Ecology, 78: 272-282.
Health Effects Support Document for Microcystins - June, 2015
115
-------
Schindler, D. W., Hecky, R. E., Findlay, D. L., et al. 2008. Eutrophication of Lakes Cannot be Controlled
by Reducing Nitrogen Input: Results of a 37-Year Whole-Ecosystem Experiment. Proceedings of
the National Academy of Sciences of the United States of America 105: 11254-11258.
Scott, T., McCarthy, M., Otten, T., et al. 2013. Comment: An alternative interpretation of the relationship
between TN:TP and microcystins in Canadian lakes. Canadian Journal of Fisheries and Aquatic
Sciences, 70: 1265-1268
Sedan, D., Andrinolo, D., Telese, L., et al. 2010. Alteration and recovery of the antioxidant system
induced by sub-chronic exposure to microcystin-LR in mice: its relation to liver lipid
composition. Toxicon, 55(2-3): 333-342.
Sekijima, M., Tsutsumi, T., Yoshida, T., et al. 1999. Enhancement of glutathione S-transferase placental-
form positive liver cell foci development by microcystin-LR in aflatoxin B1 -initiated rats.
Carcinogenesis, 20(1): 161-165.
Shapiro, J. 1984. Blue-green dominance in lakes: the role and management significance of pH and CO2.
Internationale Revue der Gesamten Hydrobiologie, 69: 765-780.
Shen, X., Shaw, G. R., Codd, G. A., et al. 2003. DNA microarray analysis of gene expression in mice
treated with the cyanobacterial toxin, cylindrospermopsin. In: S. S. Bates, (Ed.)Proceedings of the
Eight Canadian Workshop on Harmful Marine Algae. Fisheries and Oceans Canada, Monkton,
New Brunswick, pp. 49-51. Available at:
http://www.glf.dfo-mpo.gc.ca/sci-sci/cwhma-atcamn/8th_cwhma_proceedings.pdf.
Shi, Q., Cui, J., Zhang, J., et al. 2004. Expression modulation of multiple cytokines in vivo by
cyanobacteria blooms extract from Taihu Lake, China. Toxicon, 44(8): 871-879.
Shi, Y., Guo, C., Sun, Y., et al. 2011. Interaction between DNA and microcystin-LR studied by spectra
analysis and atomic force microscopy. Biomacromolecules, 12(3): 797-803.
Shirai, M., Takamura, Y., Sakuma, H., et al. 1986. Toxicity and delayed type hypersensitivity caused by
Microcystis blooms from Lake Kasumigaura. Microbiology and Immunology, 30(7): 731-735.
Sicinska, P., Bukowska, B., Michalowicz, J., and Duda, W. 2006. Damage of cell membrane and
antioxidative system in human erythrocytes incubated with microcystin-LR in vitro. Toxicon,
47(4): 387-397.
Sim, A. T. R. and Mudge, L. M. 1993. Protein phosphatase activity in cyanobacteria: consequences for
microcystin toxicity analysis. Toxicon, 31(9): 1179-1186.
Sivonen, K. 1990. Effects of light, temperature, nitrate, orthophosphate, and bacteria on growth of and
hepatotoxin production by Oscillatoria agardhii strains. Applied and Environmental
Microbiology, 56: 2658-2666.
Skulberg, O. M. 1996. Terrestrial and limnic algae and cyanobacteria. In: A. Elvebakk and P. Prestrud
(Eds.) A Catalogue of SvalbardPlants, Fungi, Algae and Cyanobacteria. Part 9, Norsk
Polarinstitutt Skrifter 198: 383-395. (Cited in WHO 1999)
Slatkin, D. N., Stoner, R. D., Adams, W. H, et al. 1983. Atypical pulmonary thrombosis caused by atoxic
cyanobacterial peptide. Science, 220: 1383-1385.
Smith, V. H. 1983. Low nitrogen to phosphorus ratios favor dominance by blue-green algae in lake
phytoplankton. Science, 221(4611): 669-671.
Health Effects Support Document for Microcystins - June, 2015
116
-------
Smith, V. H. 1986. Light and nutrient effects on the relative biomass of blue-green algae in lake
phytoplankton. Canadian Journal of Fisheries and Aquatic Sciences, 43: 148-153.
Soares, R. M., Magalhaes,V. F., and Azevedo, S. M. 2004. Accumulation and depuration of microcystins
(cyanobacteria hepatotoxins) in Tilapia rendalli (Cichlidae) under laboratory conditions. Aquatic
Toxicology, 70: 1-10.
Soares, R. M. Yuan, M., Servaites, C., et al. 2005. Sublethal exposure from microcystins to renal
insufficiency patients in Rio de Janeiro, Brazil. Environmental Toxicology, 21(2): 95-103.
Soares, R. M., Cagido, V., Ferraro, R. B., et al. 2007. Effects of microcystin-LR on mouse lungs.
Toxicon, 50(3): 330-338.
Sondergaard. M., Jenson, J. P., and Jeppesen, E. 2003. Role of sediment and internal loading of
phosphorus in shallow lakes. Hydrobiologia, 506: 135-145.
Song, L., Chen, W., Peng, L., et al. 2009. Distribution and bioaccumulation of microcystins in water
columns: A systematic investigation into the environmental fate and the risks associated with
microcystins in Meiliang Bay, Lake Taihu. Water Research, 41(13): 2853-2864.
Stewart, I., Schluter, P., and Shaw, G. 2006a. Cyanobacterial lipopolysaccharides and human health - a
review. Environmental Health, 5: 7.
Stewart, I., Robertson, I. M., Webb, P. M., et al. 2006b. Cutaneous hypersensitivity reactions to
freshwater cyanobacteria - human volunteer studies. BMC Dermatology, 6:6.
Stoner, R. D., Adams, W. H., Slatkin, D. N., and Siegelman, H. W. 1989. The effects of single L-amino
acid substitutions on the lethal potencies of microcystins. Toxicon, 27(7): 825-828.
Stotts, R. R., Twardock, A. R., Haschek, W. M., et al. 1997a. Distribution of tritiated dihydromicrocystin
in swine. Toxicon, 35(6): 937-953.
Stotts, R. R., Twardock, A. R., Koritz, G. D., et al. 1997b. Toxicokinetics of tritiated dihydromicrocystin-
LRin swine. Toxicon, 35(3): 455-465.
Stotts, R. R., Namikoshi, M., Haschek, W. M., et al. 1993. Structural modifications imparting reduced
toxicity in microcystins from Microcystis spp. Toxicon, 31(6): 783-789.
Sun, Y., Meng, G.-M., Guo, Z.-I., and Xu, L.-H. 2011. Regulation of heat shock protein 27
phosphorylation during microcystin-LR-induced cytoskeletal reorganization in a human liver cell
line. Toxicology Letters, 207(3): 270-277.
Suput, D., Zorc-Pleskovic, R., Petrovic, D., and Milutinovic, A. 2010. Cardiotoxic Injury Caused by
Chronic Administration of Microcystin-YR. Folia Biologica (Prague), 56(1): 14-18.
Suzuki, H., Watanabe, M. F., Yu, Y. P., et al. 1998. Mutagenicity of microcystin-LR in human RSA cells.
International Journal of Molecular Medicine, 2(1): 109-112.
Svoboda, M., Riha, J., Wlcek, K., et al. 2011. Organic Anion Transporting Polypeptides (OATPs):
regulation of expression and function. Current Drug Metabolism, 12(2): 139-153.
Svrcek, C. and D. Smith. 2004. Cyanobacteria toxins and the current state of knowledge on water
treatment options: a review. Journal of Environmental Engineering and Science, 3: 155-185.
Health Effects Support Document for Microcystins - June, 2015
117
-------
Takahashi, O., Oishi, S., and Watanabe, M. F. 1995. Defective blood coagulation is not causative of
hepatic haemorrhage induced by microcystin-LR. Pharmacology & Toxicology, 76(4): 250-254.
Takenaka, S. 2001. Covalent glutathione conjugation to cyanobacterial hepatotoxin microcystin-LR by
F344 rat cytosolic and microsomal glutathione S-transferases. Environmental Toxicology and
Pharmacology, 9(4): 135-139.
Takumi, S., Komatsu, M., Furukawa, T., et al. 2010. p53 Plays an important role in cell fate
determination after exposure to microcystin-LR. Environmental Health Perspectives, 118(9):
1292-1298.
Teixeira-de Mello, F., Meerhoff, M., Pekcan-Hekim,, Z., and Jeppesen, E. 2009. Substantial differences
in littoral fish community structure and dynamics in subtropical and temperate shallow lakes.
Freshwater Biology, 54: 1202-1215.
Tencalla, F., and Dietrich, D. 1997. Biochemical characterization of microcystin toxicity in rainbow trout
(Oncorhynchus mykiss). Toxicon 35; 583-595.
Teneva, I., Mladenov, R., Popov, N., and Dzhambazov. B. 2005. Cytotoxicity and apoptotic effects of
microcystin-LR and anatoxin-a in mouse lymphocytes. Folia Biologica (Prague), 51(3): 62-67.
Thiel, P. 1994. The South African contribution to studies on the toxic cyanobacteria and their toxins. In:
D. A. Steffensen and B. C. Nicholson, (Eds.) Toxic Cyanobacteria: Current Status of Research
and Management. Proceedings of an International Workshop. Adelaide, Australia, March 22-26.
Australian Centre for Water Quality Research, Salisbury, Australia, pp. 23-27.
Thompson, W. L., Allen, M. B. and Bostian, K. A. 1988. The effects of microcystin on monolayers of
primary rat hepatocytes. Toxicon. 26(1): 44.
Thomspon, W. L. and Pace, J. G. 1992. Substances that protect cultured hepatocytes from the toxic
effects of microcystin-LR. Toxicology in Vitro, 6(6): 579-587.
Toivola, D., Eriksson, J. E., and Brautigan, D. L. 1994. Identification of protein phosphatase 2A as the
primary target for microcystin-LR in rat liver homogenates. FEBSLetters, 344(2-3): 175-180.
Toivola, D. M., Goldman, R. D., Garrod, D. R. and Eriksson, J. E. 1997. Protein phosphatases maintain
the organization and structural interactions of hepatic keratin intermediate filaments. Journal of
Cell Science, 110(Pt. 1): 23-33.
Toivola, D., Omary, M., Ku, N.-O., et al. 1998. Protein phosphatase inhibition in normal and keratin 8/18
assembly-incompetent mouse strains supports a functional role of keratin intermediate filaments
in preserving hepatocyte integrity. Hepatology. 28: 116-128.
Torokne, A., Palovics, A., and Bankine, M. 2001. Allergenic (sensitization, skin and eye irritation) effects
of freshwater cyanobacteria - experimental evidence. Environmental Toxicology, 16: 512-216.
Touchette, B. W., Burkholder, J. M., Allen, E. H., et al. 2007. Eutrophication and cyanobacteria blooms
in run-of-river impoundments in North Carolina, U.S.A. Lake and Reservoir Management, 23:
179-192.
Towner, R. A., Sturgeon, S. A., and Hore, K E. 2002. Assessment of in vivo oxidative lipid metabolism
following acute microcystin-LR-induced hepatotoxicity in rats. Free Radical Research, 36(1):
63-71.
Health Effects Support Document for Microcystins - June, 2015
118
-------
Toxicology Literature Online (TOXLINE) 2012. Toxicology Data Network, National Institute of Health.
Retrieved on September 25, 2012 from the World Wide Web:
http: //toxnet.nlm .nih .gov/cgi -bin/sis/htmlgen?T OXLINE
Tsuji, K., Naito, S., Kondo, F., et al. 1993. Stability of microcystins from cyanobacteria: Effect of light on
decomposition and isomerization. Environmental Science & Technology, 28: 173-177. (Cited in
WHO 1999)
Tsuji, K., Naito, S., Kondo, F., et al. 1995. Stability of microcystins from cyanobacteria~II. Effect of UV
light on decomposition and isomerization. Toxicon, 33(12): p. 1619-31.
Turner, P. C., Gammie, A. J., Hollinrake, K., and Codd, G.A. 1990. Pneumonia associated with contact
with cyanobacteria. British Medical Journal, 300(6737): 1440-1441.
U.S. EPA (United States Environmental Protection Agency). 1986a. Guidelines for the Health Risk
Assessment of Chemical Mixtures. Fed. Reg. 51( 185):34014-34025. Available from:
http://cfpub.epa.gov/ncea/cfm/recordisplay.cfm?deid=22567
U.S. EPA (United States Environmental Protection Agency). 1986b. Guidelines for Mutagenicity Risk
Assessment. Fed. Reg. 51(185):34006-34012. Available from:
http://cfpub.epa.gov/ncea/cfm/recordisplay.cfm?deid=23160
U.S. EPA (United States Environmental Protection Agency). 1988. Recommendations for and
documentation of Biological Values for Use in Risk Assessment. EPA 600/6-87/008. Available
from: National Technical Information Service, Springfield, VA; PB88-179874/AS. Available
from: http://cfpub.epa.gov/ncea/cfm/recordisplay.cfm?deid=34855
U.S. EPA (United States Environmental Protection Agency). 1991. Guidelines for Developmental
Toxicity Risk Assessment. Fed. Reg. 56(234):63798-63826. Available from:
http://cfpub.epa.gov/ncea/cfm/recordisplay.cfm?deid=23162
U.S. EPA (United States Environmental Protection Agency). 1994a. Interim policy for particle size and
limit concentration issues in inhalation toxicity studies. Fed. Reg. 59(206):53799. Available from:
http://cfpub.epa.gov/ncea/cfm/recordisplay.cfm?deid=l 86068
U.S. EPA (United States Environmental Protection Agency). 1994b. Methods for derivation of inhalation
reference concentrations and application of inhalation dosimetry. EPA/600/8-90/066F. Available
from: National Technical Information Service, Springfield, VA; PB2000-500023, and
http://www.epa.gov/iris/backgrd.html
U.S. EPA (United States Environmental Protection Agency). 1995. Use of the benchmark dose approach
in health risk assessment. U.S. Environmental Protection Agency. EPA/630/R-94/007. Available
from: National Technical Information Service, Springfield, VA; PB95-213765, and
http://www.epa.gov/iris/backgrd.html
U.S. EPA (United States Environmental Protection Agency). 1996. Guidelines for reproductive toxicity
risk assessment. Fed. Reg. 61(212):56274-56322. Available from:
http://www.epa.gov/iris/backgrd.html
U.S. EPA (United States Environmental Protection Agency). 1998. Guidelines for neurotoxicity risk
assessment. Fed Reg 63(93):26926-26954. Available from: http://www.epa.gov/iris/backgrd.html
Health Effects Support Document for Microcystins - June, 2015
119
-------
U.S. EPA (United States Environmental Protection Agency). 2000a. Science Policy Council Handbook:
peer review. 2nd edition. Prepared by the Office of Science Policy, Office of Research and
Development, Washington, DC. EPA 100-B-00-001. Available from:
http: //www. epa.gov/iris/backgrd .html.
U S. EPA (United States Environmental Protection Agency). 2000b. Supplemental guidance for
conducting for health risk assessment of chemical mixtures. EPA/630/R-00/002. Available from:
http: //www. epa.gov/iris/backgrd .html.
U.S. EPA (United States Environmental Protection Agency). 2002. A review of the reference dose and
reference concentration processes. Risk Assessment Forum, Washington, DC; EPA/630/P-
02/0002F. Available from: http://www.epa.gov/iris/backgrd.html
U.S. EPA (United States Environmental Protection Agency). 2005a. Guidelines for carcinogen risk
assessment. Risk Assessment Forum, Washington, DC; EPA/630/P-03/00IB. Available from:
http: //www. epa.gov/iris/backgrd .html
U.S. EPA (United States Environmental Protection Agency). 2005b. Supplemental guidance for assessing
susceptibility from early-life exposure to carcinogens. Risk Assessment Forum, Washington, DC;
EPA/630/R-03/003F. Available from: http://www.epa.gov/iris/backgrd.html
U.S. EPA (United States Environmental Protection Agency). 2006a. Science Policy Council Handbook:
Peer Review. 3rd edition. Prepared by the Office of Science Policy, Office of Research and
Development, Washington, DC. EPA 100-B-06-002. Available from:
http://www.epa.gov/peerreview/pdfs/peer_review_handbook_2012.pdf
U.S. EPA (United States Environmental Protection Agency). 2006b. A framework for assessing health
risks of environmental exposures to children. National Center for Environmental Assessment,
Washington, DC; EPA/600/R-05/093F. Available from: http://www.epa.gov/iris/backgrd.html
U.S. EPA (United States Environmental Protection Agency). 2009. National Lakes Assessment. A
Collaborative Survey of the Nation's Lakes. EPA 841-R-09-001. Available from:
http://www.epa.gov/owow/LAKES/lakessurvey/pdf/nla_report_low_res.pdf
U.S. EPA (United States Environmental Protection Agency). 201 la. Exposure Factors Handbook 2011
Edition (Final). Washington, DC, EPA/600/R-09/052F. Available from:
http://www.epa.gov/ncea/efh/pdfs/efh-complete.pdf
U.S. EPA (United States Environmental Protection Agency). 2012. Benchmark dose technical guidance
document [external review draft]. EPA/630/R-00/001. Available from:
http: //www. epa.gov/iris/backgrd .html
U.S. EPA. (United States Environmental Protection Agency). 2014a. Child-Specific Exposure Scenarios
Examples (Final Report), Washington, DC, EPA/600/R-14-217F. Available from:
http://cfpub.epa.gov/ncea/risk/recordisplay.cfm?deid=262211#Download
U.S. EPA (United States Environmental Protection Agency). 2014b. Framework for Human Health Risk
Assessment to Inform Decision Making. Office of the Science Advisor, Risk Assessment Forum,
Washington, DC; EPA/100/R-14/001. Available from:
http://www2.epa.gov/programs-office-science-advisor-osa/framework-human-health-risk-
assessment-inform-decision-making
Health Effects Support Document for Microcystins - June, 2015
120
-------
Ueno, Y., Nagata, S., Tsutsumi, T. et al. 1996. Detection of microcystins, a blue-green algal hepatotoxin,
in drinking water sampled in Haimen and Fusui, endemic areas of primary liver cancer in China,
by highly sensitive immunoassay. Carcinogenesis, 17(6): 1317-1321.
Ueno, Y., Makita, Y., Nagata, S. et al. 1999. No chronic oral toxicity of a low-dose of microcystin-LR, a
cyanobacterial hepatoxin, in female Balb/C mice. Environmental Toxicology, 14(1): 45-55.
VanderKooi, S., Burdinck, S., Echols, K., et al. Algal toxin in Upper Klamath Lake, Oregon: Linking
water quality to juvenile sucker health. U.S. Geological Survey Fact Sheet 2009-3111, pp. 2.
Vareli, K., Zarali, E., Zacharioudakis, G. S. A., et al. 2012. Microcystin producing cyanobacterial
communities in Amvrakikos Gulf (Mediterranean Sea, NW Greece) and toxin accumulation in
mussels (Mytilus galloprovincialis). Harmful Algae, 15: 109-118.
Vasconcelos, V., Oliveira, S., and Teles, F. O. 2001. Impact of a toxic and a non-toxic strain of
Microcystis aeruginosa on the crayfish Procambarus clarkii. Toxicon, 39: 1461-1470.
Vesterkvist, P.S. and Meriluoto, J. A. 2003. Interaction between microcystins of different
hydrophobicities and lipid monolayers. Toxicon, 41(3): 349-355.
Wagner, C. and Adrian, R. 2009. Cyanobacteria dominance: quantifying the effects of climate change.
Limnology and Oceanography, 54: 2460-2468.
Wang, X., Parkpian, P., Fujimoto, N., et al. 2002. Environmental conditions associated with microcystins
production to Microcystis aeruginosa in a reservoir of Thailand. Journal of Environmental
Science and Health, Part A, 37: 1181-1207.
Wang, C., Kong, H., Wang, X., et al. 2010. Effects of iron on growth and intracellular chemical contents
of Microcystin aeruginosa. Biomedical and Environmental Sciences, 23: 48-52.
Wang, X., Ying, F., Chen, Y., and Han, X. 2012. Microcystin (-LR) affects hormones level of male mice
by damaging hypothalamic-pituitary system. Toxicon, 59(2): 205-214.
Wang, X., Chen, Y., Zuo, X., et al. 2013. Microcystin (-LR) induced testicular cell apoptosis via up-
regulating apoptosis-related genes in vivo. Food and Chemical Toxicology, 60: 309-317.
Watanabe M. F. and Oishi, S. 1985. Effects of environmental factors on toxicity of a cyanobacterium
Microcystis aeruginosa under culture conditions. Applied and Environmental Microbiology, 49:
1342-1344
Watanabe, M. M., Kaya, K., and Takamura, N. 1992. Fate of the toxic cyclic heptapeptides, the
microcystins, from blooms of Microcystis (cyanobacteria) in a hypertrophic lake. Journal of
Phycology, 28: 761-767. (Cited in WHO 1999)
Watanabe, M. F., Park, H-D., Kondo, F., et al. 1997. Identification and estimation of microcystins in
freshwater mussels. Natural Toxins, 5: 31-35. (Cited in WHO 1999)
Wei, Y., Weng, D., Li, F., et al. 2008. Involvement of JNK regulation in oxidative stress-mediated murine
liver injury by microcystin-LR. Apoptosis, 13(8): 1031-1042.
Weng, D., Lu, Y., Wei, Y., et al. 2007. The role of ROS in microcystin-LR-induced hepatocyte apoptosis
and liver injury in mice. Toxicology, 232(1-2): 15-23.
Health Effects Support Document for Microcystins - June, 2015
121
-------
Weyhenmeyer, G.A. 2001. Warmer winters: are planktonic algal populations in Sweden's largest lakes
affected? Ambio, 30: 565-571.
WHO (World Health Organization). 1999. Toxic Cyanobacteria in Water: A Guide to their Public Health
Consequences, Monitoring, and Management, I. Chorus and J. Bartram, (Eds.), E&FN Spon,
London, UK.
WHO (World Health Organization). 2003. Cyanobacterial toxins: Microcystin-LR in Drinking-water.
Background document for development of WHO Guidelines for Drinking-water Quality. World
Health Organization, Geneva, Switzerland.
Wickstrom, M. L., Khan, S. A., Haschek, W.M., et al. 1995. Alterations in microtubules, intermediate
filaments and microfilaments induced by microcystin-LR in cultured cells. Toxicologic
Pathology, 23(3): 326-337.
Williams, D. E., Craig, M., Dawe, S. C., et al. 1997. Evidence for a covalently 'bound form of
microcystin-LR in salmon larvae and dungeness crab larvae. Chemical Research in Toxicology,
10: 463-469. (Cited in WHO 1999)
Wilson, A.E., Gossiaux, D. C., HooK, T. O., et al. 2008. Evaluation of the human health threat associated
with the hepatotoxin microcystin in the muscle and liver tissues of yellow perch (Perca
flavescens). Canadian Journal of Fisheries and Aquatic Sciences, 65: 1487-1497.
Wolf, H.-U. and Frank, C. 2002. Toxicity assessment of cyanobacterial toxin mixtures. Environmental
Toxicology, 17(4): 395-399.
Wolf, D. C., & Mann, P. C. (2005). Confounders in interpreting pathology for safety and risk assessment.
Toxicology and applied pharmacology, 202(3), 302-308.
Wood, S. A., Briggs, L. R., Sprosen, J., et al. 2006. Changes in concentrations of microcystins in rainbow
trout, freshwater mussels, and cyanobacteria in Lakes Rotoiti and Rotoehu. Environmental
Toxicology, 21: 205-222.
WSDE (Washington State Department of Ecology). 2012. Freshwater Algae Control Program. Accessed
December 12, 2012; http://www.ecy.wa.gov/programs/wq/plants/algae/index.html.
Wu J. Y., Xu, Q. J., Gao, G., and Shen, J. H. 2006. Evaluating genotoxicity associated with microcystin-
LR and its risk to source water safety in Meiliang Bay, Taihu Lake. Environmental Toxicology,
21(3): 250-255.
Wu, J., Shao, S., Zhou, F., et al. 2014. Reproductive toxicity on female mice induced by microcystin-LR.
Environmental Toxicology and Pharmacology, 37: 1-6.
Wynne, T., Stumpf, R., Tomlinson, M., et al. 2013. Evolution of a cyanobacterial bloom forecast system
in western Lake Erie: Development and initial evaluation. Journal of Great Lakes Research,
39(Supplement 1): 90-99.
Xie, L., Xie, P., Guo, L., et al. 2005. Organ distribution and bioaccumulation of microcystins in
freshwater fish at different trophic levels from the eutrophic Lake Chaohu, China. Environmental
Toxicology, 20: 293-300.
Xing, M. L., Wang, X. F., and Xu, L. H. 2008. Alteration of proteins expression in apoptotic FL cells
induced by MCLR. Environmental Toxicology, 23(4): 451-458.
Health Effects Support Document for Microcystins - June, 2015
122
-------
Xiong X., Zhong A., Xu H. 2014. Effect of Cyanotoxins on the Hypothalamic-Pituitary-Gonadal Axis in
Male Adult Mouse. PloSone, 9(11): el06585.
Xu, L., Lam, P. K. S., Chen, J., et al. 2000. Comparative study on in vitro inhibition of grass carp
(Ctenopharyngodon idellus) and mouse protein phosphatases by microcystins. Environmental
Toxicology, 15(2) :71-75.
Xu, L., Qin, W., Zhang, H., et al. 2012. Alterations in microRNA expression linked to microcystin-LR-
induced tumorigenicity in human WRL-68 Cells. Mutation Research, 743: 75-82.
Yea, S. S., Yang, Y, I., Jang, W. H., and Paik, K. H. 2001. Microcystin-induced proinflammatory
cytokines expression and cell death in human hepatocytes. Hepatology, 34(4 Pt. 2): 516A
Yoshida, T., Makita, Y., Nagata, S., et al. 1997. Acute oral toxicity of microcystin-LR, a cyanobacterial
hepatotoxin in mice. Nat. Toxins, 5: 91-95.
Yoshizawa, S., Matsushima, R., Watanabe, M. F. et al. 1990. Inhibition of protein phosphatases by
Microcystis and nodularin associated with hepatotoxicity. Journal of Cancer Research and
Clinical Oncology, 116(6): 609-614.
Yu, S. Z., Huang, X. E., Koide, T., et al. 2002. Hepatitis B and C viruses infection, lifestyle and genetic
polymorphisms as risk factors for hepatocellular carcinoma in Haimen, China. Japanese Journal
of Cancer Research, 93(12): 1287-1292.
Yuan, G., Xie, P., Zhang, X., et al. 2012. In vivo studies on the immunotoxic effects of microcystins on
rabbit. Environmental Toxicology, 27(2): 83-89.
Yuan, L. L., Pollard, A., Pather, S., et al. 2014. Managing microcystin: identifying national-scale
thresholds for total nitrogen and chlorophyll a. Freshwater Biology, 59: 1970-1981.
Yuan, M., Namikoshi, M., Otsuki, A., et a.l. 1999. Electrospray Ionization Mass Spectrometric Analysis
of Microcystins, Cyclic Heptapeptide Hepatotoxins: Modulation of Charge States and [M 1 H] 1
to [M lNa]l Ratio. Journal ofthe American Society for Mass Spectrometry, 10: 1138-1151.
Yuan, M., Carmichael, W. W., and Hilborn, E. D. 2006. Microcystin analysis in human sera and liver
from human fatalities in Caruaru, Brazil 1996. Toxicon, 48(6): 627-640.
Zegura, B., Sedmak, B., and Filipic, M. 2003. Microcystin-LR induces oxidative DNA damage in human
hepatoma cell line HepG2. Toxicon, 41(1): 41-48.
Zegura, B., Lah, T. T., and Filipic, M. 2004. The role of reactive oxygen species in microcystin-LR-
induced DNA damage. Toxicology, 200(1): 59-68.
Zegura, B., Lah, T. T., and Filipic, M. 2006. Alteration of intracellular GSH levels and its role in
microcystin-LR-induced DNA damage in human hepatoma HepG2 cells. Mutation Research,
611(1-2): 25-33.
Zegura, B., Zajc, I., Lah, T. T., and Filipic, M. 2008a. Patterns of microcystin-LR induced alteration of
the expression of genes involved in response to DNA damage and apoptosis. Toxicon, 51(4): 615-
623.
Zegura, B., Volcic, M., Lah, T. T., and Filipic, M. 2008b. Different sensitivities of human colon
adenocarcinoma (CaCo-2), astrocytoma (IPDDC-A2) and lymphoblastoid (NCNC) cell lines to
microcystin-LR induced reactive oxygen species and DNA damage. Toxicon, 52(3): 518-525.
Health Effects Support Document for Microcystins - June, 2015
123
-------
Zegura, B., Straser, A., and Filipic, M. 2011. Genotoxicity and potential carcinogenicity of cyanobacterial
toxins - a review. Mutation Research, 727: 16-41.
Zeller, P., Clement, M., and Fessard, V. 2011. Similar uptake profiles for microcystin -LR and -RR in an
in vitro human intestinal model. Toxicology, 290: 7-13
Zhan, L., Sakamoto, M., Sakuraba, M. et al. 2004. Genotoxicity of microcystin-LR in human
lymphoblastoid TK6 cells. Mutation Research, 557(1): 1-6.
Zhang, X. X., Zhang, Z., Fu, Z., et al. 2010. Stimulation effect of microcystin-LR on matrix
metalloproteinase-2/-9 expression in mouse liver. Toxicology Letters, 199(3): 377-382.
Zhang, X., Xie, P., Li., D., et al. 201 la. Anemia Induced by Repeated Exposure to Cyanobacterial
Extracts with Explorations of Underlying Mechanisms. Environmental Toxicology, 26(5): 472-
479.
Zhang, H. Z., Zhang, F. Q., Li, C. F., et al. 2011b. A cyanobacterial toxin, microcystin-LR, induces
apoptosis of Sertoli cells by changing the expression levels of apoptosis-related proteins. The
Tohoku Journal of Experimental Medicine, 224(3): 235-42.
Zhang, Z., Zhang, X. X., Qin, W., et al. 2012. Effects of microcystin-LR exposure on matrix
metalloproteinase-2/-9 expression and cancer cell migration. Ecotoxicology and Environmental
Safety, 77: 88-93.
Zhao, J. M. and Zhu, H. G. 2003. Effects of microcystins on cell cycle and expressions of c-fos and c-jun.
Zhonghua Yu Fang Yi Xue Za Zhi, 37(1): 23-25. (Chinese)
Zhao Y., Xue Q., Su X., et al. 2015. Microcystin-LR induced thyroid dysfunction and metabolic disorders
in mice. Toxicology, 328: 135-141.
Zhou, L., Yu, H., and Chen, K. 2002. Relationship between microcystin in drinking water and colorectal
cancer. Biomedical and Environmental Sciences, 15(2): 166-171.
Zhou, Y., Yuan, J., Wu, J., and Han, X. 2012. The toxic effects of microcystin-LR of rat spermatogonia in
vitro. Toxicology Letters, 212: 48-56.
Zhu, Y., Zhong, X., Zheng, S., et al. 2005. Transformation of immortalized colorectal crypt cells by
microcystin involving constitutive activation of Akt and MAPK cascade. Carcinogenesis, 26(7):
1207-1214.
Zimba, P. V., Camus. A., Allen, E. H., and Burkholder, A. M.. (2006). Co-occurrence of white shrimp,
Litopenaeus vannamei, mortalities and microcystin toxin in a southeastern USA shrimp facility.
Aquaculture, 261: 1048-1055.
Zurawell, R. W., Chen, H., Burke, J. M., and Prepas, E. E. 2005. Hepatotoxic cyanobacteria: a review of
the biological importance of microcystins in freshwater environments. Journal of Toxicology and
Environmental Health, PartB, 8(1): 1-37.
Health Effects Support Document for Microcystins - June, 2015
124
------- |