Ecosystem Health Science-Based Solutions Canada Canadian Water Quality

Ecosystem Health Science-Based Solutions Canada Canadian Water Quality
Ecosystem Health
Science-Based Solutions
Canadian Water Quality
Guidelines for the Protection of
Aquatic Life: Nitrate Ion
Report No. 1-6
Environment
Canada
Environnement
Canada
Canada
Science-Based Solutions
Solutions fondées sur la science
Prepared and published by
National Guidelines and Standards Office
Water Policy and Coordination Directorate
Environment Canada
Ottawa
May, 2003
ISSN 1497-2689
ISBN 0-662-34079-5
Cat. No. En1-34/6-2003E-IN
Scientific Supporting Document
Canadian Water Quality Guidelines
for the Protection of Aquatic Life:
Nitrate Ion
Report No. 1-6
© Her majesty the Queen in Right of Canada, represented by the Minister of the Environment, 2003.
All rights reserved. Reproduction authorized if source is acknowledged. The reproduction must be
presented within its proper context and must not be used for profit.
NOTE TO READERS
The Ecosystem Health: Science-based Solutions series is dedicated to the
dissemination of scientific knowledge, information and tools for monitoring,
assessing, and reporting on ecosystem health to support Canadians in making sound
decisions. Documents published in this series include the scientific basis, methods,
approaches and frameworks for environmental guidelines and their implementation;
monitoring, assessing, and rehabilitating environmental quality in Canada; and,
indicator development, environmental reporting and data management. Issues in this
series are published ad libitum.
This particular issue provides a general overview of the current understanding of
nitrate ion in the Canadian aquatic environment and is based largely on the scientific
supporting documents to the Canadian Environment Quality Guidelines (sediment,
tissue, soil) for the nitrate ion. For additional information regarding this document,
please contact:
Environment Canada
Water Policy and Coordination Directorate
National Guidelines and Standards Office
351 Saint-Joseph Boulevard
Gatineau, QC K1A 0H3
phone: 819-953-1550
fax: 819-953-0461
[email protected]
http://www.ec.gc.ca/ceqg-rcqe
This scientific document is available in English only. Ce document scientifique n’est
disponible qu’en anglais avec un résumé en français.
Reference listing:
Environment Canada. 2003. Canadian Water Quality Guidelines for the Protection of
Aquatic Life: Nitrate Ion. Ecosystem Health: Science-based Solutions Report No. 1-6.
National Guidelines and Standards Office, Water Policy and Coordination
Directorate, Environment Canada. 115 pp.
Science-based Solutions No. 1-6
i
ACKNOWLEDGEMENTS
This document was prepared under the direction of the National Guidelines and
Standards Office (NGSO) of Environment Canada. The NGSO thanks J.D. Whall,
Kelly Potter, and Elizabeth Roberts for their major scientific contributions to this
document. Many supporting contributors provided valuable scientific expertise in the
development of the document, including: Pierre-Yves Caux, Uwe Schneider, Susan
Roe, Sushil Dixit and Paul Jiapizian.
The NGSO gratefully acknowledges the following reviewers for their expert input to
the scientific supporting document: Patricia Chambers (EC), Peter Champman (EVS
Environment Consultants), Tim Fletcher (OMOE), Michael Goss (U. of Guelph),
Cynthia Grant (AAFC), Gary Grove (EC), Martha Guy (EC), Steve Hecnar (Lakehead
U.), Alison Humphries (Credit Valley Conservation), Neil Hutchinson (Gartner Lee
Ltd.), Robert Morris (Credit Valley Conservation), Alan Nebeker (US EPA), Lee Nikl
(EVS Environment Consultants), Renee Paterson (NF DOE), Peter Strain (DFO),
Shaun Watmough (Trent U.). The NGSO also thanks the following individuals for
providing additional data on nitrate levels in Canadian waters: Harry Alkema (EC),
Peter Dillon (OMOE), Serge L’Italien (EC), Jean Painchaud (MENV), Chul-Un Ro
(EC), Darrell Taylor (NSDEL), Aaron Todd (OMOE), Frank Whitney (DFO), Dwight
Williamson (Manitoba Conservation), Jen Winter (OMOE).
The NGSO extends its appreciation to members of the Canadian Council of Ministers
of the Environment (CCME) Water Quality Guidelines Task Group: Bijan Aidun, Joe
Ballantyne, George Crawford, Sam Ferris, Don Fox, Connie Gaudet, Isabelle Guay,
Francis Jackson, Bryan Levia, Narender Nagpal, Bruce Raymond, Les Swain, Darrell
Taylor, Haseen Khan, and Dwight Williamson.
Science-based Solutions No. 1-6
ii
TABLE OF CONTENTS
NOTE TO READERS..............................................................................................i
ACKNOWLEDGEMENTS .................................................................................... ii
TABLE OF CONTENTS ....................................................................................... iii
LIST OF FIGURES............................................................................................... v
LIST OF TABLES................................................................................................. v
ABSTRACT ......................................................................................................... vi
RÉSUMÉ............................................................................................................ viii
LIST OF ACRONYMS......................................................................................... xi
1. INTRODUCTION...............................................................................................1
2. PHYSICAL AND CHEMICAL PROPERTIES ....................................................2
2.1 Chemistry of the Nitrate Ion .........................................................................2
2.2 Analytical Methods ......................................................................................4
3. NITRATE PRODUCTION AND RELEASE TO THE ENVIRONMENT ..............9
3.1 Nitrogen Cycle .............................................................................................9
3.2 Natural Sources.........................................................................................11
3.3 Anthropogenic Sources .............................................................................11
3.3.1 Municipal Wastewaters .......................................................................12
3.3.2 Industrial Sources ...............................................................................13
3.3.3 Agricultural Sources ............................................................................15
4. ENVIRONMENTAL FATE AND BEHAVIOUR.................................................18
4.1 Atmospheric Processes .............................................................................18
4.1.1 Wet Deposition....................................................................................18
4.1.2 Dry Deposition ....................................................................................19
4.2 Terrestrial Processes.................................................................................19
4.2.1 Adsorption...........................................................................................19
4.2.2 Leaching .............................................................................................19
4.2.3 Water-driven Erosion or Runoff...........................................................23
4.2.4 Biotic Uptake and Assimilation............................................................25
4.2.5 Microbial Transformation.....................................................................26
4.3 Aquatic Processes .....................................................................................26
4.3.1 Physico-chemical Factors and Nitrogen Speciation ............................26
4.3.2 Advective and Diffusional Movement Within a Water Body.................29
4.3.3 Microbial Nitrification ...........................................................................29
4.3.4 Microbial Denitrification .......................................................................31
4.3.5 Biotic Assimilation ...............................................................................32
4.3.6 Movement From Water to Sediments .................................................34
4.3.7 Exchanges Between Surface Waters and Groundwater .....................34
4.3.8 Anthropogenic Nitrate Removal from Ground and Surface Waters.....36
5. ENVIRONMENTAL CONCENTRATIONS.......................................................39
5.1 Nitrate Levels in Precipitation ....................................................................39
5.2 Environmental Levels in Surface Waters ...................................................39
5.2.1 Freshwater ..........................................................................................39
5.2.2 Marine .................................................................................................44
Science-based Solutions No. 1-6
iii
5.3 Environmental Levels in Groundwater .......................................................45
6. TOXICITY OF NITRATE TO AQUATIC ORGANISMS....................................47
6.1 Influence of Various Nitrate Salts on Toxicity ............................................48
6.2 Modes of Action.........................................................................................51
6.2.1 Uptake Mechanisms ...........................................................................51
6.2.2 Direct Toxicity .........................................................................................53
6.2.3 Indirect Toxicity ...................................................................................56
6.3 Toxicity to Freshwater Life.........................................................................60
6.3.1 Algae and Plants.................................................................................60
6.3.2 Invertebrates .......................................................................................61
6.3.3 Fish .....................................................................................................63
6.3.4 Amphibians .........................................................................................65
6.4 Toxicity to Marine Life................................................................................66
6.4.1 Algae and Plants.................................................................................66
6.4.2 Invertebrates .......................................................................................67
6.4.3 Fish .....................................................................................................68
6.5 Genotoxicity of Nitrate ...............................................................................69
6.6 Toxicity to Semi-Aquatic Animals ..............................................................70
7. CANADIAN WATER QUALITY GUIDELINES.................................................71
7.1 Protection of Aquatic Life..........................................................................71
7.2 Freshwater Guideline Derivation ...............................................................71
7.2.1 Recommended Freshwater Guideline.................................................74
7.2.2 Data Gaps / Research Recommendations..........................................75
7.2.3 Summary of Existing Guidelines .........................................................76
7.3 Marine Guideline Development .................................................................77
7.3.1 Recommended Marine Guideline........................................................78
7.3.2 Data Gaps / Research Recommendations..........................................79
7.3.3 Summary of Existing Guidelines .........................................................79
8. GUIDANCE ON APPLICATION OF THE GUIDELINES .................................81
8.1 General Guidance on the Use of Guidelines .............................................81
8.2 Monitoring and Analysis of Nitrate Levels..................................................81
8.3 Developing Site-Specific Objectives ..........................................................82
8.4 Trophic Status Management......................................................................84
9. REFERENCES................................................................................................85
Appendix A. Summary of freshwater toxicity studies.........................................105
Appendix B. Summary of marine toxicity studies. .............................................113
Science-based Solutions No. 1-6
iv
LIST OF FIGURES
Figure 2.1. Chemical structure of the nitrate ion. .................................................2
Figure 3.1. The nitrogen cycle............................................................................10
Figure 4.1. Schematic representation of the nitrogen cycle emphasizing
aquatic transformations. ...................................................................28
Figure 8.1 a) Redox potential (Eh) and electron potential (pE) for various
species of inorganic nitrogen, as a function of pH (note: N2 is
treated as a redox inert compound). b) Generalized vertical
distribution of redox potential and dissolved oxygen in stratified
lakes of very low and very high productivity. ....................................83
LIST OF TABLES
Table 2.1. Summary of selected physical and chemical properties for nitrate
ion and selected nitrate salts..............................................................3
Table 2.2. Conversion factors for various nitrate units to mg NO3-·L-1 ...................4
Table 2.3. Comparison of available techniques for analysis of nitrate in water. ...6
Table 3.1. Nitrogen loading estimates to Canadian surface and ground
waters from various sources, 1996 ..................................................12
Table 4.1. Factors affecting nitrate leaching through agricultural soils...............20
Table 4.2. Selected wastewater treatment processes for nitrate removal. .........38
Table 5.1. Representative nitrate concentrations in Canadian ambient
surface waters..................................................................................41
Table 6.1. Relative toxicity of sodium and potassium nitrate salts to
freshwater organisms.......................................................................49
Table 6.2. Relative toxicity of sodium and potassium chloride salts to
freshwater invertebrates...................................................................50
Table 6.3. Average total nitrogen levels in global lakes, streams and
coastal marine waters of varying trophic status. ..............................59
Science-based Solutions No. 1-6
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ABSTRACT
This scientific supporting document describes the development of Canadian Water
Quality Guidelines for the protection of aquatic life for the nitrate ion. It contains a
review of technical background information on the chemical and physical properties
of nitrate and nitrate salts, a review of sources and releases in Canada, the
distribution and behaviour of nitrate in the environment, and the toxicological effects
of nitrate on freshwater and marine aquatic life. This information is used to derive
ambient water quality guidelines for the nitrate ion, based on direct toxic effects, to
protect ecological receptors in Canadian waters. The role of total nitrogen and
nitrogen-to-phosphorus ratios in causing indirect toxic effects through eutrophication
are discussed in a separate document (CCME 2002). The guidelines in this
document are based on the best available toxicity data at the time of writing, January
2002.
Nitrate occurs naturally in the environment and is constantly produced and consumed
through the processes of the nitrogen cycle. Nitrate is also produced
anthropogenically for uses such as the production of fertilizers, steel, petroleum, pulp
and paper, organic and inorganic chemicals, plastics, nitroaromatic compounds,
nitroorganic compounds in pharmaceuticals, and explosives. Nitrate salts are used in
photography, glass making, engraving, textile dyes, and food processing. The major
anthropogenic sources of nitrate to surface waters are agricultural runoff, municipal
and industrial wastewaters, urban runoff, landfill leachate, precipitation of nitric oxide
and nitrogen dioxide from vehicular exhaust, storm sewer overflow, and septic tanks.
Nitrogen from all sources, and in all its forms, can potentially be transformed into
nitrate. It is estimated that approximately 600 kt of total nitrogen were released to
Canadian surface and groundwaters in 1996 from both natural and anthropogenic
sources.
Ambient nitrate levels in Canadian lakes and rivers are typically less than
4 mg NO3-·L-1. Concentrations less than 0.4 mg NO3-·L-1 are indicative of oligotrophic
lakes and streams. Concentrations exceeding 4 mg NO3-·L-1 are often associated with
eutrophic conditions, and are generally the result of anthropogenic inputs. North
American streams in agricultural landscapes typically have elevated levels of nitrate
due to fertilizer use, with mean nitrate concentrations ranging between 9 and
180 mg NO3-·L-1. Nitrate levels in marine waters are usually lower than in fresh
waters. In Canadian coastal waters, ambient nitrate concentrations rarely exceed
0.5 mg NO3-·L-1, but in estuaries draining agricultural land, nitrate concentrations can
reach 12 mg NO3-·L-1. Levels of nitrate in Canadian groundwater can range from 1 to
1100 mg NO3-·L-1, but in the absence of anthropogenic contamination, levels are
generally less than 13 mg NO3-·L-1.
In water, the fate of nitrate is primarily determined by the biotic processes of
assimilation, nitrogen fixation, nitrification, denitrification, ammonification, and
decomposition of organic matter. Rates of these processes are affected by pH,
temperature, and oxygen availability. Through biotic assimilation, nitrate is taken up
by aquatic plants and algae and is used for the synthesis of cellular materials, such
Science-based Solutions No. 1-6
vi
as proteins. The mode of nitrate uptake from the water by aquatic animals is unclear.
Nitrate’s mode of toxicity to aquatic life is also unclear, though two proposed
mechanisms are: a) through methaemoglobin formation, with a reduction in oxygen
carrying capacity of the blood, and b) through the inability of the organism to maintain
proper osmoregulation under high salt contents associated with elevated nitrate
levels.
Nitrate toxicity tests have been conducted through the addition of nitrate salts such
as sodium nitrate, potassium nitrate, and ammonium nitrate. Evidence suggests that
in tests with ammonium nitrate, toxic effects observed are due to the ammonium ion,
rather than the nitrate. Similarly, in fresh water, the effects of potassium nitrate are
likely due to the potassium. In marine waters, however, toxic levels of potassium
nitrate occur at potassium concentrations below background levels of potassium in
seawater, and therefore the toxicity can be attributed to the nitrate ion. Based on
these arguments, nitrate toxicity to freshwater organisms was only evaluated using
tests with sodium nitrate, while toxicity data for both sodium nitrate and potassium
nitrate were used to evaluate toxicity to marine organisms.
Nitrate has wide-ranging effects in invertebrates, fish and amphibians, with larval
stages generally showing greater sensitivity than adults. Adverse effects observed in
aquatic organisms include: mortality, growth reduction, reduced feeding rates,
reduced fecundity, reduced hatching success, lethargy, behavioural signs of stress,
bent spines and other physical deformities.
An interim water quality guideline of 13 mg NO3-·L-1 is recommended for the protection
of all stages of freshwater life against the adverse effects of the nitrate ion. This
guideline was derived by multiplying the lowest observable adverse effect
concentration of 133 mg NO3-·L-1, reported for growth reduction in the Pacific treefrog
(Pseudacris regilla), by a safety factor of 0.1. An interim water quality guideline of
16 mg NO3-·L-1 is recommended for the protection of marine aquatic life against the
adverse effects of the nitrate ion. This guideline was derived by multiplying the
median lethal concentration of 329 mg NO3-·L-1 for the temperate marine polychaete
(Nereis grubei) by a safety factor of 0.05. A more conservative safety factor was used
for the marine guideline because: the polychaete in the critical study was not tested
at its most sensitive life stage; the critical endpoint, although chronic, was based on a
median lethal effect rather than a low sublethal effect; and adverse effects have been
observed in non-indigenous tropical species exposed to much lower nitrate
concentrations.
These nitrate water quality guidelines are intended to prevent direct toxicity to aquatic
organisms, but will not necessarily prevent eutrophication. Therefore, even at nitrate
concentrations below these guideline levels, indirect toxic effects due to excess algal
growth may still occur.
Science-based Solutions No. 1-6
vii
RÉSUMÉ
Le présent document scientifique complémentaire décrit l’élaboration de
recommandations canadiennes pour la qualité des eaux visant la protection de la vie
aquatique pour l’ion nitrate. Il présente un examen des données techniques de base
sur les propriétés chimiques et physiques de l’ion nitrate et des nitrates ainsi qu’une
revue de leurs sources et de leurs rejets au Canada, indique la distribution et le
comportement des nitrates dans l’environnement et examine leurs effets
toxicologiques sur la vie aquatique d’eau douce et marine. Ces données servent à
élaborer des recommandations pour la qualité des eaux concernant l’ion nitrate en se
fondant sur les effets toxiques directs afin de protéger les récepteurs écologiques
dans les eaux canadiennes. Le rôle joué par l’azote total et les rapports
azote/phosphore dans la production d’effets toxiques indirects par eutrophisation est
discuté dans un autre document (CCME 2002). Les recommandations ici présentées
sont fondées sur les meilleures données sur la toxicité disponibles en janvier 2002,
au moment où le document fut rédigé.
Les nitrates se retrouvent naturellement dans l’environnement. Ils sont constamment
produits et consommés au cours des procédés du cycle de l’azote. Ils peuvent aussi
être d’origine anthropique et servir par exemple à la production d’engrais, d’acier, de
pétrole, de pâtes et papiers, de composés organiques et inorganiques, de matières
plastiques, de composés aromatiques azotés, de composés organiques azotés
utilisés dans les produits pharmaceutiques et d’explosifs. Les nitrates sont utilisés en
photographie, dans la fabrication du verre, en gravure, dans les teintures pour textile
et dans la transformation des aliments. Les principales sources anthropiques des
rejets de nitrates dans les eaux de surface sont le ruissellement agricole, les eaux
usées municipales et industrielles, le ruissellement urbain, la lixiviation des
décharges, les émissions d’oxyde nitrique et de dioxyde d’azote provenant des gaz
d’échappement des véhicules, le débordement des égouts pluviaux et les fosses
septiques. L’azote provenant de toutes les sources et sous toutes ses formes peut
être transformé en nitrate. On estime qu’environ 600 kt d’azote total provenant de
sources à la fois naturelles et anthropiques ont été rejetées en 1996 dans les eaux
de surface et souterraines au Canada.
En général, la teneur ambiante en nitrates des lacs et des cours d’eau canadiens est
inférieure à 4 mg de NO3-·L-1. Des concentrations inférieures à 0,4 mg de NO3-·L-1
indiquent des lacs et des cours d’eau oligotrophes. Des concentrations supérieures à
4 mg de NO3-·L-1 sont souvent associées à des conditions eutrophes et résultent
généralement d’apports anthropiques. Dans les cours d’eau nord-américains en
milieu rural, les concentrations de nitrates tendent à être élevées en raison de
l’utilisation d’engrais, et leur moyenne varie entre 9 et 180 mg de NO3-·L-1. Dans les
eaux marines, les concentrations de nitrates sont ordinairement plus faibles que dans
les eaux douces. Dans les eaux côtières canadiennes, la teneur ambiante en nitrates
dépasse rarement 0,5 mg de NO3-·L-1, mais dans les estuaires qui drainent des
terres agricoles, les concentrations peuvent atteindre 12 mg de NO3-·L-1. Dans les
eaux souterraines du Canada, les concentrations de nitrates peuvent varier de 1 à 1
Science-based Solutions No. 1-6
viii
100 mg de NO3-·L-1, mais en l’absence de contamination anthropique, elles sont
généralement inférieures à 13 mg de NO3-·L-1.
Dans l’eau, le devenir des nitrates est surtout déterminé par les procédés biotiques
d’assimilation, de fixation de l’azote, de nitrification, de dénitrification,
d’ammonification et de décomposition de la matière organique. La vitesse de ces
procédés dépend du pH, de la température et de la disponibilité en oxygène .
L’assimilation biotique fait en sorte que les nitrates sont absorbés par les plantes
aquatiques et les algues pour synthétiser des matières cellulaires, comme les
protéines. On ne sait pas exactement de quelle façon les animaux aquatiques
absorbent les nitrates présents dans l’eau ni quels sont les mécanismes de toxicité
des nitrates pour ces organismes, bien que deux aient été proposés: a) la formation
de méthémoglobine, accompagnée d’une réduction du pouvoir oxyphorique du sang,
et b) l’incapacité de l’organisme d’assurer une osmorégulation convenable à une
teneur élevée en sels, conjuguée à de fortes concentrations de nitrates.
Des essais de toxicité des nitrates ont été effectués en ajoutant des sels d’acide
nitrique, comme le nitrate de sodium, le nitrate de potassium et le nitrate
d’ammonium. Les résultats obtenus portent à croire que, dans les essais utilisant le
nitrate d’ammonium, les effets toxiques observés sont dus à l’ion ammonium plutôt
qu’à l’ion nitrate. De même, dans l’eau douce, les effets du nitrate de potassium sont
probablement dus au potassium. Par contre, dans les eaux marines, les
concentrations toxiques de nitrate de potassium correspondent aux teneurs en
potassium inférieures aux concentrations de fond de cet élément dans l’eau de mer,
ce qui veut dire que la toxicité peut être attribuée à l’ion nitrate. À la lumière de ces
arguments, la toxicité des nitrates pour les organismes d’eau douce a été évaluée
seulement au moyen d’essais avec du nitrate de sodium, tandis que les données sur
la toxicité des nitrates de sodium et de potassium ont été utilisées pour les
organismes marins.
Les nitrates produisent des effets importants chez les invertébrés, le poisson et les
amphibiens, et les stades larvaires y sont généralement plus sensibles que les
adultes. Les effets nocifs observés chez les organismes aquatiques comprennent la
mortalité, la réduction de la croissance, la réduction du taux d’alimentation, la
diminution de la fécondité, la réduction du succès d’éclosion, la léthargie, des indices
de comportement dénotant un stress, le fléchissement de la colonne vertébrale et
d’autres malformations.
Une valeur provisoire de 13 mg de NO3-·L-1 pour la qualité des eaux est recommandée
en vue de la protection de tous les stades de vie aquatique d’eau douce contre les effets
nocifs de l’ion nitrate. Cette valeur a été calculée en multipliant la plus faible
concentration produisant un effet nocif observable signalée pour la réduction de la
croissance de la rainette du Pacifique (Pseudacris regilla), soit 133 mg de NO3-·L-1, par
un facteur de sécurité de 0,1. Une valeur provisoire de 16 mg de NO3-·L-1 est
recommandée pour la protection de la vie aquatique marine contre les effets nocifs
de l’ion nitrate. Cette valeur a été calculée en multipliant la concentration létale
médiane de 329 mg de NO3-·L-1 pour le polychète marin des régions tempérées
Science-based Solutions No. 1-6
ix
(Nereis grubei) par un facteur de sécurité de 0,05. Un facteur de sécurité plus
prudent a été utilisé afin de calculer la valeur recommandée pour la vie marine parce
que le polychète utilisé dans l’étude critique n’a pas été testé à son stade de vie le
plus sensible, que le paramètre critique, bien que chronique, était fondé sur un effet
létal médian plutôt que sur un effet sublétal faible, et que des effets nocifs ont été
observés chez des espèces tropicales non indigènes exposées à des concentrations
de nitrates beaucoup plus faibles.
Ces recommandations canadiennes pour la qualité des eaux visant la protection de
la vie aquatique pour l’ion nitrate ont pour but de prévenir la toxicité directe pour les
organismes aquatiques, mais elles ne préviendront pas nécessairement
l’eutrophisation. Par conséquent, même si les concentrations de nitrates sont
inférieures aux valeurs recommandées, il se peut que des effets toxiques indirects
dus à la prolifération d’algues se produisent encore.
Science-based Solutions No. 1-6
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LIST OF ACRONYMS
ADP
ANC
ATP
CAS
CCME
CCREM
CV
[C]WQG
DIN
DO
DOC
DOM
DON
EC
EC50
EDTA
H+
H2SO4
HNO3
IC
KNO3
LC50
LO[A]EL
LOEC
MDL
MWWTPs
N2
NADH
NaNO3
NH3
NH4+
NH4NO3
NO
N2O
NO2NO3NO3- -N
NO[A]EL
NOEC
NPRI
SCs
TDS
TLm
adenosine diphosphate
acid neutralizing capacity
adenosine triphosphate
Chemical Abstracts Service
Canadian Council of Ministers of the Environment
Canadian Council of Resource and Environment Ministers
coefficient of variation
[Canadian] Water Quality Guidelines
dissolved inorganic nitrogen
dissolved oxygen
dissolved organic carbon
dissolved organic matter
dissolved organic nitrogen
effects concentration
median effects concentration
ethylenediaminetetraacetic acid
hydronium ion
sulfuric acid
nitric acid
ion chromatography
potassium nitrate
median lethal concentration
lowest observable [adverse] effects level
lowest observable effects concentration
method detection limit
municipal wastewater treatment plants
molecular nitrogen
nicotinamide adenine dinucleotide, reduced form
sodium nitrate
un-ionized ammonia
ammonium ion
ammonium nitrate
nitric oxide
nitrous oxide
nitrite
nitrate
nitrate-nitrogen
no observable [adverse] effects level
no observable effects concentration
National Pollutant Release Inventory
safe concentrations
total dissolved solids
median lethal tolerance
Science-based Solutions No. 1-6
xi
TN
US EPA
UV
total nitrogen
United States Environmental Protection Agency
ultraviolet
Science-based Solutions No. 1-6
xii
1. INTRODUCTION
This report describes the development of Canadian Water Quality Guidelines
(CWQGs) for nitrate for the protection of freshwater and marine life. CWQGs are
numerical limits based on the most current, scientifically-defensible toxicological data.
They are nationally consistent benchmarks designed to protect, sustain and enhance
the present and potential uses of a water body. CWQGs are used by provincial,
territorial, and federal jurisdictions to evaluate water quality issues and manage
competing uses of water. The guideline values derived for nitrate are intended to
protect all forms of aquatic life and all aspects of aquatic life cycles, including the
most sensitive life stage of the most sensitive species over the long term.
This document describes production and uses, sources, and pathways for the entry
of the more common nitrate salts into the Canadian environment. Available data on
environmental fate and persistence of the nitrate ion are summarised. A
comprehensive assessment of the toxicity of selected nitrate salts to aquatic life is
also presented to evaluate environmental hazards posed by these chemicals.
Together, this information is used, in accordance with “A Protocol for the Derivation
of Water Quality Guidelines for the Protection of Aquatic Life”, (CCME 1991) to derive
numerical water quality guidelines (WQGs) for aquatic organisms.
It should be noted that nitrate concentrations are generally reported in this document
in terms of the nitrate ion rather than as nitrate-nitrogen (i.e., mg NO3-·L-1, not
mg NO3--N·L-1). Where source publications have used other units, these have been
converted for consistency to mg NO3-·L-1 wherever possible. In a few cases data is
presented in this document in terms of nitrogen, rather than nitrate, because we were
unable to assume how much of the nitrogen was in the form of nitrate; where this
occurs, the information is clearly identified as referring to nitrogen.
Science-based Solutions No. 1-6
1
2. PHYSICAL AND CHEMICAL PROPERTIES
2.1 Chemistry of the Nitrate Ion
The nitrate ion (NO3-), which has a molecular weight of 62 g·mol-1, is the most
oxidised form of nitrogen (N) present in the environment, with an oxidation state of +5
(NRC 1978). The molecule has a planar and symmetrical structure. The nitrogen
atom in the centre forms sigma bonds with the three oxygen (O) atoms using sp2
hybrid orbitals (NRC 1978). Other p orbitals of the nitrogen and oxygen atoms
combine to yield a pi bond that is shared among the three sites (Figure 2.1).
O
N
O
O
O
N
122 pm
O
O
Figure 2.1. Chemical structure of the nitrate ion. The Lewis diagram on the left is
adapted from McQuarrie and Rock (1991). The diagram on the right, adapted
from Petrucci (1989), depicts the delocalized pi molecular orbital.
The nitrate salts of all common metals (e.g., NaNO3, KNO3, Ca(NO3)2, AgNO3) are
highly soluble in water, and solutions of these salts are neutral in pH (NRC 1978).
While the resulting free nitrate ion has little tendency to form coordination complexes
with metal ions in dilute aqueous solutions (NRC 1978), under acidic conditions it can
act as a good oxidizing agent, as demonstrated in the reaction below (Petrucci 1989):
4 Zn(s) + 10 H+(aq) + 2 NO3-(aq) → 4 Zn2+(aq) + 5 H20 + N2O(g)
The nitrate ion also is the conjugate base of nitric acid (HNO3), a strong acid which is
completely dissociated in solution (NRC 1978). Physical and chemical properties of
the nitrate ion and selected nitrate salts commonly used in manufacturing are
presented in Table 2.1.
Science-based Solutions No. 1-6
2
Table 2.1. Summary of selected physical and chemical properties for nitrate ion and
selected nitrate salts.
Nitrate Ion
Sodium Nitrate
Potassium
Nitrate
Ammonium
Property
CAS #
14797-55-8
7631-99-4
7757-79-1
6484-52-2
Merck (1996)
Molecular
formula
Physical
structure
NO3-
NaNO3
KNO3
NH4NO3
CRC (1986)
• chemical
structure is
trigonal
planar
• colourless
transparent
prisms, white
granular or
crystal
powder
• pungent taste
101.10
• odourless,
transparent,
hygroscopic,
deliquescent
crystals or
white granules
Merck (1996)
80.04
CRC (1986)
Nitrate
Reference
Molecular
weight (g·mol-1)
Melting point
(°C)
Boiling point
(°C)
Density /
Specific gravity
Solubility in
cold water
(g·mL-1)
pH
62.00
• colourless
transparent
prisms, white
granular or
crystal powder
• deliquescent
in moist air
84.99
__
306.8
334
196.6
CRC (1986)
__
decomposes at
400°
2.109
210°
CRC (1986)
__
decomposes at
380°
2.261
1.725
CRC (1986)
__
0.921
0.133
1.183
CRC (1986)
__
Merck (1996)
__
neutral in
aqueous
solution
fireworks
pickling meat
manufacture
of glass
gunpowder
blasting
powders
tempering
steel
5.43 in 0.1 M
solution
Notes on use
neutral in
aqueous
solution
• manufacture
of nitric acid,
sodium nitrite,
glass and
enamels
• colour fixative
in meats
• fertilizer
• manufacture
of nitrous
oxide
• freezing
mixtures
• explosives
• matches
• pyrotechnics
• fertilizers
Merck (1996)
•
•
•
•
•
•
Science-based Solutions No. 1-6
3
The amount of nitrate present in a solution is often expressed relative to the amount
of nitrogen present in the NO3- ion, where 1 mg NO3-·L-1 is equivalent to
0.226 mg NO3--N·L-1 (WHO 1996). Other, less commonly used base units for nitrate
concentration include: g-at·L-1 (or g-at N·L-1), M NO3-, eq NO3-, and N NO3-.
Conversions between these units are given in Table 2.2. For consistency in this
report, unless otherwise specified, all nitrate concentrations will be reported for the
ion only (i.e., as mg NO3-·L-1).
Table 2.2. Conversion factors for various nitrate units to mg NO3-·L-1.
Base Unit
Multiply by:
mg NO3--N·L-1
mg NaNO3·L-1
mg KNO3·L-1
mg NH4NO3·L-1
eq·L-1, M, or g-at.·L-1 *
ppm NO3ppb NO3-
4.43
0.73
0.61
0.78
62.0 x 103
1
10-3
*note: for these units the conversion factor is the same
whether they’re expressed as NO 3--N or NO3-
2.2 Analytical Methods
There are several techniques available for analysing nitrate ions in aqueous
solutions. It may be difficult, however, to select the most appropriate technique for a
given application due to the limited concentration ranges available with each of the
techniques and the potential for interference from other compounds in the sample
matrix (APHA 1998). Table 2.3 provides an outline of nitrate ion analytical
techniques, their detection ranges and potential sources of interference.
Due to the potential for transformations between nitrate, nitrite, dissolved ammonia,
organic nitrogen and ammonia gas, it is important that certain procedures be used in
the collection, storage, and preservation of samples for nitrate analysis. Standard
methodologies, such as APHA (1998), should be consulted.
In general, nitrate analysis can be divided into three categories: colorimetric analyses
(various nitrate reduction processes); potentiometric analysis (ion-selective
electrodes); and, direct ion quantification (ion chromatography, capillary
electrophoresis).
The automated cadmium reduction method is commonly used for analysing nitrate
using colorimetry (NLET 1994; US EPA 2000a). In this method, nitrate present in a
sample must first be reduced to nitrite. To do this, the water sample is passed
through a glass column packed with cadmium (Cd) granules treated with CuSO4
which completely reduces nitrate to nitrite upon contact. The resulting nitrite is then
diazotised with sulfanilamide (NH2C6H4SO2NH2) and coupled with N-(1-naphthyl)Science-based Solutions No. 1-6
4
ethylenediamine dihydrochloride to form a reddish-purple azo dye (NLET 1994). The
absorption of the monochromatic radiation by the azo dye is proportional to the nitrite
concentration and is measured using a spectrophotometer at 520 nm (NLET 1994).
The same procedure without the reduction step is also applied on a subsample to
correct for NO2- ions originally present in the sample. It should be noted that this last
step of correcting for nitrite is frequently ignored, and some measurements reported
in the literature as nitrate concentrations may actually be concentrations of
nitrate+nitrite. The amount of nitrite in most water samples, however, is generally
quite small, particularly for samples originating from well-oxygenated waters.
The major advantage of the automated cadmium reduction method is greater
analytical
sensitivity,
with
nitrate
ion
detections
ranging
from
0.004 to 44.3 mg NO3-·L-1 (APHA 1998). Appropriate dilutions are required when
analyzing samples with the higher concentrations within this analytical working range
(NLET 1994). Potential interferences include: a) suspended matter that can restrict
sample flow in the column; b) high metal concentrations (e.g., Fe, Cu, etc.
> several mg·L-1) that can decrease reduction efficiency (in which case EDTA can be
used to chelate metals prior to analysis); c) hydrocarbons such as oil and grease
(must be pre-extracted with an organic solvent); and, d) residual chlorine which
should also be removed as it can interfere by oxidising the Cd in the column (APHA
1998).
This method is recommended for levels below 0.4 mg NO3-·L-1, where other methods
lack adequate sensitivity (APHA 1998). It should be noted, however, that Cd is very
toxic and, therefore, care must be taken when handling and disposing of it (US EPA
2000a).
The nitrate electrode method uses a pH meter with a dedicated NO3- ion electrode
that develops an electric potential across a thin, porous, inert membrane that
contains a water-immiscible liquid ion exchanger. The electrode measures ion
activity over a potentially wide range between approximately 0.62 to 6200 mg NO3·L-1 (APHA 1998). Although a complex buffer solution is required to remove potential
interferences from unwanted ions (e.g., Cl-, HCO3-, NO2-, CN-, S2-, Br-, I-, ClO3-, and
ClO4-), the electrode functions satisfactorily over a pH range of 3 to 9, provided pH
and ionic strength in the solution remain constant (APHA 1998). This method cannot
be used with samples that have high ionic strength, and therefore may not be
appropriate for many brackish or saltwater samples.
Science-based Solutions No. 1-6
5
Table 2.3. Comparison of available techniques for analysis of nitrate in water.
Technique
Analytical
detector
Detection
range
(mg NO3-·L-1)
Sample precision
cadmium
reduction
spectrophotometer
0.04 to 4.43 a
1.8 (12.5) to
4.60 (1.0) a
suspended matter;
oil and grease;
residual chlorine;
sample colours in
same wavelength
APHA: 4500-NO3- E a
ASTM: D 3867 b
US EPA: 0353.2 c
automated
cadmium
reduction
spectrophotometer
0.004 to 44.3 a
0.02 to 6.65 d
0.4 (0.0) to
9.3 (2.3) a
see Cd reduction
method
APHA: 4500-NO3- F, I a
ASTM: D3867 b
US EPA: 0353.2, 0353.6 c
NLET: 01-1181 d
automated
hydrazine
reduction
spectrophotometer
0.04 to 44 a
1.73 (5.1) to
21.0 (0.6) a
sulfide ion
concentrations
< 10 mg·L-1 can
cause variations
> 10%
US EPA: 0353.1 c
APHA: 4500-NO3- H a
brucine
reduction
spectrophotometer
0.44 to 8.8 c
5.49 (17.3) c
DOM causes colour
interference; strong
oxidizing and
reducing agents
US EPA: 0352.1 c
nitratespecific
electrode
pH meter with ionspecific electrode
0.62 to 6200 a
± 0.4mV
(= ± 2.5%CV) a
other anions;
inconsistent pH
APHA: 4500-NO3- D a
US EPA: 9210 e
Colorimetry
Potentiometry
(mean ± CV%)
(mg NO3-·L-1)
Science-based Solutions No. 1-6
Potential sources
of interference
Protocol reference
6
Technique
Analytical
detector
Detection
range
(mg NO3-·L-1)
Sample precision
capillary
electrophoresis
capillary electropherograph with
UV detector
0.0008* f
0.031 (2.7) f
ion chromatography
ion chromatograph
0.009 to 61.9 g
2.7 (33.3)
4.1 (2.17) d
Potential sources
of interference
(mean ± CV%)
(mg NO3-·L-1)
Protocol reference
Table 2.3 continued:
Direct ion
quantification
APHA: 4140 a
any substance with
a similar retention
time; high
concentrations from
similar anions may
mask anion of
interest
APHA: 4110 a
ASTM: D 4327 g
US EPA: 0300.0 e
NLET: 01-1080 d
notes:
* - MDL = Method Detection Limit
a
- APHA 1998
b
- ASTM 2000a
c
- Keith 1992
d
- NLET 1994; note: the upper end of this range can be extended with adequate sample dilution
e
- USEPA 2002
f
- Bondoux et al. (2000)
g
- ASTM 2000b
Science-based Solutions No. 1-6
7
Ion chromatography (IC) is another analytical method for measuring nitrate, with
detectable concentrations reported for NO3- using IC ranging from 0.009 mg NO3-·L-1
to 62 mg NO3-·L-1 (ASTM 2000c). There are two significant advantages of using IC.
First, unlike colorimetric, electrometric, or titrimetric methods for analysing ions, ion
chromatography can be used for sequential, rapid analysis of a suite of ions without
the need for hazardous reagents. Second, it is also capable of readily distinguishing
between NO2- and NO3- ions (APHA 1998; ASTM 2000c). Anions within a water
sample are separated by the ion chromatograph and measured using a conductivity
detector. The ion chromatograph consists of a guard column (that protects the
separator column from organics or particulates) and an anion separator column and
suppressor device (that separates the anions based on their relative affinities for the
strongly basic anion exchanger).
Capillary electrophoresis is a relatively new technique for the analysis of ionic
analytes. It is similar to IC, in that it can be used to distinguish between several
anions or cations simultaneously. Ion separation is based on individual
electromigration times and is quantified by direct UV detection (for nitrate and nitrite)
and indirect UV detection using a cationic UV chromatophore for the ammonium ion
(Padarauskas et al. 2000). The advantages offered by this method for nitrate analysis
over IC include short analysis time (~4 min per sample), improved ion resolution (and
therefore sensitivity), and more recently, the ability to simultaneously identify various
nitrogen anions and cations (e.g., nitrate, nitrite and ammonium) (Padarauskas et al.
2000). Under optimised conditions for anion analysis in pure water samples, Bondoux
et al. (2000) report a nitrate detection limit of 0.8 ppb (0.0008 mg NO3-·L-1). Although
the innovative simultaneous anion/cation technique allows for precise separation of
the three nitrogen ions, further method optimization is required for direct nitrate
quantification (Padarauskas et al. 2000).
Science-based Solutions No. 1-6
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3. NITRATE PRODUCTION AND RELEASE TO THE ENVIRONMENT
3.1 Nitrogen Cycle
Although the Earth’s atmosphere is composed of approximately 80% nitrogen, the
majority of this nitrogen pool is stored as nitrogen gas (N2) that is unavailable for use
by most organisms. The nitrogen cycle (Figure 3.1) serves to convert this biologically
unreactive nitrogen into useable forms for biota that are eventually cycled back to
nitrogen gas (Chambers et al. 2001).
Natural processes, such as forest fires and decomposition of organic matter, release
un-ionized ammonia (NH3), nitrous oxide (N2O), and nitric oxide (NO) into the
atmosphere (NRC 1978). In the atmosphere, these gases may undergo various
complex reactions (Chambers et al. 2001). The ammonia will react with hydroxyl (OH) radicals to produce NO and nitrogen dioxide (NO2). These two nitrogen oxides
(NOx) are also formed through the reaction of nitrous oxide with an oxygen atom.
Nitrous oxide may also dissociate to produce N2. Nitrogen gas is quite stable, and
only through lightning discharges is it converted to NO. Molecules of NO and NO2 in
the atmosphere will cycle back and forth in a complex reaction which involves the
formation of ozone. They can also react with water vapour or OH- radicals to form
nitric acid (HNO3) that can then enter aquatic ecosystems through precipitation
(Chambers et al. 2001).
Nitrogen occurs in surface waters in numerous forms, including dissolved molecular
nitrogen (N2), a variety of organic compounds (e.g., amino acids, amines, proteins
and refractory humic compounds), un-ionized ammonia (NH3), ammonium ion (NH4+),
nitrite (NO2-), and nitrate (NO3-) (Wetzel 1983). Nitrous oxide (N2O) may also occur in
surface waters, but rarely in appreciable quantities as it is rapidly reduced to N2
(Wetzel 1983), or out gassed and returned to the atmosphere. All aquatic and
terrestrial plants will assimilate nitrogen for protein production as either NO3- or NH4+,
however, the latter form requires less energy to assimilate and is therefore often
taken up preferentially (Crouzet et al. 1999). Nitrogen is also incorporated into
organic material (typically as amine [NH2] groups in organic nitrogen-compounds)
through biological fixation. In this process, N2 is reduced to ammonia that is then
incorporated into organic nitrogen compounds (NRC 1978). Aquatic nitrogen-fixing
species are limited to selected species of cyanobacteria (blue-green algae), and
photosynthetic and heterotrophic bacteria (NRC 1978). In terrestrial systems,
nitrogen-fixing bacteria in symbiotic association with leguminous plants (e.g., beans,
peas, alfalfa, clover, soybeans, lentils, peanuts) are major contributors of nitrogen to
the soil (NRC 1978; Chambers et al. 2001).
Science-based Solutions No. 1-6
9
(from Chambers et al. 2001)
Figure 3.1. The nitrogen cycle.
Science-based Solutions No. 1-6
10
3.2 Natural Sources
Natural sources of nitrate to surface waters can include wet and dry deposition of
HNO3 or NO3-. Atmospheric deposition of nitrate and ammonium in Canada is
estimated to contribute 182 kilotonnes (kt) of nitrogen per year to surface waters
(Table 3.1) (Chambers et al. 2001). This may be a conservative estimate because
data on dry deposition is lacking for many locations. Data collected over 1984-1994
show that wet deposition of nitrogen, on an areal basis, is considerably higher in
eastern than in western Canada (see Section 5.1). It should be noted that wet and
dry deposition are not entirely natural sources, as some of the nitrate and ammonia in
the atmosphere originates from anthropogenic sources. Other natural sources of
nitrate include igneous rocks, volcanic activity, and the complete oxidation of organic
nitrogen from vegetable and animal debris in native soil (Nordin and Pommen 1986).
This latter nitrification process is the principle source of nitrate in terrestrial and
aquatic environments (NRC 1978).
3.3 Anthropogenic Sources
All forms of inorganic nitrogen released into surface waters have the potential to
undergo nitrification to nitrate. Point source discharges of nitrogen include municipal
and industrial wastewaters, septic tanks, and water discharges from mining
(explosives) activity. On a national scale, point source discharges represent a small
fraction of total input of nitrogenous compounds to ground and surface waters (NRC
1978). The National Pollutant Release Inventory (NPRI) total point source estimate of
nitrate ion release from all participating Canadian sources for the year of 1999 was
6.8 kt NO3- to air, land, and surface and groundwaters (Environment Canada 2001).
Diffuse sources, however constitute the greatest inputs of anthropogenically-fixed
nitrogen and can include agricultural runoff, feedlot discharges, urban runoff, lawn
fertilizers, landfill leachate, nitric oxide and nitrogen dioxide from vehicular exhaust,
and storm sewer overflow (NRC 1972; NRC 1978). In a review of U.S. nitrogen
discharge estimates, van der Leeden et al. (1990) reported that point sources
contributed 561 kt N·a-1 (1977 data), while non-point sources contributed
9108 kt N·a-1 (1980 data). Although point sources account for only a small fraction of
the nitrogen released to surface and groundwaters, they can result in higher
concentrations because they are released into a small area.
Science-based Solutions No. 1-6
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Table 3.1. Nitrogen loading estimates to Canadian surface and ground waters from
various sources, 1996.
Total Nitrogen (103 t·a-1)
Nutrient
Source
Municipality
MWWTPs1
Sewers
Septic Systems
Industry
Agriculture
(residual in
the field after
crop harvest)
Aquaculture
Atmospheric
Deposition to
Water
(NO3- N and
NH4+ N only)
Total Loadings
Atlantic
Québec
Ontario
Prairies
British
Columbia
Territories
Canada
4.6
19.9
31.7
13.2
10.6
0.3
2.2
0.12
18
3.7
0.33
46
5.0
9.9
14
2.6
0.6
188
1.9
0.9
29
0.05
0
n/a
80.3
11.8
15.4
11.8
294
0.8
11.9
0.04
60.7
0.2
54.4
0.04
13.9
1.2
1.6
n/a
39.9
2.3
182
37.6
130.64
115.2
218.34
45.2
40.25
597.6
1
MWWTPs: municipal wastewater treatment plants
data from Newfoundland only
3
data for industries discharging to the St. Lawrence River
2
* (Industrial N loads are based on NO3- + NH3 and not total nitrogen; industrial data are not available for NB, NS
and PEI and Québec industries that do not discharge to the St Lawrence River. Agricultural residual is the
difference between the amount of nitrogen added to cropland and the amount removed in the harvested crop; data
are not available as to the portion of this residual that moves to surface or ground waters.)
(from Chambers et al. 2001)
3.3.1 Municipal Wastewaters
Humans excrete virtually all nitrogen obtained in protein from food sources. This
translates to an average excretion rate of 5.4 kg N per person per annum (NRC
1972). As of 1999, 86% of Canada’s population were served by municipal sewer
systems; the remaining 14% were served by septic disposal systems and lagoons
(Environment Canada 1999). Of those served by sewer systems, 97% were
connected to municipal wastewater treatment plants (MWWTPs) employing primary
(or better) treatment processes (Environment Canada 1999). The remaining 3% were
serviced by sewage collection structures that were not connected to treatment
facilities such that untreated wastewater was discharged directly into lakes, rivers or
oceans. Canadian loading estimates for nitrogen from wastewater sources for 1996
include 80.3 kt·a-1 from municipal water treatment plants, 11.8 kt·a-1 from storm
sewers and combined stormwater overflows, and 15.4 kt·a-1 from septic systems
(Table 3.1, Chambers et al. 2001).
Science-based Solutions No. 1-6
12
Among the facilities in the Canadian NPRI database which reported releases of
nitrate, sewage treatment facilities recorded the largest discharges of NO3-, due to
the nitrification of ammonia wastes (Environment Canada 2001). For example, the
Regional Municipality of Ottawa-Carleton released 0.66 kt of NO3- directly to
receiving waters in 1999; the City of Toronto’s Humber and Ashbridges Bay
MWWTPs each reported releases of 0.48 kt of NO3-; and the City of Medicine Hat
MWWTP reported a nitrate release of 0.44 kt (Environment Canada 2001). These
four MWWTPs all use secondary treatment or better. Average nitrate concentrations
measured between 1987 and 1994 in effluents from selected MWWTPs from across
Canada, with varying types of treatment, ranged from 0.05 to 27 mg NO3-·L-1
(Chambers et al. 1997). Examining nitrate levels alone in effluent, however, may only
give an indication of the degree of nitrification in the effluent. Concentrations of total
inorganic nitrogen in MWWTP effluents give a better indication of nitrate loading, as
ammonia and nitrite are readily transformed to nitrate in the receiving waters.
Nitrate levels in urban stormwater runoff can be highly variable depending on land
use patterns. In a review of 25 years of international runoff data from urban areas,
Makepeace et al. (1995) report a range in nitrate concentrations of 0.04 to
53 mg NO3-·L-1. Mean nitrate concentrations from storm event samples monitored
over a one-year period in the Brunette River watershed in British Columbia did not
exceed 4.0 mg NO3-·L-1 (Hall et al. 1999). Airports also contribute nitrate to
stormwater runoff through the breakdown of urea-based de-icing agents (DND 1998).
A review of nitrate levels between 1992 and 1996 from monitoring stations at federal
civil and military airport facilities reported nitrate levels in stormwater runoff of up to
116 and 1465 mg NO3-·L-1, at respective facilities (DND 1998).
3.3.2 Industrial Sources
Ammonium nitrate production in Canada began during the Second World War for use
in explosives. It was not until after the end of the war that large quantities were
available for use in fertilizers (McBeath 1987). Ammonium nitrate is produced by an
exothermic reaction between ammonia and nitric acid (McBeath 1987):
NH3 + HNO3 → NH4NO3
Natural gas is one of the primary raw materials in ammonia synthesis, and as such,
the majority of Canadian nitrogen fertilizer production is centred in Western Canada
where natural gas reserves are plentiful (SENES 2001). In 1999, there were twelve
Canadian facilities producing nitrogen fertilizers, six of which produced either
ammonium nitrate (totalling 498 kt·a-1) or solutions of urea [CO(NH2)2] and
ammonium nitrate (1273 kt·a-1) (SENES 2001). The other six facilities produced
ammonia, urea, and/or ammonium sulphate. As urea contains significantly higher
levels of fixed nitrogen than ammonium nitrate, on a unit mass basis, this product is
displacing traditional ammonium nitrate fertilizer markets, and since 1987, five
ammonium nitrate production facilities have ceased operations (McBeath 1987;
SENES 2001). Approximately one-half of ammonium nitrate and urea production is
used nationally, while the remainder is exported to the U.S. (SENES 2001).
Science-based Solutions No. 1-6
13
Provincial limits for nitrate in fertilizer plant wastewater in Alberta and British
Columbia are 88 and 45 mg NO3-·L-1, respectively (McBeath 1987). In 1980/81
however, effluent monitoring from selected Canadian fertilizer producers revealed
that nitrate levels ranged from 0.13 to 3400 mg NO3-·L-1 (McBeath 1987). By 1999,
six of the twelve Canadian fertilizer plants were “zero discharge” facilities that either
directed their effluents to municipal water treatment plants, or used on-site
evaporation ponds (SENES 2001). Of the remaining plants for which data exist,
nitrate concentrations in effluents discharged directly to receiving waters ranged from
0.4 to 56.2 mg NO3-·L-1 (SENES 2001).
Nitrate metal salts such as potassium nitrate, calcium nitrate, silver nitrate and
sodium nitrate are used in a variety of industrial applications, including oxidising
agents in explosives, matches and pyrotechnics, photography, glass making,
engraving, textile dyes, food processing (e.g., meat preservatives), and as a raw
material for manufacturing nitric acid (Nordin and Pommen 1986; WHO 1996).
Industrial sources with high concentrations of inorganic nitrogen effluents include
steel production, petroleum production and refining, pulp and paper, plastics and
fertilizer production (Heathwaite et al. 1996). Other industrial processes that are
known to result in high nitrate concentrations in their wastestreams include the
production of nitroaromatic compounds, the synthesis of nitroorganic compounds in
pharmaceuticals, and wastewaters from nuclear fuel processing (Pinar et al. 1997).
Mining activities can also be a source of nitrate to Canadian waters. Nitrate, resulting
from the use of explosives containing ammonium nitrate, may enter surface waters
through mine drainage from pits and spoil piles, and through seepage from tailing
ponds (Pommen 1983). Elevated levels of nitrate have been noted downstream from
several Canadian mines (Pommen 1983). For example, on the Fording River in
southeastern British Columbia, Nordin (1982) found that nitrate concentrations
upstream from a surface coal mine ranged from 0.22 to 0.31 mg NO3-·L-1, while river
concentrations within the minesite were as much as 200 times higher, ranging from
4.4 to 44 mg NO3-·L-1.
Total Canadian industrial loading of inorganic nitrogen (nitrate and ammonia) to
surface waters is estimated at 11.8 kt N·a-1 (Table 3.1, Chambers et al. 2001). This
value, however, underestimates actual loads as not all industries are monitored
nationally, and monitoring data were not available for industries in New Brunswick,
Nova Scotia, and Prince Edward Island, nor for industries in Québec which do not
discharge directly into the St. Lawrence River Basin (Chambers et al. 2001).
Science-based Solutions No. 1-6
14
3.3.3 Agricultural Sources
During the last six months of 1998 and the first six months of 1999, a total of 1600 kt
of nitrogen as fertilizer were sold (and assumed to be consumed) in Canada (Korol
and Rattray 2000). Of this, 90 kt of nitrogen were nitrate compounds, with 82% as
ammonium nitrate; remaining forms included calcium nitrate, calcium ammonium
nitrate and potassium nitrate (Korol and Rattray 2000). The other 1500 kt of nitrogen
sold in Canada was contained in fertilizers such as urea, anhydrous ammonia, and
monoammonium phosphate, among others. These levels correspond with 1999
estimates of total nitrogen consumption by plants in Canada of 1626 kt (Korol and
Rattray 2000). The amount of nitrogen fertilizer applied to Canadian cropland has
increased considerably over the past century, due to both increased fertilizer
application rates and increased land usage (Chambers et al. 2001). For example, the
amount of nitrogen applied to the western Canadian grain crop in 1986 was four-fold
greater than the average amount applied annually between 1883 and 1953
(Chambers et al. 2001). Annual nitrogen fertilizer use in the United States has also
increased dramatically from 450 kt to 9980 kt in less than 50 years (Lanyon 1996).
Although the total Canadian use of nitrogen fertilizer continues to rise, within the
provinces of Ontario and British Columbia sales in recent years have been
decreasing, after hitting peaks in 1985 and 1989, respectively (Korol and Rattray
2000).
Among the various regions of Canada, the greatest loadings of nitrogen per unit area
of agricultural land in 1996, through the application of fertilizer, occurred in Québec
and the Atlantic region, with 89 and 86 kg N·ha-1 applied, respectively (Chambers et
al. 2001). In Manitoba, British Columbia, Ontario, Alberta, and Saskatchewan, the
amounts of nitrogen applied as fertilizer in 1996 were 82, 75, 72, 61, and
52 kg N·ha-1, respectively (Chambers et al. 2001).
The practice of spreading animal waste slurries (manure) as organic fertilizer also
constitutes a significant source of agricultural-nitrogen loading. In 1994, more than
34 000 kt of manure (containing approximately 141 kt N) were generated in Ontario
alone (OMAFRA 1996). Nationally, approximately 384 kt of nitrogen were applied to
crop land as manure in 1996 (Chambers et al. 2001).
Nutrient contents of manure vary according to animal source. Solid manure from
broiler chicken litter contains 29 kg N·t-1, whereas pig and cattle manure contains
6 kg N·t-1. For liquid slurries applied directly to fields, pig manure contains 5 kg N·m-3
as opposed to cattle slurry with 3 kg N·m-3 (Hooda et al. 2000). Within a species,
nutrient manure may also vary depending on the diet of the livestock. For example,
dairy cattle from Ontario, which are primarily corn-fed, produce manure with a
typical nitrogen content of 1.5 kg N·t-1, whereas the manure from dairy cattle in
Alberta, which are generally grain-fed, typically contains 4.5 kg N·t-1 (Hilborn and
Brown 1996; Statutes of Alberta 2001). Manure processing also affects nitrate
composition. At a beef cattle feedlot, for example, fresh manure used for crop
applications can contain 0.115 kg NO3-·t-1, while composted manure allowed to
undergo nitrification can contain 5.33 kg NO3-·t-1 (Eghball and Gilley 1999).
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15
Manure produced on intensive livestock farms often far exceeds the agronomic
requirements, resulting in large amounts of unutilised, or surplus, nitrogen. In a study
of seven different farming systems in Ontario, Goss and Goorahoo (1995) found that
larger surpluses of nitrogen were likely to occur on dairy farms than on swine farms,
or farms with crops only. Examination of nitrogen inputs and outputs for a dairy farm
in the Waterloo region of Ontario showed a surplus of 77 kg N·ha-1 (Millman 1999).
Millman noted that the Ontario farm was very efficient compared with other farms
from the United States and Europe which reported higher nitrogen surpluses. Hooda
et al. (2000) cited an example of 177 Dutch dairy farms that showed an average
nitrogen surplus of 486 kg N·ha-1. In a study of the effect of fertilizer type on nitrate
levels in agricultural runoff, Eghball and Gilley (1999) found that NO3--N accounted
for 21%, 25% and 37% of total nitrogen (TN) found in field runoff waters fertilized with
inorganic fertilizers, fresh manure and composted manure, respectively. Several
Canadian provinces currently have regulations for manure storage and land
application on intensive livestock farms to reduce impacts on aquatic systems.
Canadian soil nitrogen surpluses for 1996, based on national application rates and
crop removal from harvesting, are estimated at 294 kt N·a-1 (Table 3.1, Chambers et
al. 2001). Due to high production levels of nitrogen-intensive crops such as corn and
soybeans, Ontario and Québec contained the greatest share (37 and 27%,
respectively) of agricultural lands at risk of having > 60 kg N·ha-1 residual nitrogen
remaining after harvesting (MacDonald 2000a). As soils in these areas also
experience water surpluses, they are at the greatest risk of exporting excess nitrogen
to the watershed. Using data for soil water-holding capacity and regional 30-year
precipitation averages, MacDonald (2000b) determined that 17% and 6% of the
agricultural lands in Ontario and Québec, respectively could generate runoff or
seepage water with > 14 mg N·L-1. Between 1981 and 1996, the nitrogen content of
water moving off agricultural land to surface and groundwater was estimated to
increase by at least 1 mg N·L-1 on 68% of Ontario’s and 77% of Québec’s farmlands
(MacDonald 2000b). However, it should be noted that MacDonald’s estimates are
based on modelling, without measurements to evaluate the reliability of the
predictions; actual groundwater analyses in rural Ontario have not shown a temporal
increase in the proportion of farm wells with nitrate contamination. A survey of
domestic well water from Ontario farms in 1992 showed approximately 14% of wells
contained nitrate concentrations above the provincial drinking water guideline, the
same percentage of exceedances that were observed in a survey conducted in 19501954 (Goss et al. 1998a).
Although national estimates quantifying nitrogen loss to surface and groundwaters
from agricultural lands are not available (Chambers et al. 2001), NO3--N has been
shown to account for 97-98% of sub-surface nitrogen in leaching loss studies from
Quebec and Georgia (Lowrance 1992; Gangbazo et al. 1995). As such, losses from
residual nitrogen from agricultural soils (Table 3.1), can provide a major source of
nitrate to surface or groundwaters. In some regions of the United States, up to 54%
of nitrogen in surface waters is thought to originate from agricultural runoff or other
rural sources (NRC 1972). Mean annual total nitrogen losses to rivers from
agricultural subcatchments within the Lake Simcoe, Ontario watershed were highest
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16
from low-land cultivated marshes (or polders) used in the production of vegetables, at
25 (± 24) kg N·ha-1·a-1, followed by mixed agricultural lands, at 2.2 (± 0.7) to
7.9 (± 3.6) kg N·ha-1·a-1 (Winter et al. 2002). By comparison, forested areas in the
watershed generally exported the least amount of nitrogen, at 1.7 (± 0.5) to
2.7 (± 0.8) kg N·ha-1·a-1 (Winter et al. 2002).
Aquaculture is a $355 million per year industry in Canada, with finfish and shellfish
production totalling 53 and 19 kt, respectively in 1996 (DFO 1998). Nutrient loading
from animal wastes and decomposition of unused food in semi-closed and open
culturing systems are estimated to contribute 1.0 and 1.3 kt N·a-1 to inland and
coastal surface waters, respectively (Table 3.1, Chambers et al. 2001).
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4. ENVIRONMENTAL FATE AND BEHAVIOUR
4.1 Atmospheric Processes
4.1.1 Wet Deposition
Anthropogenic processes such as fossil fuel burning and ore smelting release SOx
and NOx to the atmosphere where they undergo hydrolysis and oxidation to form the
acid rain causing compounds H2SO4 and HNO3 (Galloway and Dillon 1983).
Subsequently, nitrate is one of the dominant ions present in precipitation (Fowler et
al. 1999). In the early 1980s, nitric acid deposition contributed approximately 35% of
the acidity of acid rain in eastern Canada and the northeastern United States (with
the other 65% contributed by sulfuric acid) (Galloway and Dillon 1983). However, due
to reductions in the emission of sulphur oxides since the early 1980s, nitric acid has
accounted for an increasing proportion of the acidity. Between 1976-77 and 1985-86,
the ratios of NO3- to SO42- in atmospheric deposition have increased in central
Ontario from 0.43 to 0.68 (Dillon and Molot 1989). A long-term study of atmospheric
inputs to Heney Lake, situated in Ontario on the Canadian Precambrian Shield,
showed almost no change in the amount of nitrate in precipitation over the period
from 1976 to 1997, but by the 1990s, nitrate and sulphate ions were present in
precipitation at almost equal amounts (Dillon and Evan 20021).
Wet deposition of ammonium is another major atmospheric source of nitrogen. In
some parts of Canada, deposition of ammonia can be as great as, or greater than,
deposition of nitrate (Chambers et al. 2001). Once deposited in aquatic or terrestrial
ecosystems, some ammonium may be taken up by plants, but the remainder is
generally converted to nitrate through nitrification.
In some catchments, atmospheric deposition accounts for the majority of nitrate
concentrations in surface waters, with very little export from the terrestrial system
(Lovett et al. 2000). There are also areas of Canada where less than 10% of the total
deposition of atmospheric nitrate to surface waters occurs through direct deposition,
with the majority of the deposition occurring on land with subsequent transport of the
nitrate ion from the terrestrial basin to surface waters (Elder 1984). Most deposited
atmospheric nitrogen, however, is likely retained in the terrestrial ecosystem and
assimilated into biomass (Jeffries and Semkin 1983). Aquatic systems are most at
risk of acidification if the terrestrial system is already saturated with nitrogen, in which
case atmospherically deposited nitrate will be released along with an equivalent
amount of cations. If the cation is H+ or Al3+, then acidification of the water will result
(Galloway and Dillon 1983). The maximum deposition of nitrogen compounds (NOx
and NHx) that will not cause eutrophication or acidification is referred to as the critical
load of nitrogen (RIVM 1991). Using critical load modelling, extensive mapping has
been conducted in Europe to determine which terrestrial and freshwater ecosystems
are at risk of acidification and eutrophication due to excess nitrogen deposition (RIVM
2001a).
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4.1.2 Dry Deposition
Dry deposition of oxidised nitrogen generally occurs in the form of nitrogen dioxide
(NO2) or nitric acid (HNO3) (Fowler et al. 1999). Ammonia may also enter aquatic and
terrestrial systems through dry deposition. The nitrate form of nitrogen is only
precipitated from the atmosphere in the form of wet deposition. The other nitrogen
species that do undergo dry deposition, however, may form nitrate once in the
receiving environments.
4.2 Terrestrial Processes
4.2.1 Adsorption
The nitrate ion is negatively charged, and therefore does not adsorb to clay minerals
or organic matter in soils unless they have a significant anion exchange capacity
(Jury and Nielsen 1989). Soils with large anion exchange capacities are very
uncommon, except in tropical areas. Therefore, with respect to the Canadian
environment, it can be assumed that nitrate does not adsorb to soil particles and has
a high potential for mobility. Both leaching and surface runoff are major fate
processes of nitrate in the terrestrial environment.
4.2.2 Leaching
In soils, the nitrate ion is highly mobile, readily moving with the soil water, and,
therefore, can potentially leach below the rooting zone (Hooda et al. 2000). Leaching
is the most significant process by which nitrate can enter groundwaters and is
dependent on the water supply from precipitation and irrigation, evaporation and
drainage rates, tillage practices, the type of fertilizer applied (organic vs. inorganic),
the type of ground or crop cover, and the soil structure and porosity (Table 4.1).
Moisture and temperature are major factors affecting the leaching of nitrate in soils.
Nitrate is moved downward in the soil with rainfall and irrigation, while upward
movement may occur in the very upper layers of the soil through evaporation (NRC
1978). Downward movement of nitrate is reduced at low temperatures because water
drains more slowly through cold soils; this effect is only significant, however, when
temperatures are below freezing, at which point water completely ceases to drain
(NRC 1978). Extreme variations in temperature, such as freezing of soil following by
thawing, can lead to greater leaching of nitrate (Mitchell et al. 1996). Saturated soil
conditions due to high water tables will enhance denitrification (see Section 4.2.5),
while all other processes occur at faster rates when the soil moisture content is below
field capacity (Madramatoo et al. 1997).
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Table 4.1. Factors affecting nitrate leaching through agricultural soils.
Factor
Less Leaching
More Leaching
Climate
Low rainfall
Cold temperatures
High or irregularly
distributed rainfall
Warm temperatures
Crop
Vigorous crop
Poor crop
Time of Application
At the beginning of the main
growing period or during
active crop growth
Established crop
At the end of the growing
season or out of season
Seedbed application
Application Rate
Rate appropriate for crop
use
Over-application
Soil
Fine soil (e.g., clayey)
Poor drainage
Limited soil tillage
Coarse soil (e.g., sandy)
Good drainage
Intensive soil tillage
(adapted from Ritter et al. 2001)
Leaching of nitrate from soil into groundwater appears to follow seasonal trends.
Through the use of field lysimeters, Roy et al. (2000), in Guelph, Ontario, found that
very little leaching of nitrate occurred following spring and summer applications of
ammonium-nitrate fertilizer to turfgrass, but an average of 16.5% of the applied
nitrogen was lost through leaching in late autumn and early winter. Possible reasons
for greater nitrate leaching in late autumn include increased precipitation coupled with
reduced uptake of water by plants. Roy et al. (2000) speculate that washing out of
nitrate that has accumulated in soil during the spring and summer could occur as a
single autumn pulse to the water table, resulting in high transient concentrations of
nitrate in groundwaters. Ezeonu and Okaka (1996) have also observed seasonal
trends in the occurrence of nitrate in Nigeria’s groundwater. Concentrations of nitrate
entering the aquifers are highest at the beginning of the rainy season, decrease
throughout the rainy season, and remain at relatively constant low levels during the
dry season.
During dry periods, nitrate may accumulate in soil due to decreased transport to
streams, decreased uptake by plants, and, with the declining water table, increased
capacity for storage of nitrate as the thickness of the unsaturated zone above the
water table increases (Lucey and Goolsby 1993). Under wetter conditions, the water
table rises, and nitrate stored in what was previously the unsaturated zone becomes
mobilised and may be transported by subsurface flow to surface waters. In a test of
this nitrate flushing theory, nitrate-nitrogen release from soils was modeled for the
forested catchments in the Turkey Lakes Watershed of Ontario (Creed et al. 1996).
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Two mechanisms were suggested for producing significant concentrations of nitrogen
in catchment discharge waters: (1) a rapid flushing of nitrogen from throughwaters
entering a previously unsaturated zone high in nitrate from either a period of low
biological activity (e.g., spring snowmelt and autumn stormflow), or soils that had
previously undergone enhanced nitrification (e.g., after summer droughts), or (2)
through a slow draining of nitrogen from the bioactive soil layers to non-active layers
through percolation to be released slowly throughout the year (Creed et al. 1996). Of
these two processes, rapid flushing is the dominant mechanism.
In an examination of soils and groundwater beneath an agricultural field receiving
nitrate fertilizer applications, nitrate concentrations were generally found to decrease
exponentially with soil depth (Schuh et al. 1997). Elevated concentrations at all soil
depths occurred temporarily following large rainfall events. During these brief periods
of large water influx, concentrations of nitrate in groundwater were observed to
increase by an order of magnitude or more (Schuh et al. 1997). In some cases, the
downward movement of nitrate during rainfall or flooding events can be quite rapid,
due to the vertical hydraulic gradients that are created. For example, stable isotopelabeled 15N sodium nitrate applied to the surface of an Illinois agricultural field was
found to travel 4.5 m vertically in the soil horizon within 16 hours following an annual
flooding event from the nearby Illinois River (Kelly and Wilson 2000).
The type of vegetation or forest cover in a watershed can affect the amount of nitrate
retention in the soil. For example, in the Catskill Mountains, New York, Lovett et al.
(2000) found that forests where red oak and beech trees dominate had higher stream
nitrate concentrations than forests dominated by maples. They attributed this
difference to the quality of the different leaf litters in terms of lignin-to-nitrogen ratios
and potential rates of nitrification.
The type of cropping system used on agricultural lands can have a large influence on
the amount of nitrate lost through leaching. Randall et al. (1997) found that row-crop
systems, such as continuous corn, or annually alternating corn and soybean
systems, resulted in nitrate losses about 45 times higher than that in perennial crops,
such as alfalfa or mixtures of alfalfa and grasses. Annual crops such as corn and
soybeans allow for greater losses of nitrate because they are shallower rooted, have
shorter growing seasons, and use water less efficiently (Randall et al. 1997). The
water balance in fields planted with annual crops will generally favour drainage rather
than evaporation, hence nitrate will also tend to leach downwards.
Certain agricultural practices, such as tilling, fertilizer and manure application, and
improved subsurface drainage through tile lines also contribute to greater loss of
nitrate through leaching (Randall et al. 1997). A study of rivers in Ireland concluded
that the major factor affecting nitrate levels in the rivers was the proportion of
ploughed land area in the catchment (Neill 1989). Mean nitrogen loss from ploughed
land was estimated at 75.9 kg·ha-1·a-1 compared with only 1.9 kg·ha-1·a-1 from
unploughed land (Neill 1989). In a study of soil plots with a drainage system, the
amount of nitrate leached from plots that were ploughed was 21% more than from
direct-drilled (untilled) plots (Goss et al. 1993). In plots with subsoil drains, five times
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more nitrate was lost through leaching than from undrained soils (Goss et al. 1993).
Greater nitrate leaching has been observed with grassland that is used for grazing
livestock than when the grass is cut, due to the additional nitrogen inputs from the
livestock manure (Ryden et al. 1984).
Timing of fertilizer application can have a large effect on nitrate leaching. To reduce
the amount of leaching, it is important to synchronize nitrogen additions (through
fertilizer or manure applications) with nitrogen mineralization in the soil and nitrogen
uptake by the crop (Izaurralde et al. 1995). Application methods for organic fertilizers
may also affect the amount of leaching that occurs. Leaching of nitrate is more likely
when the injection method for manure application is used than when it is broadcast
on the soil surface (Sutton et al. 1982). The injection method, however, is better for
reducing volatilization. Differences in leaching have also been noted among different
forms of fertilizers. For example, Sutton et al. (1978) observed greater downward
movement of nitrate through soil that had received inorganic fertilizer than soil that
had received swine manure, despite the higher nitrogen content of the manure. The
original form of nitrogen in the fertilizer was urea, while the manure contained both
ammonium and organic forms of nitrogen. Therefore, less inorganic nitrogen may
have been available for leaching from the swine manure due to the slower
decomposition of the organic matter (Sutton et al. 1978). The higher carbon content
in manure than in inorganic fertilizers may also promote increased denitrification in
the soil profile, reducing the potential for nitrate contamination of groundwater (Burton
et al. 1994).
Winter crop covers can also aid in reducing nitrate runoff from agricultural fields.
During a three year study on winter soil nitrate leaching under sweet corn (Zea mays
L.) or broccoli (Brassica oleracea var. italica Plenck) crops in Oregon, Brandi-Dohrn
et al. (1997) found that recommended crop-specific nitrogen application rates (up to
280 kg N·ha-1·a-1) resulted in flow-weighted mean nitrate levels in winter leachate
under fallow fields of 77 mg NO3-·L-1. Planting a winter cereal rye (Secale cereale L.)
cover crop, however, significantly (p < 0.05) reduced nitrate levels in soil leachate by
34 to 39% (Brandi-Dohrn et al. 1997). On the Western Canadian prairies, continuous
cropping has been found to result in less nitrate leaching than that observed for crop
rotations that include a fallow season (Campbell et al. 1984). Goss et al. (1998b) also
found that, compared with leaving fields fallow, winter cover crops decreased nitrate
leaching by 36% in the periods in which they were growing. However, they also found
that over the long term, growth of winter cover crops could result in greater net levels
of nitrate leaching due to nitrogen releases from the cover crop residues in the
following autumn (Goss et al. 1998b).
A reduction of vegetative cover through forest fires, logging, or insect defoliation can
result in increased inputs of nitrate to surface waters. For example, a study of
peatlands in northern Alberta that were razed by fire showed that the water of lakes
from burnt catchments contained three-fold higher nitrate concentrations than
reference lakes (McEachern et al. 2000). Clear-cutting of a watershed in the Hubbard
Brook Experimental Forest increased streamwater nitrate concentrations by
approximately 50-fold (Likens et al. 1970). Increased stream export of nitrate was
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observed in Appalachian hardwood forests during periods of intense defoliation by
cankerworms and gypsy moth (Swank et al. 1981; Webb et al. 1995).
Soil type is another factor that affects the amount of nitrate leaching. Coarse-textured
soils generally support greater leaching, or infiltration, and, therefore, favour transport
of nitrate to groundwater (Druliner 1989; Spalding and Exner 1991). The largest
nitrate losses occur in sandy and peat soils, moderate nitrate leaching occurs in
loamy soils, and smaller losses occur in clay soils (Bergstrom and Johansson 1991).
Although there may be less leaching of nitrate from fine-textured, less permeable, or
poorly-drained soils, these soils may lose more nitrate to streams through surface
runoff (Hooda et al. 2000). In agricultural fields comprised of fine textured soils,
significant amounts of nitrate may also be transported to surface waters through tile
drain systems.
Geochemical characteristics of the soil may also affect the degree of nitrate leaching.
Robertson et al. (1996) found that where reduced sulphur compounds were present
at higher concentrations in the soil, greater attenuation of nitrate leaching occurred.
They reasoned that the sulphur provided an electron donor for autotrophic
denitrification of the nitrate. Again, less leaching of nitrate would be expected in silt
and clay-rich soils as these typically have higher sulphur contents than sandy soils
(Robertson et al. 1996).
4.2.3 Water-driven Erosion or Runoff
During heavy precipitation and snowmelt episodes, when soils are water-saturated,
or where the ground is impermeable, surface runoff will occur. Runoff may transport
dissolved nitrate to surface waters, or, where soils are unstable, it may result in the
erosion of soil containing nitrate into surface waters.
Lamontagne et al. (2000) examined the fate of 15N-labelled nitrate applied to a Boreal
Shield catchment at the Experimental Lakes Area in northwestern Ontario. NaNO3
was applied to the test area at a rate of 40 kg N·ha-1·a-1 over a two year time period.
The fate of the nitrate was then determined by measuring the amount of 15N stored in
the biomass of trees, ground vegetation, litter and soil, and by estimating 15N loss
through runoff. Elevated levels of nitrate in runoff were associated with snowmelt and
small rain events that followed a dry period. Approximately 16% of the 15N added to
the experimental area was lost through runoff (Lamontagne et al. 2000). Estimates
from similar temperate forest experiments suggest that approximately 10% of
elevated nitrogen inputs are lost through leaching or volatilisation (Nadelhoffer et al.
1999). Large-scale manipulations of forests in Europe have indicated that there is a
critical threshold for nitrogen loading (Dise and Wright 1995; Bredemeier et al.
1998). At inputs below the threshold (of approximately 10 kg N·ha-1·year-1), the forest
ecosystems were capable of retaining most of the N, but when the threshold was
exceeded, saturation occured and the ecosystems responded rapidly with high N
outputs in runoff (Dise and Wright 1995). In saturated forests, it was possible for N
exports to equal, or even exceed, inputs (Bredemeier et al. 1998).
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Short-term increases in acidity of lakes may occur during periods of heavy surface
runoff, for example, during the snow-melt period (Elder 1984). Analyses of the
snowpack in Algoma, Ontario showed that nitrate concentrations in the snow
gradually increase throughout the winter months, in tandem with an increase in water
content, reaching a maximum in March (Jeffries and Semkin 1983). Both the water
content and nitrate concentration of the snow plummet in April as the snow melts.
The nitrate content decreases more rapidly than the water content with the result that
the discharge of the early meltwater has a much higher nitrate content, and lower pH,
than the snowpack or the later snowmelt. With a brief, but intense pulse of nitrate in
the watershed, an associated pH depression can occur. In some Adirondack lakes of
Vermont that exhibit low baseline acid neutralising capacity throughout the year,
nitrate pulses are more likely to reduce pH than a concomitant increase in the dilution
of base cations (Stoddard and Kellog 1993). Although these acidification episodes
are generally short-lived, the timing is cause for concern because many aquatic
organisms are at sensitive life stages during the spring (Harvey et al. 1981).
Mueller et al. (1997) found that land use and hydrologic basin characteristics can be
used to predict areas where high nitrate concentrations are likely to occur in streams.
Logistic regression modelling was used with the predictive variables being
streamflow, the amount of surrounding land area in corn production (or, alternatively,
the amount of fertilizer application), soil texture and water drainage characteristics,
and population density. In a study of the Duffin Creek drainage basin (just east of
Toronto, Ontario), nitrate losses from soils were highly correlated with the amount of
land area used in crop production, and to a lesser extent, with the area of imperfectly
drained soils, sandy loam soils, main stream channel gradient and drainage basin
relief ratio (Hill 1978). The factor most highly correlated with mean annual nitrate
concentrations in the stream water was crop area (Hill 1978). Examination of a
watershed in Massachusetts showed that nitrate concentrations were positively
correlated (R2 = 0.68) with the percentage of the catchment area classified for human
use, i.e., agricultural, residential, commercial, industrial, urban open, and
transportation areas (Rhodes et al. 2001). Through studies in Maryland and
Pennsylvania, Correll et al. (1995) also found a strong relationship between nitrate
concentrations in streams and the dominant land use of the watershed. Streams
surrounded by cropland and pasture had consistently higher concentrations of nitrate
than streams in forested watersheds (Correll et al. 1995). Similar observations have
been made in Alberta where the amount of inorganic nitrogen exported from
agricultural watersheds was more than an order of magnitude higher than that in
forested watersheds (Cooke and Prepas 1998). The speciation of inorganic nitrogen
also differed with land use. Nitrate was the predominant nitrogen species in runoff
from cropland, comprising 98% of the total inorganic nitrogen pool (Cooke and
Prepas 1998). In forested watersheds, approximately half of the inorganic nitrogen
was nitrate, and in a mixed agricultural watershed (comprising cropland and two
cattle operations), 94% of the nitrogen in runoff was NH4+ (Cooke and Prepas 1998).
In this case, the authors speculated that the large nitrate inputs from cropland could
be attributed to excessive inorganic fertilizer use, whereas the large ammonium
inputs from the mixed agricultural land were likely due to poor manure management.
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The amount of nitrate loss from agricultural land can be reduced by certain cropping
practices. On sensitive landscapes, reduced or zero tillage and the planting of
perennial forages can help to alleviate erosion. Vegetative buffer strips along the
edges of water courses can also help to reduce the amount of nitrate entering the
water through erosion and runoff.
4.2.4 Biotic Uptake and Assimilation
There are several forms of inorganic nitrogen (i.e., nitrate, ammonium, dinitrogen)
and organic nitrogen (i.e, urea, amino acids) available to plants in soils (Crawford and
Glass 1998). Under typical aerobic conditions found in agricultural soils, nitrate is far
more prevalent, as shown in a review of 35 agricultural soils where nitrate levels
(6.0 mM NO3-) greatly exceeded those of ammonium (0.77 mM NH4+) (Crawford and
Glass 1998, and references therein). Nitrate in soil is rapidly absorbed by plant roots
for assimilation into proteins (Viets, Jr. 1965; Jury and Nielsen 1989). The rate of
absorption will depend somewhat on the rate of water uptake by the plant due to
transpiration; however, it is not entirely a passive process as plants are also able to
regulate nitrate uptake rates (Viets, Jr. 1965). To compensate for large seasonal and
regional variations in soil nitrate concentrations, plants have evolved genetically
regulated transport systems that take up nitrate from the soil against an
electrochemical gradient (Crawford and Glass 1998). The energy required by the
plant (even when external concentrations are relatively high), is provided from proton
gradients (or proton motive forces), which facilitates the transport of the nitrate ion
and two accompanying protons from the external medium into the cell (Crawford and
Glass 1998).
Because terrestrial plants can absorb nitrate against a concentration gradient,
bioaccumulation can occur. Nitrate concentrations in the roots or stems of plants may
become hundreds of times higher than that in the surrounding soil or culture solution
(Viets, Jr. 1965). For example, cytoplasmic nitrate levels in barley seedlings, which
were below detection limits in nitrate-deprived conditions, increased to 620 to
2170 mg NO3-·L-1 when available nitrate levels were increased from 0.62 to
62 mg NO3-·L-1 (Siddiqi et al. 1991).
Riparian zones, or buffer strips, between agricultural fields and streams can help to
reduce nitrate loadings from shallow groundwaters to the stream (Cook 1999). Hunt
et al. (1995) found that a riparian zone removed substantial amounts of nitrate from
the shallow groundwater of a swine wastewater disposal site. Nitrate levels of up to
97 mg NO3-·L-1 in subsurface water, passing through either grassland or woodland
buffer zones, are consistently reduced to less than 9 mg NO3-·L-1 (Muscutt et al.
1993). Subsurface nitrate removal below buffer zones appears to occur over short
distances, as the majority of nitrate removal in studies by Cooper (1990), and
Haycock and Burt (1993) occurred within the first 5 m and 8 m of the zones,
respectively (Muscutt et al. 1993). In a study of woody and grassy riparian zones
separating agricultural fields from both Carroll Creek and the Speed River, in
southern Ontario, nitrate concentrations in shallow groundwater were essentially
100% depleted, with most of the decrease occurring within the first 20 to 30 m of the
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riparian zone (Martin et al. 1999). Woody riparian zones appear to be slightly more
effective than grassy ones at removing nitrate from groundwater (Martin et al. 1999).
Some of the nitrate removal that occurs beneath buffer strips is due to root uptake of
the nitrate by vegetation. The vegetation also increases nitrate removal indirectly by
providing a carbon source for anaerobic microbial denitrification in the root zone
(Gold et al. 1999). In geologically recent groundwater reserves, Spruill (2000) found
that there was a 95% reduction in nitrate levels in young groundwater beneath
vegetative buffer strips relative to groundwater in areas without buffer strips. Spruill
(2000) attributed approximately 70% of this difference to denitrification processes that
are facilitated by the higher levels of dissolved organic carbon (DOC) provided by the
decaying vegetation from the buffer strips. Using isotopic tracers, Mengis et al.
(1999) also confirmed that denitrification, as opposed to plant uptake, was the major
route for nitrate removal from groundwater flowing through a grassed buffer strip in
an agricultural watershed.
In locations where leaching of nitrate to deep groundwaters occurs, or where artificial
underdrainage has been constructed, buffer strips may be “underpassed” and thus
ineffective at preventing nitrate loss to streams (Cook 1999).
4.2.5 Microbial Transformation
In soils, nitrate is relatively stable except when biologically transformed by
denitrification. Denitrification, in which NO3- is converted by bacteria to gaseous
nitrogen, occurs under low oxygen or anaerobic conditions, and in the presence of a
carbon source. As such, it is most likely to occur in very wet soils, inside of soil
aggregates at high moisture content, or in other anaerobic microsites within the soil
(Jury and Nielsen 1989). It is an important soil process primarily in wetlands or after
spring snowmelt and heavy rainstorm events (Melillo et al. 1983; Post et al. 1985).
Unlike aquatic ecosystems, the role of denitrification in the nitrogen dynamics of
terrestrial ecosystems is relatively minor (Stoddard 1994).
4.3 Aquatic Processes
4.3.1 Physico-chemical Factors and Nitrogen Speciation
The predominant form of nitrogen present in a water body (Figure 4.1) is dependent
on a number of factors, including pH, temperature, oxygen availability, plant uptake,
and mineralisation rates of organic nitrogen (Johnes and Burt 1993). Because many
of these factors are largely a function of season, it can be said that season indirectly
controls the speciation balance of nitrogen in waters (Johnes and Burt 1993).
Dvir et al. (1999) examined the influence of pH on nitrogen speciation in a marine
model ecosystem. Variations in pH affected the rates of oxidation of ammonia to
nitrite and nitrate by nitrifying bacteria in the test vessels. Nitrate production rates
were similar at pH 7 and pH 8, but lower at pH 9. Overall, nitrification was optimal at
pH 8, resulting in greater nitrate+nitrite production rates (Dvir et al. 1999).
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Season not only influences nitrogen speciation, but also the total concentrations of
nitrogen that are present in surface waters. In fresh and marine waters, seasonal
variations in nitrate concentrations occur. Numerous researchers in northern
temperate climates have found that nitrate concentrations in fresh surface waters are
highest in the fall and winter months, particularly when there is greater precipitation
(Hill 1978; Neill 1989; Haycock and Burt 1993; Johnes and Burt 1993). In marine
waters, nitrate concentrations are also highest in the late fall and winter, largely due
to the breakdown of offshore stratification that results in the entrainment and mixing
of deep nutrient-rich waters into the surface layer (Louanchi and Najjar 2000). In
some nearshore coastal waters, runoff of nutrient-rich water from the land can also
contribute to higher nitrate concentrations in the fall and winter (Louanchi and Najjar
2000). Nitrate concentrations in marine waters are lowest in the spring and summer,
reflecting the greater biological uptake (Louanchi and Najjar 2000).
Schindler et al. (1971) demonstrated the influence of seasonal biological uptake on
nitrate concentrations in a whole-lake enrichment study in the Experimental Lakes
Area of northwestern Ontario. Weekly additions of 0.66 mg NO3-·L-1 resulted in water
column nitrate concentrations of up to 0.88 mg NO3-·L-1 in the early spring,
> 1.3 mg NO3-·L-1 in the fall and winter, but generally only 0.04 to 0.22 mg NO3-·L-1
were present in the productive late spring and summer months (Schindler et al.
1971).
Science-based Solutions No. 1-6
27
Loss to
Atmosphere
Wet
Deposition
N2
Surface Runoff
(DON)
(1)
Surface Runoff /
Soil Leaching
(DIN)
Sewage
Algae
(2)
(2)
Terrestrial
Plants
Organic N
(6)
(3)
(6)
N2
NH3
(4)
(6)
Groundwater
Seepage
NO3-
NO3
Aquatic
Plants
-
(4)
(5)
(5)
NO2-
(5)
(5)
N2O
Animals
(2)
(4)
(2)
Organic N
NO3-
NH3
Bacteria
N in Sediments
(5)
4 - Nitrification
1 - Fixation
2 - Decomposition / Excretion 5 - Denitrification
6 - Assimilation
3 - Ammonification
Transport / Uptake
(Adapted from NRC 1978)
Figure 4.1 Schematic representation of the nitrogen cycle emphasizing aquatic
transformations.
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28
4.3.2 Advective and Diffusional Movement Within a Water Body
In marine waters, nitrate concentrations are typically very low in the upper euphotic
zone due to rapid assimilation by phytoplankton. Movement of nitrate from deeper
waters to the surface may occur seasonally, or sporadically through upwelling and
mixing caused by surface cooling, wind, or other processes affecting thermal
stratification (Ryther and Dunstan 1971).
In examining lake-wide responses to manipulated nutrient levels, Levine and
Schindler (1989) found that nitrate levels within in-lake enclosures were about
0.3 mg NO3-·L-1, compared to enclosures with solid plastic bottoms
(< 0.02 mg NO3-·L-1), showing that in this shallow study lake with a mean depth
1.5 m, nitrate levels in the water column were directly affected by mobilisation from
the sediments.
Using 15N tracers, Peterson et al. (2001) determined that nitrate regeneration from
sediments of small (< 10 m wide) headwater streams contributed significantly to
inorganic nitrogen concentrations in the overlying water. Once in the stream, nitrate
molecules travelled approximately 5 to 10 times as far as ammonium molecules
before being assimilated by biota, or undergoing denitrification (Peterson et al. 2001).
4.3.3 Microbial Nitrification
Nitrification is a two-step microbial process by which ammonium is oxidised to nitrite
and then nitrate (Figure 4.1). This oxidation is primarily conducted by autotrophic
bacteria under aerobic conditions. Certain heterotrophic bacteria are capable of
carrying out nitrification, but at a much slower rate than autotrophic nitrification
(Verstraete and Alexander 1973; Bock 1978; Killham 1986; Wolfe et al. 1988). Fungi
are also known to carry out nitrification (Stoddard 1994). Other than nitrate formed
from nitrogen oxides in the atmosphere, nitrification is the sole natural source of
nitrate in the biosphere (NRC 1978).
In the first step of nitrification, ammonium is oxidised to nitrite (Wolfe et al. 1988):
NH4+ + 3/2 O2 ¤ NO2- + H2O + 2 H+
(∆G° = -272 kJ·mol-1)
The genera of bacteria most frequently associated with this step are Nitrosomonas,
Nitrosolobus, Nitrosococcus, Nitrosovibrio, and Nitrospira (Watson et al. 1981).
During the production of NO2- from NH4+, several intermediate products are formed,
including hydroxylamine (NH2OH), pyruvic oxime (N2H2O2) and nitrous acid (HNO2)
(Wetzel 2001). Nitrous oxide (N2O) can subsequently be produced from the
breakdown of NH2OH (Kaplan1983).
The second step involves the oxidation of nitrite to nitrate (Wetzel 2001):
NO2- + ½ O2 ¤ NO3-
(∆G° = -75 kJ·mol-1)
This process is carried out primarily by members of the genus Nitrobacter.
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29
As there is more free energy liberated per mole of NH4+ than NO2- during the
nitrification process, Nitrosomonas obtains more energy per mole of nitrogen oxidised
than Nitrobacter. Maximum growth rates, however, for Nitrobacter (0.8 day-1) are
much greater than Nitrosomonas at 20°C (0.5 day-1), and therefore the intermediate
nitrite form will not accumulate in large amounts as it is generally oxidised as rapidly
as it is formed (NRC 1978; Halling-Sorensen and Jorgensen 1993). Nitrification is a
strongly acidifying process, producing two moles of hydrogen ions for each mole of
ammonium that is nitrified. This oxidation process can also be costly to oxygen
budgets in surface waters, as 4.57 mg O2 are consumed per mg NH4+-N oxidized to
NO3--N.
Temperature, dissolved oxygen, and pH have all been found to affect rates of
nitrification (Dvir et al. 1999). Most strains of nitrifying bacteria grow optimally at a pH
of 7.5-8.0, warm temperatures of 25-30°C, and in darkness (Watson et al. 1981;
Alleman et al. 1987; Wolfe et al. 1990). Submersed macrophytes can enhance rates
of nitrification in the water column by providing a substrate for epiphytic communities
of microbial nitrifiers (Eriksson and Weisner 1999). Nitrification in epiphytic
communities is greater in light than in dark, presumably due to increased oxygen
concentrations at macrophyte surfaces produced during photosynthesis (Eriksson
and Weisner 1999). The presence of macrophytes may also stimulate nitrification in
sediments through the release of oxygen from their roots into sediments that might
otherwise be anoxic (Iizumi et al. 1980). Bioturbation by benthic invertebrates may
also enhance nitrification rates in sediments (Seitzinger 1988). Nitrifying organisms
are typically slow-growing.
The rates of nitrification per unit volume that occur in sediments are typically at least
an order of magnitude greater than nitrification rates in the water column (Seitzinger
1988). For example, Kaplan (1983) found that typical nitrification rates in coastal
sediments were 0.28 mg N·L-1·h-1, whereas in coastal waters nitrification rates were
generally less than 0.014 mg N·L-1·h-1.
Half-saturation constants for Nitrosomonas range from 0.2 to 8.0 mg NH4+-N·L-1,
whereas phytoplankton range from 1.4 to 140 µg NH4+-N·L-1 (NRC 1978). As growth
rates between the two types of organisms are similar (i.e., 1 to 3 doublings per day),
and because ammonia concentrations in the euphotic zone of lakes and oceans are
typically less than 100 µg·L-1, phytoplankton can outcompete the nitrifying bacteria for
ammonia (NRC 1978).
During the early summer, following stratification, nitrification in the hypolimnion of
lakes can consume a significant amount of oxygen, and the resulting nitrate produced
is denitrified as the water becomes anoxic. This process of nitrification-denitrification
provides an important pathway for the ultimate removal of fixed nitrogen from surface
waters (NRC 1978).
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30
4.3.4 Microbial Denitrification
Denitrification (also known as dissimilatory reduction) occurs in the presence of
facultative heterotrophic bacteria under extremely low oxygen conditions (Kapoor and
Viraraghavan 1997; Dvir et al. 1999). In the absence of oxygen, anaerobic bacteria
use oxidized forms of nitrogen (e.g., NO3-, NO2-) as a terminal electron acceptor
during the oxidation of an organic substrate (e.g., methanol in MWWTPs or DOC in
surface and groundwaters) to produce gaseous forms of nitrogen, such as N2, that
are then lost to the atmosphere (Seitzinger 1988). Denitrification occurs along the
following pathway:
NO3- → NO2- → NO → N2O → N2
The two-step dissimilatory reduction process can be illustrated using methanol as the
electron provider (Halling-Sorensen and Jorgensen 1993):
1) NO3- + 1/3 CH3OH → NO2- + 1/3 CO2 + 2/3 H2O
2) NO2- + 1/2 CH3OH → 1/2 N2 + 1/2 CO2 + 1/2 H2O + OHCertain genera of bacteria such as Pseudomonas, Micrococcus and Bacillus first
reduce nitrate to nitrite and then subsequently to two intermediates (NO and N2O)
before being lost to the atmosphere as N2 gas (Halling-Sorensen and Jorgensen
1993).This denitrification process provides an important pathway for nitrogen
removal. In almost all inland and coastal ecosystems, more nitrogen is lost via
denitrification than is gained through direct N2 fixation (Seitzinger 1988).
In both freshwater and marine systems, an oxygen concentration of about 0.2 mg·L-1
or less, or an electron activity level (pE) of ~ 10 - 14 is required for denitrification in
water or sediment (Seitzinger 1988; Hemond and Fechner 1994). Open-water
denitrification may occur, but bottom sediments are the main site for denitrification in
aquatic systems (Keeney et al. 1971; Seitzinger 1988).
Both heterotrophic bacteria (e.g., Pseudomonas nitrificans) and autotrophic bacteria
(e.g., Thiobacillus denitrificans, Micrococcus denitrificans) are capable of
denitrification by using nitrate as a terminal electron acceptor in place of oxygen in
the respiratory process (Halling-Sorensen and Jorgensen 1993; Kapoor and
Viraraghavan 1997). Heterotrophic bacteria require a carbon source from organic
substrates such as methanol, ethanol or acetic acid to provide an electron donor for
the reduction process; however, autotrophic bacteria can use hydrogen or reduced
sulfur compounds (Kapoor and Viraraghavan 1997). In shallow groundwater,
dissolved organic carbon (DOC) provides the energy source for bacteria, and
elevated levels of DOC are associated with increased denitrification (Spruill 2000).
Anaerobic bacteria in sediments may also reduce nitrate to ammonium under the
appropriate conditions, utilizing a portion of that ammonium as a nitrogen source for
growth (Hattori 1983). The most efficient denitrification pathway for the bacteria which
Science-based Solutions No. 1-6
31
are reducing nitrate in the presence of H2 is dependent on limiting amounts of
substrates. When organic matter (the electron donor in the reaction) is limiting, the
ultimate production of N2 is more energetically favorable (as amount of free energy).
However, when nitrate levels are limiting and organic matter is abundant, the
reduction of nitrate to ammonium would be more advantageous (Hattori 1983).
Although dissimilatory reduction is a possible pathway for ammonium production, the
primary source of ammonium in aquatic systems is via waste products from the
breakdown of organic matter (i.e., the deamination of proteins, urea, amino acids,
etc.) by heterotrophic bacteria (Wetzel 2001).
Numerous researchers have found that denitrification rates increase with increasing
temperature (Cavari and Phelps 1977; Messer and Brezonik 1984). Other factors that
affect rates of denitrification in aquatic systems include oxygen concentration, and
the supply of nitrate and organic matter (Seitzinger 1988). Denitrification rates
reported for freshwater lake and river sediments range from 0 to 4.8 mg N·m-2·h-1
(Seitzinger 1988). The reported range of denitrification rates for coastal marine
sediments is greater, ranging from 0 to 14.9 mg N·m-2·h-1, but are most commonly
between 0.7 and 3.5 mg N·m-2·h-1 (Seitzinger 1988).
Christensen et al. (2000) found that denitrification in marine waters occurred in late
autumn when NO3- levels in the water column were high. During the summer months,
little denitrification occurred based on water column NO3-; however, denitrification did
occur during this time in the sediments based on nitrate produced through nitrification
in the sediments (Christensen et al. 2000).
Through the use of wetland mesocosm studies, Crumpton et al. (1993) found that
nitrate concentrations decreased rapidly in water overlying wetland sediments, even
under highly aerobic conditions. With initial NO3- application rates of approximately
12 and 33 mg NO3-·L-1, the nitrate was completely removed from the water in 3 and
5 days, respectively. It is likely that the nitrate removal was due, in part, to both
microbial denitrification and assimilation by macrophytes growing in the mesocosms.
4.3.5 Biotic Assimilation
Assimilatory nitrate reduction is the process by which plants (including phytoplankton)
and a number of aerobic bacteria and fungi endogenously reduce nitrate to
ammonium that then provides the nitrogen source for the synthesis of cellular
materials (Hattori 1983). Aquatic plants will preferentially take up NH4+ because it is
more energetically favourable than NO3- (Stoddard 1994). Nonetheless, large
quantities of nitrate may be removed from surface waters through assimilation by
algae and macrophytes (Johnes and Burt 1993). Diatoms, for example, have been
observed to actively accumulate nitrate so that the internal concentration within their
cells is more than 100 times higher than concentrations in the surrounding medium
(Cresswell and Syrett 1981). Large nitrate removals from a stream in central Ontario,
as a result of assimilation, were observed by Devito and Dillon (1993). Their study of
a beaver pond located along the stream showed that annual inputs of NO3- to the
pond exceeded outputs, whereas annual outputs of organic nitrogen from the pond
Science-based Solutions No. 1-6
32
exceeded inputs, suggesting transformation through biotic assimilation (Devito and
Dillon 1993).
In temperate zones, assimilation rates vary with season, and consequently nitrate
levels will also vary seasonally. Hunt et al. (1995) found that establishment of an instream wetland was effective at the removal of nitrogen from the stream water in
warmer months. Summer concentrations of nitrate immediately downstream from the
wetland site dropped from 5.5 mg NO3-·L-1, prior to establishment of the wetland, to
1 mg NO3-·L-1 or less (Hunt et al. 1995). The wetland was less effective at nitrate
removal during cooler months, presumably due to slower denitrification and less plant
growth.
Diurnal patterns have also been observed for rates of assimilatory nitrate reduction.
In a marine tank system containing seaweed, nitrate levels in the water were found to
increase during the day, but decrease at night (Dvir et al. 1999). Diurnal nitrate
fluctuations have also been observed in the Neversink River, New York, but with an
opposite trend. Nitrate concentrations decreased during the day due to uptake by
photoautotrophs that were actively photosynthesizing; nitrate concentrations in the
water increased during the night, peaking in the early morning before sunrise (Burns
1998). These results are supported by a study in which a lack of nitrate uptake by
diatoms was observed when the culture was incubated in darkness (Cresswell and
Syrett 1981). Also, when the diatoms were exposed to light but aerated with CO2-free
air, nitrate uptake was inhibited (Cresswell and Syrett 1981). The authors speculated
that nitrate uptake requires a supply of ATP from either photophosphorylation or
oxidative phosphorylation.
According to Stumm and Morgan (1981), the form of nitrogen assimilated by aquatic
autotrophs will strongly influence the chemistry of surrounding waters. When nitrate is
used as the nitrogen substrate, more oxygen is produced in the surrounding water
than with the ammonium ion, which can result in super-saturated conditions (Crouzet
et al. 1999). Similarly, alkalinity will also increase with nitrate assimilation due to the
consumption of H+ (Crouzet et al. 1999). This is demonstrated in the following
equations:
106 CO2 + 16 NO3- + HPO42- + 122 H2O + 18 H+
photosynthesis
respiration
(C106H263O110N16P1) + 138 O2
“algae”
106 CO2 + 16 NH4+ + HPO42- + 108 H2O
photosynthesis
respiration
(C106H263O110N16P1) + 107 O2 + 14 H+
“algae”
(after Crouzet et al. 1999)
Yamaguchi and Itakura (1999) found that out of 26 different forms, or sources of
inorganic and organic nitrogen, the dinoflagellate Gymnodium mikimotoi showed the
greatest yield and growth rates when supplied with nitrate or nitrite. The authors
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33
speculated that the high concentrations of ammonia and urea used in the assays
(250 µM) may have inhibited the dinoflagellates, whereas these nitrogen species
might be used more in the natural environment where they would occur at lower
concentrations. The diatom Phaeodactylum tricornutum was observed to actively
take up nitrate, but this uptake was inhibited in the presence of ammonium
(Cresswell and Syrett 1981).
Eventual decomposition of biota will release organically bound nitrogen to the water
again where it will be mineralised to ammonium, and if the waters are sufficiently
oxic, will be oxidised to nitrate (Johnes and Burt 1993).
Microbial assimilation also occurs, in which nitrate is reduced to ammonia and
incorporated into organic compounds, such as amino acids, that may subsequently
be used in the production of nucleic acids and proteins (Brezonik 1975). The general
pathway for bacterially mediated assimilatory nitrate reduction is:
NO3- → NO2- → X (unknown ) → NH2OH → Organic N
(Halling-Sorensen and Jorgensen 1993)
4.3.6 Movement From Water to Sediments
Most of the nitrate found in sediments is produced in situ through the biodegradation
of organic matter to NH4+ that is then oxidized to NO3- (Seitzinger 1988). Smaller
quantities of nitrate, however, may enter the sediments from the water column.
Christensen et al. (2000) examined the flux of nitrate across the marine sedimentwater interface at locations below fish farm cages and at reference sites. No
significant differences were observed between the two types of sites, and generally
there was only a minor influx to the sediments of < 62 mg NO3-·m-2·d-1.
Stammers et al. (1978) found that sediment with a high organic matter content was
very effective as an agent for the removal of nitrate from stream water, with removal
occurring through denitrification. Within reduced sediments, anaerobic bacteria, such
as sulfate-reducing bacteria, may reduce nitrate to ammonium (Christensen et al.
2000).
4.3.7 Exchanges Between Surface Waters and Groundwater
Nitrate in surface waters can move downwards through sediments and the hyporheic
zone into groundwater. The hyporheic zone is a biologically active subsurface
ecotone between the surficial streambed and groundwater, where surface and
subsurface waters may mix. The downward movement of surface water into the
hyporheic zone occurs where the altitude of the water table is lower than the stream
or lake water surface; within streams this is typically at the head of riffles (Winter et
al. 1998; Biksey and Brown 2001). Downward movement of nitrate to groundwater is
Science-based Solutions No. 1-6
34
largely controlled by hydraulic recharge/discharge processes. Therefore, factors that
affect groundwater recharge rates, such as the permeability of surface water
sediments, can also influence the movement of nitrate. For example, Grimaldi and
Chaplot (2000) observed downstream decreases in nitrate concentrations, with loss
to the underlying groundwater, for a stream flowing on granite, but not on schist. On
granite, exchanges with the hyporheic zone were favoured by coarse-grained
sediments with a high permeability, whereas on schist the grain-size distribution is
much finer and permeability is reduced, thus preventing exchanges between surface
and subsurface waters (Grimaldi and Chaplot 2000). Downwelling zones are
characterized by high oxygen levels and aerobic processes (Biksey and Brown
2001); therefore the production of nitrate through nitrification is likely to occur in these
areas.
Movement of nitrate can also occur in the opposite direction, with seepage of
groundwater up into surface water bodies. Discharge and upwelling of groundwater
occurs where the altitude of the water table is higher than the stream or lake water
surface, such as at the base of pools within streams (Winter et al. 1998; Biksey and
Brown 2001). Nitrate present in groundwater may be advected through freshwater
sediments (Keeney et al. 1971), or coastal marine sediments (Slater and Capone
1987). In temperate regions, the greatest flux of nitrate from groundwater to surface
waters occurs in the spring. For example, the spring that feeds Swifts Brook, a small
headwater stream within the Grand River Watershed of Southern Ontario, has its
highest concentrations of nitrate during peak flow rates in March and April, and
lowest nitrate concentrations in October or November following the periods of lowest
flow (August or September) (Stammers et al. 1978). The movement of chemical
constituents, such as nitrate, between groundwater and surface water is affected by
biogeochemical processes in the hyporheic zone (Winter et al. 1998). Upwelling
zones are characterized by anoxic conditions and anaerobic processes (Biksey and
Brown 2001); therefore, much of the nitrate present in discharged groundwater will
likely undergo denitrification within this zone. Tobias et al. (2001) tracked the fate of
15
N-labelled nitrate that had been introduced into a groundwater plume upgradient of
a salt marsh in Virginia. Up to 90% of the groundwater nitrate load discharging into
the marsh was reduced rapidly in the upper 10 cm of sediment. Denitrification
(primarily to N20) accounted for 70% of the total nitrate loss rate, and the other 30%
was due to dissimilatory nitrate reduction to ammonium (Tobias et al. 2001). Another
study using nitrogen isotope tracers compared the fate of groundwater nitrate in two
different drainage basins in Maryland (Böhlke and Denver 1995). The groundwater
nitrate concentrations in the two basins were similar when recharged, but the basins
differed in terms of the depths at which reducing sediments occurred. Lower nitrate
concentrations were observed in groundwater discharges to the stream where the
reducing sediments were shallower because a larger fraction of the groundwater was
able to pass through those sediments, and therefore more denitrification took place
(Böhlke and Denver 1995).
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4.3.8 Anthropogenic Nitrate Removal from Ground and Surface Waters
Nonpoint sources of nitrate (such as leaching and surface runoff from agricultural
land, and urban stormwater runoff) pose the greatest source of contamination to
surface waters (NRC 1978). Nitrate reaching surface waters can subsequently be
consumed by vegetative uptake (algae and macrophytes), denitrification, and
assimilation by microorganisms (Laposata and Dunson 1998). Efforts to remove
nitrate before entering receiving waters in agricultural areas can include the use of
vegetative buffer strips to assimilate nitrate from shallow groundwaters and runoff
(see section 4.2.2), reducing field slopes to slow runoff and facilitate greater
biological uptake, and by collecting and treating runoff from feedlots and crop fields in
holding ponds (NRC 1978). Other measures for reducing nitrate export from
agricultural land include the use of zero tillage to reduce erosion and runoff, planting
of perennial forages in marginal areas, and encouraging grassed waterways. Fencing
off access for livestock to waterways assists in the regeneration of plant growth, and
increases habitat availability for littoral aquatic species (Magilligan and McDowell
1997).
There are several biological, physical, and chemical processes available for the
removal of nitrogen from point source discharges such as MWWTPs (Table 4.2).
Biological denitrification is the most commonly used technique to remove nitrate from
municipal and industrial wastewaters before they are released into receiving waters
(NRC 1978; Kapoor and Viraraghavan 1997). This involves a two-step process that
can be carried out in conjunction with secondary or tertiary waste treatment, whereby
wastewater is first oxygenated to convert any ammonia-nitrogen present to nitrate
using nitrifying bacteria, followed by denitrification with heterotrophic bacteria under
anoxic conditions and a readily usable carbon energy source (e.g., methanol) to
reduce nitrate to nitrogen gas (N2) (Halling-Sorensen and Jorgensen 1993). Nitrate
removal efficiency using this process ranges from 80-90%; however, the second step
involving denitrification is less efficient at ambient temperatures < 6°C and in the
presence of dissolved oxygen (Kapoor and Viraraghavan 1997). In a study on the
removal of nitrate from dairy wastewaters, Zayed and Winter (1998) found that a
mixed bacterial culture was able to completely denitrify loads of 4000 mg NO3-·L-1·d-1
for 15 days using existing organic compounds as electron donors, suggesting that
more costly methanol-addition operations may not be necessary for all applications.
Reactive barriers have been investigated as a low-cost, low-maintenance method for
in situ removal of nitrate from septic systems or farm field drainage (Robertson et al.
2000). These barriers, which consist of waste cellulose solids such as wood mulch,
sawdust and leaf compost, reduce nitrate levels by providing a carbon source for
heterotrophic denitrification. Under varying conditions, the reactive barriers can result
in nitrate removal rates ranging from 3 to 142 mg NO3-·L-1·d-1 (Robertson et al. 2000).
Non-biologically mediated denitrification techniques include ion exchange, reverse
osmosis and electrodialysis (Table 4.2). Ion exchange resin beds substitute nitrate
ions from contaminated water with chloride or bicarbonate ions until the resin’s
exchange capacity is exhausted, at which point the resin must be regenerated
(Kapoor and Viraraghavan 1997). Ion exchange has been shown to be effective for
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36
the removal of nitrate from groundwater, drinking water, agricultural subsurface
drainage, and activated sludge plant effluent; however, the ion exchange efficiency is
reduced from the presence of organic matter and by competition with SO42- (Eliassen
et al. 1965; Magette et al. 1990; Halling-Sorensen and Jorgensen 1993; Kapoor and
Viraraghavan 1997). Close to 100% nitrate removal is possible through ion exchange
(Clifford and Liu 1993). Ion exchange can also be used in combination with biological
denitrification of the spent brine to reduce the salt consumption and waste discharge
(van der Hoek et al. 1988; Clifford and Liu 1993).
The reverse osmosis process excludes ions by forcing water across a
semipermeable membrane at pressures exceeding the ionic species’ osmotic
pressure. Water is forced through cellulose acetate or polyamide membranes at
pressures ranging from 2070 to 10 350 kPa (Kapoor and Viraraghavan 1997). Such
high pressures require a greater expenditure of energy, resulting in much larger
operating costs than ion exchange (Kapoor and Viraraghavan 1997).
Electrodialysis is another membrane separation technique that uses a direct electric
current to transfer ions from a less concentrated to a more concentrated solution
through a semipermeable membrane. This process is not very widely used for nitrate
removal as it is also costly, works only for soft waters, and requires considerable
pretreatment of the influent to remove organics (Kapoor and Viraraghavan 1997).
Although NO3- stripping through resin columns is widely available, global drinking
water treatment processes are generally not equipped to remove nitrate, and as
such, drinking water concentrations frequently contain nitrate levels similar to that of
source waters (Heathwaite et al. 1996).
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Table 4.2. Selected wastewater treatment processes for nitrate removal.
Treatment process
Biological
Denitrification using
methanol (following
nitrification stage)
Physical/Chemical
Ion exchange
Reverse osmosis
Electrodialysis
% Removal of nitrogen form
Organic N
NH3/NH4+
NO3-
Process advantages
Process disadvantages
-
-
80 - 90
• rapid denitrification
• high degree of nitrogen removal
possible
• up to 3 weeks for start-up
• methanol required
• high operational space
requirements
slight
slight
75-90
• immediate start-up
• not influenced by climatic
conditions (i.e., low
temperatures)
• low TDS in effluent
• ease of product quality control
60 - 90
60 - 90
60 - 90
100
(suspended
organic
nitrogen)
30 - 50
30 - 50
• simultaneously removes all
forms of nitrogen
• large amounts of nitrogen
removed
• not affected by lower
temperatures
• simultaneously removes all
forms of nitrogen
• pre-treatment by filtration
required
• organic matter and other anions
reduce efficiency
• disposal of regeneration
material (brine)
• higher capital costs
• requires highly skilled operator
• membrane elements easily
fouled by colloidal material
• pre-treatment of secondary
effluent required
• high maintenance
Science-based Solutions No. 1-6
• precipitation of salts on
membrane surface
• clogging of membrane from
residual colloidal organic matter
• ~10% of feed volume required
to continuously wash
membrane
38
5. ENVIRONMENTAL CONCENTRATIONS
5.1 Nitrate Levels in Precipitation
Atmospheric deposition can provide a substantial route for nitrate contamination of
surface and groundwaters, especially in urban areas with little ground cover or
natural vegetation to take up the deposited nitrate and ammonium that can then
accumulate in groundwater (via leaching processes), or in surface waters as a result
of runoff (Rouse et al. 1999). Areal estimates of nitrate deposition vary widely across
Canada. Annual total deposition (dry + wet) of nitrate at the Abbotsford Aquifer,
British Columbia is estimated at 192 mg NO3-·m-2·a-1 (= 1.92 kg NO3-·ha-1·a-1)
(McGreer and Belzer 1999). Meteorological sampling between 1995 and 1998
suggested that 9.2 (± 1.6) kg N·ha-1·a-1 were being deposited in Lake Simcoe,
Ontario (Winter et al. 2002). In general, atmospheric deposition of NO3- and NH4+ is
greater in Eastern Canada, with a ten-year average for 1984-1994 of
3.44 kg N·ha-1·a-1 occurring east of the Manitoba-Ontario border, compared to
0.80 kg N·ha-1·a-1 west of the border (Chambers et al. 2001).
Heidorn (1979) showed a link between days with high nitrate deposition in suspended
particulate matter (> 9.9 µg·m-3) in the Southern Ontario corridor and high-pressure
systems originating from south of the lower Great Lakes area. These periods of
higher nitrate concentrations were biased towards colder months when greater
quantities of NOx gases are released due to larger energy demands (e.g., space
heating). Nitrate is then formed from the nitrogen oxides collected in air masses over
the Great Lakes, and is precipitated out of the atmosphere, with deposition
decreasing as distance from the Great Lakes increases (Heidorn 1979). Nitrate
levels in precipitation around the heavily populated Great Lakes often exceeds
2 mg NO3-·L-1, resulting in loading estimates in excess of 20 kg NO3-·ha-1·a-1 for this
region, compared to less than 1 kg NO3-·ha-1·a-1 for more remote locations such as
Snare Rapids in the Northwest Territories (Ro et al. 1995; CNACD 2001).
Volume-weighted concentrations of nitrate in precipitation of the Muskoka-Haliburton
region from 1976 to 1986 ranged from ~1.9 to 2.5 mg NO3-·L-1 (Dillon et al. 1988). In
2000, annual weighted-mean nitrate concentrations from selected Canadian Air and
Precipitation Monitoring Network (CAPMoN) locations ranged from 0.25 mg NO3-·L-1
in Snare Rapids, Northwest Territories to 2.23 mg NO3-·L-1 in Longwoods, near Lake
Erie, Ontario (CNACD 2001).
5.2 Environmental Levels in Surface Waters
5.2.1 Freshwater
Inorganic nitrogen is the predominant form of nitrogen in surface waters, of which
nitrate is the most abundant form in well-oxygenated systems (Wetzel 1983). In
general, nitrate-nitrogen constitutes two-thirds to four-fifths of the total available
nitrogen in surface waters (Crouzet et al. 1999).
Science-based Solutions No. 1-6
39
Nitrate levels in Canadian lakes and rivers rarely exceed 4 mg NO3-·L-1 (Table 5.1).
In oligotrophic lakes and streams, where primary productivity is low, nitrate
concentrations are generally < 0.4 mg NO3-·L-1 (NRC 1978; Nordin and Pommen
1986). High nitrate concentrations (i.e., exceeding 4 mg NO3-·L-1) tend to be
associated with eutrophic conditions and waters experiencing algal blooms (NRC
1978). In the U.S., stream nitrate concentrations above the national background level
of 2.7 mg NO3-·L-1 are considered to have been affected by human activities (USGS
1999). In a study of streams in agricultural regions of Alberta in 1996, flow-weighted
mean nitrate concentrations were 5.3 mg NO3-·L-1 in regions of high agricultural
intensity compared to 0.10 mg NO3-·L-1 in low intensity regions (Anderson et al.
1998). Reference values reported by the European Environment Agency for nitrate in
non-impacted European rivers range from 0.4 to 4.4 mg NO3-·L-1 (Crouzet et al.
1999). In Canada, average 1990 nitrate levels in raw (pre-treated) municipal water
supplies ranged from 0.1 to 3.3 mg NO3-·L-1 (Government of Canada 1996).
Correlations often exist between nitrate concentrations in a waterbody and factors
such as human population growth, or the percentage of a catchment that has been
altered by anthropogenic land uses (Rhodes et al. 2001). High nitrate levels have
been noted in surface waters as a result of various human activities. In areas
downstream of open pit coal mining operations, explosives residues result in elevated
nitrate concentrations (Nordin and Pommen 1986). Inorganic fertilizer use in rural
areas can also result in excessive localized nitrate levels. Mean nitrate
concentrations of North American streams in agricultural landscapes generally range
between 9 and 180 mg NO3-·L-1, and levels above 45 mg NO3-·L-1 can persist for
several weeks (Rouse et al. 1999; Castillo et al. 2000). Irrigation water used in crop
fertilization studies carried out on a Nebraska farm contained nitrate concentrations
of 93 mg NO3-·L-1 (Eghball and Gilley 1999). Sewage treatment plants may also
contribute to elevated nitrate levels; concentrations ranging from 19 to 42 mg NO3-·L-1
found in the Cootes Paradise wetland in Dundas, Ontario in 1997, were primarily
attributed to anthropogenic loading from a local sewage treatment plant (Rouse et al.
1999).
5.2.1.1 Seasonal Variation
Nitrate concentrations are seasonally variable, with increased biological uptake in
warmer productive months reducing ambient surface water concentrations. In 1983,
nitrate levels in the tributaries of the inner bay of Rondeau Provincial Park, on the
north shore of Lake Erie, declined between winter and spring (range: 31 to
58 mg NO3-·L-1) and summer (18 mg NO3-·L-1) (OMOE 1983). On the St. Lawrence
River downstream of the Montreal Archipeligo, higher nitrate concentrations in the
winter and spring (1.90 and 1.55 mg NO3-·L-1, respectively),
Science-based Solutions No. 1-6
40
Table 5.1. Representative nitrate concentrations in Canadian ambient surface
waters.
Province/Territory
Water Body
[NO3-]
(mg NO3-·L-1)‡
Reference
FRESHWATER
British Columbia
Arrow Lake
Thompson River
Flathead River
0.69
1.37
0.09
(< DL - 0.487)
NLET (2000)*
NLET (2000)*
McDonald et al. (1987)*
Alberta
Athabasca River
various Boreal
Plains headwater
lakes (wetland
dominated)
various Boreal
Plains headwater
lakes (upland
dominated)
0.27
0.05
(0.005 - 0.44)
NLET (2000)*
Prepas et al. (2001)*
0.02
(0.005 - 0.06)
Prepas et al. (2001)*
Saskatchewan
Battle River
0.190 - 0.602
NLET (2000)*
Manitoba
Assiniboine River
1.27
(< DL - 14.2)
0.24
(< DL - 0.93)
0.59
(< DL - 21.5)
Manitoba Conservation
(2000)*
Manitoba Conservation
(2000)*
Manitoba Conservation
(2000)*
Lake Winnipeg
Red River
Ontario
Ausable River
Grand River
Lake Ontario
Lake Superior
Mississagi River
Turkey Lakes
Quebec
Richelieu River
St. Lawrence River
Trois-Rivières
New Brunswick
Miramichi River
Nova Scotia
Gold River
28.5
(4.7 - 86.4)
13.4
(1.22 - 29.0)
1.46 - 2.04
1.38
0.56
(0.02 [trace] - 1.57)
0.252 - 61.0
13.7
1.13
(0.22 - 7.93)
0.31 - 0.66
0.920
0.02
(< DL - 0.09)
Science-based Solutions No. 1-6
OMOE (2001)*
OMOE (2001)*
NLET (2000)*
Bennett (1982)
OMOE (2001)*
NLET (2000)*
NLET (2000)*
Hudon and Sylvestre
(1998)*
NLET (2000)*
NLET (2000)*
Dalziel et al. (1998)*
41
Province/Territory
Water Body
[NO3-]
(mg NO3-·L-1)‡
Reference
Table 5.1 continued
Annapolis River
2.3
(0.67 - 6.49)
0.04
(< DL - 2.22)
Dalziel et al. (1998)*
Great Slave Lake
(western basin)
0.46
(0.41 - 0.85)
Evans (1997)*
coastal waters
Bay of Fundy
(depth: 0 - 5 m)
Bay of Fundy
(depth: 100 - 275 m)
< DL to 0.37
0.41
(0.01 to 1.12)
0.76
(0.26 to 1.34)
Keizer et al. 1996)
Petrie et al. 1999)
various lakes
Northwest Territories
MARINE
Nova Scotia
British Columbia
coastal waters summer
coastal waters winter
off-shore waters
(depth: 0 - 100 m)
NSDEL (2001)*
Petrie et al. 1999)
0.11
(~0 to 0.31)
1.1 - 1.7
Ahn et al. 1998)
0.9
(0.5 to 1.7)
Whitney 2001)
Whitney 2001)
note: NLET (2000) samples represent median values from a large, interlaboratory quality assurance study (n = 20 - 50); OMOE
(2001) data from 1996-2000; Manitoba Conservation (2000) data from 1980-2000; < DL = below detection limit, i.e, < 0.02
- -1
- -1
- -1
mg NO3 ·L for OMOE (2001), < 0.04 mg NO3 ·L for Manitoba Conservation (2000), and < 0.006 mg NO3 ·L for Dalziel et al.
‡
(1998); concentrations are means, with ranges indicated in brackets; * - concentrations are reported as NO2 + NO3 , but
are considered to consist entirely of NO3 as NO2 concentrations were not detected in surface water samples (Alkema 2000).
N/A = not available (i.e., a description of the watershed use was not provided in the reference document)
declined in the summer and autumn months to 0.84 and 1.06 mg NO3-·L-1,
respectively, due to greater biological productivity and nitrate uptake (Hudon and
Sylvestre 1998).
5.2.1.2 Temporal Trends
Contrary to decreasing trends in other nutrients, such as phosphorus, which have
been specifically targeted for removal from municipal sewage treatment plants, there
has been a general increasing trend in nitrate levels in the surface waters of the
Great Lakes. Comparison of mean spring and summer nitrate levels in the western
basin of Lake Erie between 1983-87 and 1989-93 showed significant increases from
2.53 to 3.54 mg NO3-·L-1 (Makarewicz et al. 2000). Ranges in spring and summer
nitrate concentrations in the Grand River, Ontario, which empties into Lake Erie have
also increased from 0.22 - 2.8 mg NO3-·L-1 in 1966 to 0.04 - 18.0 mg NO3-·L-1 in 1994
(Rott et al. 1998). Caution should be used in interpreting trends from this data for the
Grand River, however, as the authors only compared two years; nitrate levels can
Science-based Solutions No. 1-6
42
vary considerably on a year-to-year basis. Mean lake-wide spring surface
nitrate + nitrite concentrations in Lake Superior have been increasing from
~1.17 mg NO3-·L-1 in 1970 to ~1.56 mg NO3-·L-1 in 1992, with a predicted increase of
0.014 mg NO3-·L-1·a-1 (Williams and Kuntz 1999). Mean annual spring nitrate + nitrite
concentrations in Lake Ontario have also been steadily increasing from 0.95 (±
0.075) mg NO3-·L-1 in 1968 to 1.74 (± 0.035) mg NO3-·L-1 in 1993 (Williams et al.
1998a). Nitrogen-to-phosphorus ratios (N:P) are increasing in Lake Ontario (ratios,
expressed on a weight basis, currently range from 36 to 40) due to decreasing
phosphorus and increasing nitrogen concentrations. This could result in changes to
Lake Ontario’s algal species composition, as the prevalence of cyanobacterial
dominance tends to decrease at N:P ratios > 29 (by weight), and are replaced by
diatoms and chlorophytes (Williams et al. 1998a [cf Smith 1983]; CCME 2002). In
contrast with observations from the Great Lakes, mostly downward trends have been
observed for nitrate concentrations in Québec rivers and streams for the period from
1988 to 1998 (MENV 2001b).
International estimates for nitrate concentrations in surface waters are generally
consistent with Canadian levels. Increasing trends have also been observed in lakes
surrounded by intensive agricultural production in the English Lake District, with
mean nitrate concentrations increasing from approximately 1.8 mg NO3-·L-1 in 1945 to
6.2 mg NO3-·L-1 in 1980 (Heathwaite et al. 1996). Results from a European-wide
survey revealed that approximately 15% of rivers exceeded an annual average
concentration of 33 mg NO3-·L-1 between 1992 and 1996 (Crouzet et al. 1999).
Surface water samples from the Netherlands have been shown to range from 2.7 to
24.3 mg NO3-·L-1 (Brinkhoff 1978), and the average nitrate concentrations from 584
Norwegian lakes is 0.48 (± 0.46) mg NO3-·L-1 (maximum = 3.08 mg NO3-·L-1) (Bulger
et al. 1993). The much higher nitrate concentrations in the Netherlands’ waters,
compared to Norway, may be due in part to the higher population density and higher
percentage of agricultural land that is intensively farmed. For comparison, the
population densities of the Netherlands, Norway, and Canada are approximately 378,
14, and 3 persons per square km, respectively (based on data from Times Books
1999). The land base proportions that are used for agriculture in the Netherlands,
Norway, and Canada are approximately 53%, 3%, and 8%, respectively (World Atlas
2002).
In a review of water quality in U.S. rivers between 1974 and 1981, Smith et al. (1987)
reported trends of increasing nitrate concentrations at 116 monitoring stations versus
27 stations which showed decreasing nitrate trends. The majority of stations showing
nitrate increases were located in the eastern half of the country and were strongly
associated with agricultural activities. Total nitrogen loads delivered to the Gulf of
Mexico from intensive agricultural areas of the Mississippi Basin have increased
three-fold since the 1970’s with a mean annual nitrogen-flux for 1980 to 1996 being
1600 kt·a-1 (Goolsby et al. 2001). Nitrate-nitrogen accounted for 61% of the total
nitrogen, with the remainder being comprised of organic N (37%) and ammonium-N
(2%) (Goolsby et al. 2001).
Science-based Solutions No. 1-6
43
5.2.1.3 Spatial Trends
Longitudinal trends in nitrate concentrations for lotic waters are highly variable and
depend on site-specific factors such as catchment basin size (Johnes and Burt
1993), land-use activities (Rott et al. 1998; Van Herpe and Troch 2000), floodplain
lithology (Grimaldi and Chaplot 2000), and stream size, substrate composition and
geochemistry (Devito et al. 2000; Peterson et al. 2001). For example, longitudinal
gradients have been noted for Québec rivers flowing from the Appalachians and the
Laurentians through the St. Lawrence lowlands. Nitrate levels in headwaters were in
the range of 0.09 to 0.4 mg NO3-·L-1, whereas in the lowlands concentrations ranged
from 0.4 to 22 mg NO3-·L-1 (MENV 2001b).
During biologically productive seasons, standing waters consistently show lower
nitrate levels in the upper euphotic zones where the nitrate is readily assimilated by
phytoplankton and heterotrophic bacteria. Evans (1997) suggested that elevated
nitrate levels can also occur at greater depths due to nutrient regeneration; in study
sites > 60 m deep in the western basin of Great Slave Lake, NT, nitrate + nitrite
concentrations observed near the lake floor (up to 0.19 mg NO3-·L-1) were greater
relative to surface waters (~0.10 mg NO3-·L-1). Photoinhibition in surface waters of
light-sensitive nitrifying bacteria such as Nitrosomonas may also explain observed
increases in nitrification rates with depth (Hall 1986).
5.2.2 Marine
Although gaseous N2 is the most abundant species of nitrogen in ocean waters,
nitrate is the most abundant biologically-reactive form (Sharp 1983). Nitrogen
budgets for coastal marine waters indicate that more biologically available nitrogen is
lost through denitrification than gained through N2 fixation, resulting in an overall
ecosystem-level nitrogen deficiency (Paerl 1993). In contrast, microorganisms and
benthic invertebrates living within the sediments effectively recycle phosphorus in
coastal marine sediments and overlying water, which results in a nitrogen-limited
environment that is often reliant on external nitrogen inputs to maintain ecosystem
productivity (Paerl 1993).
Naturally occurring nitrate concentrations in temperate region seawater can reach up
to 2.4 mg NO3-·L-1 (Spencer 1975), the majority of which is due to nitrification
processes (Muir et al. 1991). Nitrate levels in European and North American
estuaries of rivers draining agricultural and urbanised areas can exceed
12 mg NO3-·L-1 (Sharp 1983). These concentrations tend to decrease when there is
increased mixing with more saline waters (Sharp 1983).
Nitrate levels from the central Scotian Shelf off the Canadian Atlantic coast follow
seasonal trends with the highest surface water concentrations (up to
0.535 mg NO3-·L-1) being found in the winter months (Petrie et al. 1999). By midspring, nitrate is largely depleted in the surface waters (~0.038 mg NO3-·L-1) due to
biological assimilation, with increasing concentrations (up to ~1.24 mg NO3-·L-1),
occurring beyond 30 m depth. Nitrate levels remain low throughout the summer
Science-based Solutions No. 1-6
44
(< 0.031 mg NO3-·L-1) and do not increase again until the late fall (Petrie et al. 1999).
Nitrate levels from two near-shore sampling locations in Nova Scotia from 1992 to
1994 ranged from below detection limits to 0.37 mg NO3-·L-1 (Keizer et al. 1996). On
the Canadian Pacific coast, nitrate levels tend to be higher in the winter months than
in the summer, and they also typically increase with depth (Whitney 2001). Nitrate
concentrations measured at various depths in February 2001, in the Strait of Georgia
(between Vancouver Island and mainland British Columbia) and in a transect running
west from the southwestern end of Vancouver Island, ranged from 0.18 to
2.9 mg NO3-·L-1 (Whitney 2001). Ambient nitrate levels in seawater near a salmon
farm in British Columbia were typically less than 0.31 mg NO3-·L-1 between May and
July (Ahn et al. 1998) (Table 5.1).
Nitrate levels in European marine waters are also generally below 1 mg NO3-·L-1 with
average concentrations reported for the U.K. at 0.44 to 0.88 mg NO3-·L-1, and from
the North Sea coast at ~0.4 to 1.0 mg NO3-·L-1 (Wickins 1976; van Duijvenbooden
and Matthijsen 1989).
5.3 Environmental Levels in Groundwater
Nitrate levels in groundwater are primarily a human health concern as well water
systems with elevated nitrate concentrations could pose a risk of
methaemoglobinaemia (see Section 6.2.2.1) to infants that do not have sufficient
gastric acids to control nitrate-reducing bacteria in their guts (Hill 1999). Groundwater
can, however, impact aquatic biota through discharge into streams and other surface
waters. Nitrate concentrations in groundwater tend to exceed those of surface waters
due to increased accumulation of nitrate leaching through soils under intense
agricultural and livestock production. Generally, up to 13 mg NO3-·L-1 can be found
naturally in groundwaters; levels above this indicate anthropogenic contamination
(Rouse et al. 1999).
Nitrate concentrations in well water in Canada can often exceed the guideline for
Canadian drinking water quality of 45 mg NO3-·L-1. In a summary of nitrate levels in
rural wells from each of the provinces, 1.5% to 64% of wells surveyed had greater
than 45 mg NO3-·L-1 (Chambers et al. 2001). Nitrate levels up to 1100 mg NO3-·L-1
have been reported in semi-arid regions of western Canada and the United States
(Rodvang et al. 1998). In 1991-1992, 14% of 1292 wells sampled in Ontario had
nitrate levels greater than 45 mg NO3-·L-1; this trend appears to have remained
relatively consistent from the 1950s (Goss et al. 1998a). During the 1980s and
1990s, mean groundwater concentrations in the Maritime provinces ranged from 8.9
to 132.9 mg NO3-·L-1, with up to 44% of dairy farm wells in Prince Edward Island
exceeding 45 mg NO3-·L-1 (AAFC 2000). Nitrate concentrations measured in
groundwater samples from Nova Scotia range from approximately 1 to
204 mg NO3-·L-1, with a mean that is likely less than 20 mg NO3-·L-1 (Moerman and
Briggins 1994). In western Canada, the Abbotsford aquifer, which spans southern
British Columbia and northern Washington State, is also dominated by agricultural
activity. Here, 54% of 117 domestic, municipal, and monitoring wells exceeded
Science-based Solutions No. 1-6
45
45 mg NO3-·L-1 in 1993, and it is estimated that 80% of all groundwater exceeds
40 mg NO3-·L-1 (Wassenaar 1994).
Reported groundwater nitrate concentrations from other international jurisdictions are
comparable to Canadian levels. In 53% of shallow groundwater studies in U.S.
agricultural and urban areas, median nitrate concentrations exceeded the U.S.
national background concentration estimate of 8.9 mg NO3-·L-1, and median
concentrations in 13 of 36 agricultural areas were > 22 mg NO3-·L-1 (USGS 1999).
These elevated nitrate levels in groundwater were strongly related to agricultural land
use and the widespread application of fertilizers in excess of crop uptake. Of thirtythree U.S. aquifers tested, the four that exceed the US EPA drinking water standard
of 10 mg NO3--N·L-1 (approximately 45 mg NO3-·L-1) were all shallow, composed of
sand and gravel and situated beneath agricultural areas (USGS 1999).
Similar inorganic fertilizer contamination of the shallow Sparta aquifer in Greece
resulted in 65% of samples exceeding the 50 mg NO3-·L-1 European drinking water
standards, with mean and maximum nitrate concentrations of 63 and
177 mg NO3-·L-1, respectively (Antonakos and Lambrakis 2000). High nitrate levels
persist in this aquifer due to the influx of large quantities of oxygenated water and the
presence of carbonate formations that resulted in strong oxidising conditions that
inhibit denitrification (Antonakos and Lambrakis 2000).
A province-wide survey of Ontario farmstead domestic wells illustrated that nitrate
concentrations in groundwater typically decrease exponentially with depth (Rudolph
et al. 1998). Likewise, in a review of factors influencing aquifer nitrate levels, Kolpin
et al. (1994) found a consistent decrease in the percentage of samples with nitrate
concentrations > 13 mg NO3-·L-1 with increasing aquifer depth (> 40 m below the
earth’s surface). Aquifers from areas of unconsolidated materials (i.e., glacially
deposited sand and gravel, or alluvium deposits) also had significantly higher nitrate
levels (p < 0.001) than those on sandstone, limestone or dolomite bedrock (Kolpin et
al. 1994). Kolpin et al. (1994) explain this difference as resulting from less lowpermeability material overlying the wells of the unconsolidated aquifers such that
contamination from the surface occurs more readily; also, in the unconsolidated
aquifers, the groundwater flow paths for recharge of the wells is shorter than in
bedrock, resulting in faster recharge rates.
Science-based Solutions No. 1-6
46
6. TOXICITY OF NITRATE TO AQUATIC ORGANISMS
Nitrate is considerably less toxic to aquatic organisms than ammonia or nitrite, with
acute median lethal concentrations of NO3--N being up to two orders of magnitude
higher than for NH3-N and NO2--N (Colt and Armstrong 1981). Nitrate is generally
considered to be of low toxicity to aquatic organisms due to its limited uptake and
absence of major physiological effects (Russo 1985; Jensen 1996).
There is a wide response in aquatic biota to nitrate exposure, both between
taxonomic groups, and between life stages. In general, based on acute median lethal
concentrations, amphibians and invertebrates are typically more sensitive than fish,
(though there are broad ranges in tolerance among species within each taxonomic
group). One to fifteen-day LC50 values for the nitrate ion in freshwater range from 73
to 7752 mg NO3-·L-1 for amphibians, from 24 to 3070 mg NO3-·L-1 for invertebrates
and from 847 to 9344 mg NO3-·L-1 for fish (Appendix A). For marine species, LC50
values for invertebrates range from 496 to > 19 840 mg NO3-·L-1, while those for fish
range from 2538 to 22 372 mg NO3-·L-1 (Appendix B). Nitrate concentration ranges at
which chronic effects occur are comparable for these three taxonomic groups.
Early life stages are more sensitive than juvenile or adult stages. While Westin (1974)
reported median nitrate lethal concentrations of 5800 and 6000 mg NO3-·L-1 for
chinook
salmon
(Oncorhynchus tshawytscha)
and
rainbow
trout
(Oncorhyncus mykiss), respectively, Kincheloe et al. (1979) found that concentrations
as low as 10 and 20 mg NO3-·L-1 could significantly increase egg and fry mortality in
these species. In addition, early instars of two net-spinning caddisfly species had
consistently lower LC50s when exposed to NaNO3 relative to late instar stages
(Camargo and Ward 1992).
There is very little information on the influence of water quality parameters such as
water hardness, pH, temperature and DO on nitrate toxicity to aquatic organisms
(Scott and Crunkilton 2000), and no studies found to date have tested these potential
interactions specifically. There is, however, anecdotal evidence in the literature for
influences from some of these variables.
Temperature does not appear to affect the toxicity of nitrate to freshwater fish. Colt
and Tchobanoglous (1976) concluded that median lethal concentrations observed for
channel catfish (Ictalurus punctatus) exposed to nitrate were independent of
temperatures at 22, 26 and 30 ºC. It should be noted, however, that this is a small
range of temperatures, and catfish are fairly robust.
Anecdotal evidence suggests that nitrate uptake may be pH-limited. While Jensen
(1996) reported that the freshwater crayfish Astacus astacus exhibited limited nitrate
uptake at pH ≈ 8.3 (e.g., nitrate concentrations in the haemolymph were below
ambient water values), the authors refer to McMahon and Stuart (1989) who found
extracellular NO3- concentrations higher than ambient water values in the crayfish
Procambarus clarki held in water acidified to pH 4 with nitric acid.
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Higher chloride concentrations tend to reduce nitrite toxicity to fishes, as the chloride
ion will bind competitively with chloride cells (the primary site of nitrite uptake),
thereby limiting the amount of nitrate entering the blood stream (Wedemeyer and
Yasutake 1978; Russo et al. 1981; Lewis and Morris 1986). These same chloride
interactions however, do not appear to reduce the toxicity of nitrate to salmonids. For
chinook salmon and rainbow trout exposed to nitrate in both freshwater and 15‰
salinity salt water, nitrate was more toxic (p < 0.05) in saltwater by a factor of up to
1.4 (Westin 1974; for comparisons, see Appendices A, B). No explanation however,
was provided for the increased toxicity in trials with greater salinity.
6.1 Influence of Various Nitrate Salts on Toxicity
The toxicity of nitrate ions to aquatic organisms is assessed using either NaNO3,
KNO3 or NH4NO3 salts. As there are differing responses between organisms in
response to the type of salt used (Dowden and Bennett 1965; Schuytema and
Nebeker 1999c), it was necessary to screen toxicity assays based on salt type.
Ammonium nitrate is often used in amphibian toxicity assays due to its potential to
collect in the runoff from fertilizer applications in agricultural regions and, therefore,
provides a potentially concentrated source of nitrate to sensitive developing
amphibian embryos and larvae (Hecnar 1995; Oldham et al. 1997; Schuytema and
Nebeker 1999a). Acute lethality values (96-h LC50s), however, for amphibian larvae
exposed to ammonium nitrate can be an order of magnitude lower than for larvae
exposed to sodium nitrate (Schuytema and Nebeker 1999a). As the ammonium ion
can cause adverse effects on larval survival or growth at lower concentrations than
required for adverse effects from nitrate ions, this suggests that the toxicity of
ammonium nitrate compounds are due to the influence of the ammonium ion rather
than nitrate (Schuytema and Nebeker 1999c). Therefore, toxicity studies using
ammonium nitrate as the test compound were excluded from the data set used for
the development of the CWQGs for nitrate. Canadian Water Quality Guidelines for
ammonia already exist (CCME 2000).
Sodium salts are generally used in the study of the physiological effects of anions,
due to their high degree of solubility and low toxicity from the cation relative to the
anion (Jones 1941). For freshwater benthic insect larvae (Hydropsyche occidentalis
and Cheumatopsyche pettiti), Camargo and Ward (1992) demonstrated that toxicity
from exposure to NaNO3 was due to NO3- rather than Na+ ions. No mortality was
observed in test organisms exposed to NaCl at 1000 mg NaCl·L-1 (= 393 mg Na+·L-1),
whereas the most sensitive LC50 with NaNO3 was 290 mg NO3-·L-1, which represents
a sodium concentration 3.7 times lower (= 108 mg Na+·L-1). Similarly, Baker and
Waights (1994) found no statistically significant effect on the growth or survival of tree
frog (Litoria caerulea) tadpoles exposed to NaCl at the same Na+ concentrations as
those required to produce an effect using NaNO3. Therefore, toxic effects from
exposure to NaNO3 are likely due to the nitrate ion, and studies using NaNO3 were
included in the dataset for the derivation of the nitrate WQGs.
Potassium nitrate (sometimes used in inorganic fertilizers) has also been used to
assess the toxicity of the nitrate ion to aquatic organisms. In freshwater studies
Science-based Solutions No. 1-6
48
exposing animals to nitrate of both potassium and sodium salts, the former are often
found to be more toxic than the latter (Table 6.1). The only exception was for the
freshwater hydra (Hydra attenuata), for which sodium nitrate was more toxic (Tesh et
al. 1990). As animals in the Tesh et al. (1990) study were kept in distilled water,
possible disruptions in normal osmoregulatory functions may have contributed to the
observed differences in toxicity.
Table 6.1.
Relative toxicity of sodium and potassium nitrate salts to freshwater
organisms.
[NO3-] (mg NO3-·L-1)
Duration
(h)
96
Endpoint
K+ Salt
Na+ Salt
LC50
1840
8753
Trama (1954)
24
LC50
3373
9338
Dowden and
Bennett (1965)
Daphnia magna
(water flea)
96
TLm
552
3069
Dowden and
Bennett (1965)
Polycelis nigra
(planaria)
48
survival
555
2696
Jones (1940)
Gasterosteus
aculeatus
(stickleback)
240
lethal
concentration limit
79
1348
Jones (1939)
Hydra attenuata
(hydra)
288
NOEL
150 - 250
< 50
Tesh et al.
(1990)
Organism
Lepomis macrochirus
(bluegill)
Reference
A review of the relative toxicity of K+ and Na+ ions from chloride salts to freshwater
organisms also indicates that potassium salts are between 1.6 and 8.7 times more
toxic than the corresponding sodium salt (Table 6.2). Using sulfate as the associated
anion, the potassium salt was 4.7 to 11.7 times more toxic than the sodium salt for
Ceriodaphnia dubia, D. magna and Pimephales promelas (Mount et al. 1997). Using
a stepwise logistic regression model, Mount et al. (1997) found that the K+ ion
contributed significantly to observed mortality in both invertebrate and vertebrate
organisms, while the Na+ did not. Although Mount et al. (1997) found that the toxicity
of K+ decreased with the addition of other cations to the test solution, it is not known
whether a threshold exists for physiological effects from the K+ ion.
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49
Table 6.2. Relative toxicity of sodium and potassium chloride salts to freshwater
invertebrates.
Organism
Daphnia magna
(water flea)
Ceriodaphnia
dubia
(water flea)
Pimephales
promelas
(fathead minnow)
Duration Endpoint
(h)
Salt concentration
KCl
NaCl
(mg·L-1) (mg·L-1)
[NaCl]/
[KCl]
Reference
24
EC50
2184
1127
1.9
Lilius et al. 1994)
24
EC50
1023
625
1.6
24
24
48
EC50
LC50
EC50
3606
6380
1023
548
740
271
6.6
8.6
3.8
48
LC50
4770
660
7.2
Khangarot and Ray
1989)
Calleja et al. 1994)
Mount et al. 1997)
Khangarot and Ray
1989)
Mount et al. 1997)
24
LC50
3380
630
5.4
Mount et al. 1997)
48
LC50
1960
630
3.1
Mount et al. 1997)
24
LC50
8280
950
8.7
Mount et al. 1997)
48
96
LC50
LC50
6510
6390
910
880
7.2
7.3
Mount et al. 1997)
Mount et al. 1997)
These various lines of evidence suggest that the concentrations at which toxic effects
are observed in freshwater organisms exposed to KNO3 are primarily a function of
the potassium ion. This is also supported by Demaël et al. (1980) who stated that the
metabolic and hormonal effects, indicative of osmoregulatory stress, that they
observed when the freshwater fish Tinca tinca (tench) was exposed to potassium
nitrate at 8.5 mg K+·L-1, were due to the K+, not the NO3-. Therefore toxicity data from
studies using KNO3 were not considered in the development of the freshwater nitrate
guideline.
The salinity of the world’s seawater largely ranges from 33 to 37‰, while most fresh
inland waters have salinities ranging from 0.1 to 0.5‰ (Stumm and Morgan 1981;
Wetzel 1983). The ionic salinity of water is largely determined by the concentrations
of four cations (Ca2+, Mg2+, Na+, K+) and four anions (HCO3-, CO32-, SO42-, Cl-)
(Wetzel 1983). Therefore, additions of sodium nitrate or potassium nitrate in toxicity
tests can affect the overall salinity of the test solution. Mean naturally occurring
seawater concentrations of Na+ and K+ are 10 770 and 399 mg·L-1 (or 10.8 and
0.40‰), respectively (Stumm and Morgan 1981). In contrast, fresh North American
river water contains mean concentrations of Na+ and K+ of 9 and 1.4 mg·L-1 (or 0.009
and 0.0014‰), respectively (Wetzel 1983). Mean ambient levels of potassium found
Science-based Solutions No. 1-6
50
in the Great Lakes and a variety of rivers from the Canadian maritimes are generally
less than 2 mg K+·L-1 (Dalziel et al. 1998; Williams et al. 1998a,b; Williams and Kuntz
1999).
At concentrations of nitrate salts that elicit toxic responses in aquatic organisms,
sodium ion levels tend not to greatly exceed ambient sodium concentrations in fresh
water. For example, at the LOEC for growth reduction in frog embryos of
129 mg NO3-·L-1 (Schuytema and Nebeker 1999b), the corresponding concentration
of Na+ was 48 mg Na+·L-1, which is less than 5 times greater than ambient Na+ levels
(Wetzel 1983). Alternatively, at the lowest LOEC for a primary study using potassium
nitrate (55 mg NO3-·L-1; Marco et al. 1999), the corresponding potassium
concentration of 35 mg K+·L-1 is approximately 25 times higher than ambient levels.
In contrast, potassium ion concentrations found in marine toxicity studies which elicit
a response are within natural ranges of K+ in seawater (Stumm and Morgan 1981).
For example, the nitrate LC50 of 496 mg NO3-·L-1 observed in polychaetous annelids
exposed to KNO3 (Reish 1970), corresponds to a K+ concentration of 312 mg·L-1,
which is within ambient potassium levels for seawater. The levels of potassium
administered to freshwater test organisms at nitrate concentrations required to elicit a
response may be an order of magnitude higher than those normally encountered,
while marine organisms respond to nitrate toxicity at potassium levels normally
encountered. Therefore, it appears that potassium concentrations encountered in
toxicity tests are unlikely to cause adverse effects in marine organisms, and any
adverse effects observed are likely due to the nitrate ion.
The only available study exposing both Na+ and K+ ions to a marine species found a
significant increase (~20%) in larval shrimp mortality for both cations at the lowest
treatment concentration of 1 mg NO3-·L-1 (Muir et al. 1991). As there is no available
data to suggest that K+ ions are more toxic in saline environments, and because ion
fluxes in marine fish are an order of magnitude higher than in freshwater fish (Heath
1995), KNO3 studies were included in CWQG development for marine environments.
Unless otherwise specified, discussions of toxic responses by organisms in this
report are a result of exposure to the sodium nitrate salt.
6.2 Modes of Action
6.2.1 Uptake Mechanisms
The mechanisms regulating nitrate uptake in aquatic vertebrates and invertebrates
are not fully understood; however, elevated levels of nitrate have been found in bodily
fluids and tissues of invertebrates (crayfish and shrimp), and fish (rainbow trout)
exposed to high ambient nitrate levels (Jensen 1996; Stormer et al. 1996; Cheng et
al. 2002). Nitrate uptake was minor in crayfish (Astacus astacus) and rainbow trout,
each exposed to 62.0 mg NO3-·L-1 (Jensen 1996 and Stormer et al. 1996,
respectively). Crayfish had significantly increased nitrate concentrations (p < 0.01) in
haemolymph relative to control animals when exposed to sodium nitrate for seven
Science-based Solutions No. 1-6
51
days; however, these levels were still far below exposure concentrations (Jensen
1996). Similarly, Stormer et al. (1996) found that the nitrate concentrations in rainbow
trout plasma increased significantly from less than 1.9 mg NO3-·L-1 in control fish to
12.4 mg NO3-·L-1 in exposed fish, and remained constant over an eight-day exposure
period. As per the crayfish studied by Jensen (1996), the amount accumulated
accounts for only a fraction of the ambient concentration, suggesting only a weak
uptake route. This limited NO3- uptake did not measurably influence the electrolyte
balance or haematology in the rainbow trout (Jensen 1996). In addition to increases
in haemolymph nitrate levels, Cheng et al. (2002) found significant relationships (p ≤
0.001) between increasing ambient nitrate levels (48 to 2237 mg NO3-·L-1) and tissue
concentrations in the tropical marine prawn Penaeus monodon. At lower exposure
levels (i.e., 48 and 226 mg NO3-·L-1), the majority of nitrate accumulation in
P. monodon occurred within the first 12 h, and tissue levels were still increasing after
24 h at higher nitrate levels (i.e., 1317 and 2237 mg NO3-·L-1) (Cheng et al. 2002). At
the lowest exposure level of 48 mg NO3-·L-1, nitrate levels in tissues (muscle,
hepatopancreas, foregut, midgut, heart, gill) were 20 to 80% ambient levels, while
concentrations in eyestalks were 1.2 times greater than ambient levels (Cheng et al.
2002).
Although nitrite is actively transported into tissues via branchial chloride cells, nitrate
ion uptake through this route is either severely limited, or absent (Stormer et al. 1996;
Jensen 1996; Cheng et al. 2002). As plasma Cl- concentrations in rainbow trout were
shown to decrease under nitrite exposure (due to competitive exclusion at chloride
cell uptake sites), an associated decrease of plasma Cl- would also be expected if
nitrate shared the same uptake mechanism (Stormer et al. 1996). A lack of change in
plasma Cl- concentrations under nitrate exposure therefore suggests that uptake is
not likely to occur via chloride cells (Stormer et al. 1996). Another possible route of
nitrate influx may be via the diffusion of nitric acid (HNO3). However, due to the
readily dissociable properties of the nitrate ion, the proportion of nitrate as nitric acid
is negligible, and the accumulation of nitrate in tissues is thought to be attributed to
some type of active uptake mechanism (Cheng et al. 2002).
Mechanisms for nitrate uptake in amphibians have not been investigated. Due to the
permeability of amphibian skin, however, it is likely that dissolved nitrate could readily
enter trans-dermally (Hecnar 2001). There is also the potential for nitrate uptake
through the diet if tadpoles are feeding on algae or macrophytes that have
accumulated nitrate (Hecnar 2001).
There is little information on nitrate excretion rates in aquatic animals. In mammals
however, kidneys have been shown to accumulate ~60% of 15N-labelled nitrate
doses (Packer 1995), and as such the majority of nitrate in animals is lost via urine
within 24 hours (WHO 1986). Nitrate concentrations in crayfish haemolymph
remained high over the 7-d exposure period despite a very large osmotic gradient
relative to the surrounding water, suggesting a slow rate of depuration, most likely
through urine (Jensen 1996). In rainbow trout, nitrate is most likely excreted through
bile and urine (Doblander and Lackner 1997). Stormer et al. (1996) suggest that
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52
urinary loss plays a larger role in trout than in crayfish, with nitrate levels reaching a
quasi-steady balance between passive branchial influx and removal.
6.2.2 Direct Toxicity
6.2.2.1 Methaemoglobin formation
In animals, uptake of nitrate can ultimately inhibit the ability of haemoglobin, a
pigment in the blood, to carry oxygen to the various tissues of the body (WHO 1986).
This inhibition occurs through several steps. First, nitrate is reduced to nitrite within
the alimentary canal and guts of animals via bacteria such as Nitrobacter which use
NADH as an electron donor for the oxidative phosphorylation of ADP to ATP:
NO3- + NADH2 + 2 ADP + 2 Pi → NO2- + NAD+ + 2 ATP + H2O
(deSaint-Blanquat 1980)
Nitrite that is produced from this reaction is then free to be taken up into the blood
stream where it will react with the haem iron (as Fe2+) in oxyhaemoglobin (HbO2),
oxidizing it to Fe3+, and thereby creating methaemoglobin (Hb+):
4HbO2 + 4NO2- + 4H+ → 4Hb+ + 4NO3- + O2 + 2H2O
(Stormer et al. 1996)
As methaemoglobin binds irreversibly with oxygen molecules, transfer of oxygen from
the blood to cells in the body is inhibited, and appreciable levels of Hb+ can result in
hypoxia.
Background levels of methaemoglobin in fish blood, and the response in
methaemoglobin levels when fish are exposed to nitrate, can vary, and may be
related to exposure conditions, or the duration of exposure. Salmon blood normally
contains between 3.3 to 17.5% methaemoglobin in the absence of nitrite (Lewis and
Morris 1986; Brauner et al. 1993). Grabda et al. (1974) found that exposure to
potassium nitrate at 31 mg NO3-·L-1 for up to eleven weeks increased
methaemoglobin levels in rainbow trout to approximately 28%, relative to 1% in
controls. In contrast, methaemoglobin levels in rainbow trout exposed to
62 mg NO3-·L-1 (as sodium nitrate) for eight days, remained below 3% of total
haemoglobin (Stormer et al. 1996).
At blood levels of 20-25% methaemoglobin, hepatic tissue respiration rates decrease,
potentially leading to serious liver damage (Grabda et al. 1974). Methaemoglobin
levels above 50% inhibit “the cough response”, thereby preventing salmon from
purging sediment collected in the buccal cavity. At levels above 70% the fish
becomes torpid which could lead to anoxic death if the fish suddenly has increased
oxygen demands (Lewis and Morris 1986). Other effects observed due to increased
methaemoglobin include serious damage to the peripheral blood, and hematopoietic
(blood production) centres of the kidney (Grabda et al. 1974). Low haemoglobin
levels in fish could reduce survival, as Jones (1971) has demonstrated that induced
Science-based Solutions No. 1-6
53
hemolytic anaemia (abnormally low haemoglobin levels) resulted in a 34 to 40%
reduction in maximum sustained swimming speeds for rainbow trout.
Long-term sub-lethal toxicity from elevated methaemoglobin levels are unlikely as
fish possess defense mechanisms, such as the NADH-reductase system, which will
reduce methaemoglobin back to haemoglobin (Kamstra et al. 1996). Methaemoglobin
levels in rainbow trout exposed to 0.32 mg NO2-·L-1 increased from approximately 3%
in control fish to 27% after 14 days, however, then declined to near control-levels
after 48 days (Doblander and Lackner 1997). Huey and Beitinger (1982)
demonstrated that the NADH-methaemoglobin reductase enzyme in catfish
(I. punctatus) provides a rapid detoxification mechanism, with a 5-fold decrease in
catfish methaemoglobin levels occurring within 24-h of placing the animals in a nitritefree medium. Doblander and Lackner (1997) also determined that nitrite present in
blood plasma can be taken up by erythrocytes and oxidized to nitrate under oxic
conditions, thereby preventing the nitrite from oxidizing the haemoglobin to
methaemoglobin. It is estimated that erythrocytes, and other cells such as
hepatocytes, could detoxify almost 20% of nitrite taken up (Doblander and Lackner
1997). Enhanced activation of these defence mechanisms however, have an
associated metabolic cost for the fish that may redirect energies obtained from food
sources and, therefore, limit growth rates (Kamstra et al. 1996).
It is not known whether fish possess the same capability as mammals for
endogenous nitrate reduction, in which the bacterial flora within the animal reduce
nitrate to nitrite, or whether nitrate must first be converted to nitrite in the surrounding
water prior to uptake. In a review of the Grabda et al. (1974) study, Colt and
Armstrong (1981) suggest that because nitrite levels were not monitored in the water,
it was possible that bacteria in the water surrounding the fish were reducing nitrate to
nitrite (Colt and Armstrong 1981). Supporting evidence by Anuradha and Subburam
(1995) showed that for carp (Cyprinus carpio) exposed to 36 mg NO3-·L-1 (as
NaNO3), methaemoglobin levels were significantly higher (43.7%, p = 0.01) when
held in water containing nitrate reducing sewage bacteria, than in water without
bacteria (10.0%), or control water without nitrate (6.5%). Nitrate reducing bacteria
present in sewage, such as Pseudomonas (Anuradha and Subburam 1995), are
numerous in all natural surface waters (McCoy 1972).
Another potential link between nitrate and methaemoglobin formation has been
shown in the physiological response of freshwater mosquito fish (Gambusia affinis)
exposed to sodium nitrate (Nagaraju and Ramana Rao 1983, 1985). Nagaraju and
Ramana Rao (1983) found that exposure to 29 mg NO3-·L-1 resulted in an increase of
succinic dehydrogenase activity, and a decrease in lactate dehydrogenase activity.
These changes indicate that the fish were likely using an enhanced glycolysis
process to produce the H+ required to reduce methaemoglobin (formed due to nitrate
exposure) back to haemoglobin. At this level of nitrate exposure, fish were also found
to have significantly elevated enzyme levels which would aid in the conversion of
methaemoglobin back to haemoglobin (Nagaraju and Ramana Rao 1985). These
results suggest a biochemical response by the fish to counteract stresses induced by
nitrate toxicity (Nagaraju and Ramana Rao 1985).
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Methaemoglobinemia is also a likely mode of toxicity in amphibians (Huey and
Beitinger 1980a,b). In studies with bullfrog larvae (Rana catesbiana) and channel
catfish (Ictalurus punctatus), Huey and Beitinger (1980b) observed increased blood
levels of methaemoglobin in both species when exposed to nitrite; however, they
noted that the tadpoles were more resistant than the fish to nitrite-induced
methaemoglobin formation. The authors speculated that there may be less nitrite
uptake in tadpoles, and/or tadpoles may have a more efficient methaemoglobin
reductase system than fish.
The mechanism of nitrate toxicity in invertebrates has yet to be determined, but
evidence suggests that, similar to vertebrates, nitrate may affect the oxygen carrying
pigments (Muir et al. 1991). For example, histological examination of penaeid larvae
has shown that exposure to 10 mg NO3-·L-1 elicited vacuolative change and tissue
damage to the midgut and hypodermis that are thought to be the sites of
haemocyanin synthesis and uptake/removal, respectively, in decapods (as per
Senkbeil and Wriston 1981a,b). Such sub-lethal histopathological changes may affect
the survival of larval forms in the environment (Muir et al. 1991).
6.2.2.2 Osmoregulation Disruption
Although the physiological mechanisms are not fully known, it appears that the lethal
toxicity of nitrate may be related, in part, to the inability of the animal to maintain
adequate osmoregulation under waters with high salt contents (Brownell 1980; Colt
and Armstrong 1981). Acute mortality estimates for freshwater fish exposed to
NaNO3 range from 1300 to 9300 mg NO3-·L-1 (Appendix A). At these concentrations,
it may be difficult to determine whether the toxic response is due to the cation or
anion, as lethal NaNO3 levels at this magnitude are comparable to lethal NaCl levels
(Colt and Armstrong 1981). For example, 24-h LC50s for bluegills exposed to NaNO3
(3200 and 3500 mg Na+·L-1), are similar to those for NaCl (5100 and
5600 mg Na+·L-1) (Trama 1954; Dowden and Bennett 1965). Similarly, Brownell
(1980) found that acutely toxic levels of NaNO3 for marine fish (24-h LC50s
> 15 283 mg NO3-·L-1) raised the salinity of the test waters from 35‰ to 59 - 83‰.
When seawater salinity was increased to 50 and 70‰ using NaCl, 15% and 100% of
test fish (n = 20 each) died, respectively (Brownell 1980).
Sodium ions are normally passively taken up through the guts of marine fish, and
actively pumped out of the body via chloride cells in the gills, while freshwater fish
actively take up Na+ across the gill surface via chloride cells in exchange for other
monovalent waste products in the blood (e.g., ammonium, hydrogen ions) (Heath
1995). Fish tend to maintain plasma Na+ concentrations of approximately 150 to
160 mM in fresh- and marine waters, while ambient concentrations range from
approximately 0.3 mM in fresh waters to 520 mM in marine waters (Bone and
Marshall 1986). Marine fish generally have a greater number of chloride cells than
freshwater fish to help accommodate these greater ionic fluxes (Heath 1995). Fish
subjected to a higher osmotic gradient from the surrounding water than normal may
undergo cellular stress from loss of water.
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At high concentrations nitrate is also able to remove proteins from cell membranes
(Manzano et al. 1976). No information was available on osmoregulatory disruption in
amphibians or invertebrates due to nitrate exposure.
6.2.3 Indirect Toxicity
6.2.3.1 Role of Nitrate in Nutrient Enrichment
Nitrate serves as the primary source of nitrogen for aquatic plants in well oxygenated
systems, and excessive concentrations have been shown to result in algal blooms
and eutrophication in ponds (Nordin and Pommen 1986; Meade and Watts 1995).
While it is generally accepted that phosphorus is the nutrient that most limits primary
production in freshwater systems, and nitrogen is limiting in marine systems (Paerl
1993; Crouzet et al. 1999; US EPA 2000b), the role of nitrogen in eutrophication may
vary considerably in both types of systems. The dependence of the relative
contributions of both nutrients (i.e., N:P ratios) are examined in a separate CCME
discussion paper (CCME 2002).
Adverse ecological effects associated with eutrophication include a loss of water
clarity, changes in plankton and fish species composition, physical obstructions in
waterways which can impede fish migration or rearing, and potentially fatal oxygen
depletion (Environment Australia 2000b). Increased phytoplankton, and/or aquatic
plant biomass can lead to increased biological oxygen demands (BOD) on a system
for two main reasons: a) plants and algae consume oxygen when not undergoing
photosynthesis, which results in greater diurnal respiration rates, and b) after
senescense, or death, greater populations of bacteria are required to break down the
additional organic matter from excess plants/algae, which requires greater oxygen
consumption. Therefore, the risks of low oxygen (hypoxia), or complete lack of
oxygen (anoxia) events can increase, and fish kills may result if critical oxygen levels
are not maintained.
Over-stimulation of phytoplankton production in the pelagic zone can reduce the
amount of light penetrating the water column, and as a result, primary production of
benthic algae (periphyton) can be adversely affected. Nutrient enrichment studies on
small (≤ 3.4 ha), relatively shallow (mean depth ≤ 5.7 m) lakes in Michigan
demonstrated that increased phytoplankton production accompanied reductions in
periphyton production (Vadeboncoeur et al. 2001). In nutrient-enriched coastal
waters where light penetration is adequate, over-stimulation of epiphytic algae has
been linked to the widespread loss of seagrass communities, as epiphytes can also
limit the photosynthetic capabilites of the underlying macrophytes (Coleman and
Burkholder 1994). In mesocosm experiments, nitrate supply levels were found to
have a controlling influence on the community structure and species dominance of
epiphytes on the eelgrass (Zostera marina L.). Additions of 0.2 and 0.4 mg NO3-·L-1
stimulated total epiphyte productivity (primarily as blue-green algae and diatoms)
over a period of 6 weeks (170 ± 47 and 157 ± 10 mg C·m-2·d-1, respectively, versus
102 ± 9 mg C·m-2·d-1 in controls; p < 0.05) (Coleman and Burkholder 1994).
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Nutrient enrichment can lead to the proliferation of algae and photosynthetic bacteria
that produce toxic metabolites. Ingestion of these algal toxins can impair the health of
aquatic organisms and they may accumulate in shellfish to levels that are toxic to
consumers, including humans (Smith et al. 1999). Of the algae that produce toxins,
cyanobacteria, or blue-green algae (Cyanophyta) are of primary importance in fresh
waters, and diatoms and dinoflagellates are important sources in marine waters
(Chambers et al. 2001). Cyanobacteria are unique in that all species will assimilate
fixed inorganic nitrogen (i.e., nitrate, nitrite and ammonia), but some species are also
capable of directly fixing atmospheric nitrogen (N2) into organic nitrogen
(Environment Australia 2000b). This provides a competitive advantage over other
primary producers in low nitrogen environments, and as such, cyanobacteria tend to
dominate the algal species assemblage when N:P ratios (by weight) fall below 29:1
(Smith 1983). Cyanobacteria which are known to produce toxins in Canadian inland
surface waters include Anabaena, Aphanizomenon, Microcystis and Phormidium
(Chambers et al. 2001). Although passive ingestion of cyanobacterial toxins have not
been known to be fatal to humans, severe skin irritations can occur, and their
neurotoxic and hepatotoxic properties have been responsible for liver damage and
death of livestock (Environment Australia 2000b; Health Canada 1998).
Diatoms (Bacillariophyceae) are a very large, diverse group of primarily sessile
marine and freshwater phytoplankton that occur in both unicellular and colonial forms
(Wetzel 1983). The diatom Nitzschia pungens produces domoic acid, a toxin that can
cause amnesiac shellfish poisoning in humans consuming mussels from
contaminated waters (Chambers et al. 2001). In 1987, 108 cases of acute poisoning
(including three deaths), were reported in Prince Edward Island after people ingested
blue mussels (Mytilus edulis L.) contaminated with domoic acid (Bates et al. 1989).
The cause of the bloom of N. pungens responsible for the elevated toxin levels was
thought to be related to inorganic nitrogen enrichment. Nitzschia pungens population
levels, and domoic acid production have been shown to respond positively to both
nitrate and ammonium in in situ experiments (Bates et al. 1993), and blooms of
N. pungens in eastern Prince Edward Island occurred only when ambient nitrate
levels exceeded 1.1 µg NO3-·L-1 (Smith et al. 1990). As a result, the massive bloom of
N. pungens which led to the accumulation of domoic acid in in 1987, was attributed to
a long dry summer followed by heavy nitrate runoff during an intensely wet autumn
(Chambers et al. 2001).
Dinoflagellates (Dinophyceae) are unicellular flagellated algae and most have a
conspicuous armoured cell wall with large spines (Wetzel 1983). Large colonies of
dinoflagellates can produce ‘red tides’ in coastal marine waters, leading to
widespread fouling of waterways and the production of shellfish toxins (Chambers et
al. 2001). Isolated outbreaks of shellfish toxicity from dinoflagellate blooms such as
Gonyualax acatenalla have been documented along the coast of British Columbia,
however, causal links to nutrient additions were difficult to demonstrate (Chambers et
al. 2001). In a review of factors influencing global red tide occurrences, Hodgkiss and
Ho (1997) reported that decreasing N:P ratios in Tolo Harbour, Hong Kong were
associated with an increase in red tide events, and that occurrences were highly
probable when dissolved nitrogen and phosphorus levels exceeded 0.1 mg N·L-1 and
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0.02 mg P·L-1, respectively. An increase in dinoflagellate abundance, however does
not always result in increased toxic effects. Isolated population increases of the toxinproducing dinoflagellate Alexandrium catenella in Hong Kong were not followed by
paralytic shellfish poison contamination of the resident shellfish (Siu et al. 1997).
The relationship between increasing nitrogen concentrations in both marine and fresh
waters and eutrophication are not clearly defined. For example, there is a wide range
in nitrate concentrations that produce optimal growth of the marine dinoflagellate
Alexandrium catenella, from 14 to 548 mg NO3-·L-1 (Siu et al. 1997), which would
make predicting a population response based on nitrate exposure levels alone
extremely difficult. Total nitrogen levels are also poor predictors of algal biomass
(measured as Chl a) in lakes and coastal regions; algal biomass can be predicted
better from either total phosphorus, or a combination of the two nutrients (Mazumder
and Havens 1998; Meeuwig et al. 2000). It should also be noted that other factors
can affect plant and algal growth, so in some cases a relationship between nutrient
levels and primary productivity may not exist. For example, where there is light
limitation due to very high turbidity, added nutrients might not necessarily stimulate
growth. In a study of the potential for eutrophication in coastal inlets in Nova Scotia,
Strain and Yeats (1999), found that eutrophic inlets were associated with poor
flushing characteristics, and tended to have more than 50% of the water trapped
behind the inlet sill, while non-eutrophic inlets were at, or near 0% entrainment. To
better predict how altering water column nitrogen and phosphorus levels will
influence eutrophication processes, further research is required in understanding
factors regulating internal nutrient cycling, and in the complex interactions between
nutrients and food webs (Smith et al. 1999).
In many aquatic ecosystems, eutrophication-related effects will occur at nitrate
concentrations that are lower than those required to cause direct toxicity. Total
nitrogen levels associated with highly eutrophied lakes, rivers, and coastal waters
around the world are often below 1 mg N·L-1 (Table 6.3). If all nitrogen were in the
form of nitrate, this would correspond to a level of 4.4 mg NO3-·L-1; well below the
levels at which the majority of direct toxic effects have been documented
(Appendix A). As part of the whole-lake fertilization program of the Experimental
Lakes Area, northwestern Ontario, one-half of Lake 226 was fertilized with carbon
and nitrogen (as nitrate) over an eight year period (Findlay and Kasian 1987). The
increase in ambient total nitrogen concentrations in the nitrate-fertilized portion of
Lake 226 (= 0.46 ± 0.09 mg TN·L-1, compared to 0.31 ± 0.04 mg TN·L-1 in an
unfertilized control lake), resulted in overall phytoplankton biomass increasing by a
factor of 2 to 4 over unfertilized years (Findlay and Kasian 1987). Mean
phytoplankton biomass levels (3070 ± 1210 mg·m-3) were also substantially higher
than those found in the control lake not undergoing nitrate fertilization
(720 ± 200 mg·m-3) (Findlay and Kasian 1987). Similarly, in enclosure experiments in
a eutrophic Hungarian reservoir, phytoplankton production responded quickly to
nitrate-nitrogen additions. Within one week of nitrate additions (bringing the
mesocosm nitrate level to 13 mg NO3-·L-1), total phytoplankton biomass increased
from 24 to 59 mg·L-1 (Présing et al. 1997). By the end of the week, all nitrate-nitrogen
supplied to the mesocosm had been used in algal production (primarily by diatoms
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and cryptomonads), and levels had returned to those seen in controls (Présing et al.
1997).
Table 6.3. Average total nitrogen levels in global lakes, streams and coastal marine
waters of varying trophic status.
Trophic State
Lakesa
TN (mg N·L-1)
Streamsb
Marinec
Oligotrophic
< 0.35
< 0.7
< 0.26
Mesotrophic
0.35-0.65
0.7-1.5
0.26-0.35
Eutrophic
0.65-1.2
> 1.5
0.35-0.40
Hypereutrophic
> 1.2
> 0.40
a
Nürnberg 1996 [North American, European and Asian lakes];
Dodds et al. 1998 [North American and New Zealand Streams];
c
Håkanson 1994 [source waters not known]; fromSmith et al. 1999.
b
Increasing nitrate levels in surface waters may also lead to changes in algal species
compositions. The Grand River in southern Ontario is situated in a lowland area
dominated by heavy urban and agricultural development, and is subject to
increasingly high nitrate loads (e.g., up to 18 mg NO3-·L-1 in 1994) (Rott et al. 1998).
Multivariate analyses of benthic diatom species assemblages along the river showed
that Surirella brébissonii and Navicular lanceolata were associated with higher nitrate
values, while N. gregaria and N. tripunctata were associated with moderate nitrate
levels (Rott et al. 1998). From factorial enrichment experiments exposing natural
Lake Huron phytoplankton assemblages to nitrate (0.27 to 4.3 mg NO3-·L-1) and
phosphorus (4 to 16 µg P·L-1), Pappas and Stoermer (1995) determined that
populations of cyanophytes, flagellates, and the diatom Cyclotella commensis,
responded positively to increasing nitrate additions, while other species were either
not affected by, or as in the case of Cyclotella pseudostelligera, were inhibited by
higher nitrate levels (Pappas and Stoermer 1995). The authors suggest that
increasing nitrate levels in the Great Lakes would therefore affect algal species
composition in these waters (Pappas and Stoermer 1995).
In coastal regions, phytoplankton has been shown to readily respond to nitrate
enrichment. In nutrient limitation studies using mesocosms in coastal lagoons in
Narragansett Bay, Rhode Island, additions of high levels of nitrate (514 mg NO3-·L-1)
resulted in substantial phytoplankton blooms, with Chl a levels 12 times greater than
in controls, and 3 times greater than in mesocosms enriched with phosphorus alone
(22 mg P·L-1) (Fisher’s LSD test, p < 0.010) (Taylor et al. 1995). Enrichment
experiments performed on waters collected from a variety of salinity levels (0 - 30‰)
in Waquoit Bay, Massachusetts, showed that addition of 6.2 mg NO3-·L-1 in highly
saline waters (23 - 30‰) increased Chl a levels from ~5 µg·L-1 in controls to
~18 µg·L-1 (Tomasky et al. 1999). However, in fresh (0‰), and brackish waters (10 19‰), phytoplankton growth responded to phosphorus additions only (Tomasky et al.
1999).
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6.2.3.2 Role of Nitrate in Acidification
Acid neutralising capacity (ANC) is a measure of surface water’s capacity to
consume H+ and therefore buffer against acidification (Laudon et al. 2000).
Increased inputs of HNO3 to surface waters from precipitation could potentially
decrease the neutralising capacity of the water body through H+ inputs. Driscoll and
Van Dreason (1993) linked an increasing trend in nitrate levels of 0.1 mg NO3-·L-1·a-1
between 1982 and 1990 in Constable Pond (Adirondack mountains, New York) with a
simultaneous decrease in the ANC of the system. Decreases in the ANC of the pond
corresponded with spring snowmelt when high concentrations of nitrate were
released from the snowpack (Heathwaite et al. 1996). The nitrate itself would not
have contributed any acidity to the system, as it is the conjugate base to a strong
acid, and therefore is a neutral ion. This suggests that HNO3 precipitated in snow
contributed the H+ which was responsible for lowering the ANC. In contrast, a review
of studies on the Muskoka-Haliburton lakes in Ontario between 1976 and 1980
showed no relationship between H+ concentrations and NO3- (Elder 1984). This
suggests that either the nitrate in these lakes was primarily due to sources other than
atmospheric HNO3 deposition, or the H+ primarily originated from some other source
(such as atmospheric deposition of H2SO4). Similarly, a study quantifying the sources
to pH reductions in spring melt waters of 12 Swedish streams found no correlation
between nitrate levels and pH decline; in this case, organic acids were the primary
contributors to the acidity of the streams (Laudon et al. 2000).
6.3 Toxicity to Freshwater Life
6.3.1 Algae and Plants
Nitrate is a required element for plant growth, and due to its greater abundance in
surface waters relative to other fixed nitrogen species (e.g., ammonium), it is the
most widely used form of nitrogen by vascular plants and algae (Pinar et al. 1997;
Crouzet et al. 1999). As nitrate is actively taken up by aquatic primary producers, its
uptake is generally not limited by low environmental concentrations (Cresswell and
Syrett 1981; Pinar et al. 1997).
Results from the tissue analysis of half a dozen macrophyte species suggest that a
minimum of 1.3% nitrogen per dry weight of plant tissue is necessary for macrophyte
growth (Gerloff and Krombholz 1966, as cited in Forsberg 1975). No effect on the
yield occurred when tissue nitrogen content was above this critical concentration.
The critical nitrogen concentration for the blue-green algae Microcystis aeruginosa
was determined to be 4% (Gerloff and Skoog 1954, as cited in Forsberg 1975).
No studies were located that directly tested nitrate toxicity to aquatic primary
producers. Incubation studies using the alga Scenedesmus subspicatus showed that
all levels of sodium nitrate that were added to the test medium (from 4 to
285 mg NO3-·L-1) increased algal growth, with maximum growth occurring at
55 mg NO3-·L-1 (Hund 1997).
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Although not directly toxic to the plants, nitrate taken up by aquatic plants could prove
to be an environmental hazard to herbivorous consumers. From agricultural studies,
it is known that an excess of nitrate in fodder can be toxic to livestock. A nitratenitrogen content of around 0.2% dry wt. is generally accepted as the upper limit for
forage crops used for livestock feeds; however, toxic effects may occur at nitrate
concentrations as low as 0.07% if that crop is the sole food source (Tucker and
Debusk 1983). Aquatic plants can sequester nitrate to levels above the safe level for
livestock. For example, Tucker and Debusk (1983) examined NO3--N uptake in water
hyacinth (Eichhornia crassipes) cultured for one year in a flow-through system with
an ambient concentration of 1.4 mg N·L-1. Plant tissue nitrate-nitrogen content ranged
from 0.05 to 0.21% dw (= 3.2% total nitrogen dw, assuming NO3--N accounts for
6.6% of the total nitrogen), with the greatest concentrations accumulating in the plant
during the slow growing fall and winter months. For ten of the twelve study months
(April and May excluded) E. crassipes grown in water with 1.4 mg NO3--N·L-1 had
NO3--N contents ≥ 0.07% dry wt. Unfortunately, no information is available on the
effects of elevated nitrate levels in aquatic plants to aquatic and terrestrial consumers
of those plants. Nonetheless, the possibility exists that secondary poisoning through
elevated plant nitrate levels could occur even though ambient water levels of nitrate
are not directly toxic to aquatic life.
6.3.2 Invertebrates
Freshwater invertebrates are relatively sensitive to nitrate exposure, with primary
studies showing LOEC values being comparable to those of amphibians (see
Appendix A). Toxic responses include mortality, reduction in fecundity and
immobilisation.
Available studies on benthic invertebrates include caddisflies, hydra, and planaria.
Short-term (up to 120-h) static bioassays for caddisflies in soft water (hardness
= 42.7 mg CaCO3·L-1) were used to determine the toxicity of nitrate (as NaNO3) to
two species of common North American benthic insect larvae (Camargo and Ward
1992). Acute LC50 values decreased with increasing exposure time (72- to 120-h) and
from last to early instar stage. For early instars of Hydropsyche occidentalis, and
Cheumatopsyche pettiti, the 120-h LC50s were 290 and 472 mg NO3-·L-1,
respectively, suggesting a differential response to toxicity between species (Camargo
and Ward 1992). The caddisflies were also exposed to high NaCl levels (up to
1100 mg Na+·L-1). As no mortality was observed, it is likely that the toxic effects seen
in the study were fundamentally due to the nitrate ion (Camargo and Ward 1992).
Using mortality data from the above study, Camargo and Ward (1995) determined
safe concentrations (SCs = 8760-h LC0.01s) for the two caddisfly species. These
values are analogous to NOECs and are intended to be protective of animals
throughout their entire larval stage (approximately 1 year or 8760 h). Calculated SCs
for early instars of H. occidentalis and C. pettiti are 6.2 and 10.6 mg NO3-·L-1,
respectively (Camargo and Ward 1995). These values are lower than estimated safe
concentrations for salmonid fish at 25 to 35 mg NO3-·L-1 (see Westin 1974, Section
6.4.3).
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Jones (1940, 1941) determined the toxicity of a variety of anions to the freshwater
planaria, Polycelis nigra, using distilled water in the test media. When exposed to
NaNO3 at pH 6.4, the planaria in both studies responded in a very similar fashion; the
concentrations corresponding to a median survival time of 48 hours were
2666 mg NO3-·L-1 (Jones 1941) and 2697 mg NO3-·L-1 (Jones 1940).
Sodium and potassium salts were used to determine the toxicity of nitrate to the
growth of hydra (Hydra attenuata) populations (Tesh et al. 1990) and the survival of
Lymnea snails (Dowden and Bennett 1965). The no-effect level of the nitrate ion on
hydra population growth when exposed to KNO3 was > 150 mg NO3-·L-1, based on
NOECs of between 150 and 250 mg NO3-·L-1; with NaNO3, the NOEC for population
growth was less than 50 mg NO3-·L-1 (Tesh et al. 1990). Snails also exhibited a
differential response to sodium and potassium salts, with median lethal tolerance
limits (TLms; 50% hatching success of eggs) of 2373 and 671 mg NO3-·L-1,
respectively (Dowden and Bennett 1965). Dowden and Bennett (1965) speculate that
the firm gelatin-like covering of the egg masses for these snails may afford extra
protection to the developing embryos.
Effects data for cladocerans exposed to nitrate ranged from 189 to 6205 mg NO3-·L-1
(Appendix A). Scott and Crunkilton (2000) exposed two common North American
cladocerans, Ceriodaphnia dubia and Daphnia magna to sodium nitrate in daily
renewal tests at concentrations up to 501 mg NO3-·L-1 in moderately hard waters
(150 mg CaCO3·L-1, pH 7.5). Ceriodaphnia dubia was more susceptible to nitrate
toxicity than D. magna, with 7-d LOECs for neonate production of 189 and
3176 mg NO3-·L-1, respectively. The LOEC for C. dubia is within the range reported
for surface waters draining agricultural lands (44 to 266 mg NO3-·L-1; McCoy 1972).
Scott and Crunkilton (2000) also determined the short-term toxicity (48-h LC50) of
C. dubia and D. magna neonates to be 1657 and 2047 mg NO3-·L-1, respectively.
Acute toxicity (96-hr LC50) values for D. magna exposed to KNO3 and NaNO3 in
standard reference water were 549 and 3070 mg NO3-·L-1, respectively (Dowden and
Bennett 1965). In studies where Daphnia magna were exposed to sodium nitrate in
centrifuged Lake Erie water, concentrations required to produce a threshold limit that
would just fail to immobilise D. magna (analogous to a NOEC) after 16 and 48-h
exposures were 6205 and 3650 mg NO3-·L-1 (Anderson 1944, 1946, respectively).
The giant freshwater prawn (Macrobrachium rosenbergii), a native of the Indo-Pacific
region is grown extensively in aquaculture operations (Eldredge 2001). The tolerance
of M. rosenbergii to high sodium nitrate levels (up to 4483 mg NO3-·L-1) in
recirculating aquaculture tanks was determined under freshwater - brackish
conditions (salinity ranging from 0.5 - 4‰) (Wickins 1976). For a three week
exposure period using growth as the endpoint, the EC50 was 775 mg NO3-·L-1 and the
LC50 was 709 mg NO3-·L-1 (Wickins 1976). The author noted that the EC50 value may
have been slightly elevated due to the extremely slow growth of the prawns; perhaps
with a longer exposure period, growth effects would have been seen at a lower
concentration.
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6.3.3 Fish
Eggs were found to be the most sensitive stage of freshwater fish to nitrate exposure.
Eggs and fry of two salmon and three trout species were exposed to NaNO3
concentrations ranging from 3 to 30 mg NO3-·L-1 in flow-through systems with low
water hardness (25 to 39 mg CaCO3·L-1) for a period lasting from egg fertilization to
30 days past yolk absorption (first feeding stage) (Kincheloe et al. 1979). Significant
increases (p ≤ 0.05) in total mortality for anadromous steelhead and freshwater
rainbow trout (both Oncorhynchus mykiss) were found at nitrate concentrations of 5
and 10 mg NO3-·L-1, respectively. Significant mortality was also found for chinook
salmon fry at 20 mg NO3-·L-1 and Lahontan cutthroat trout (Salmo clarki) eggs and fry
at 20 and 30 mg NO3-·L-1, respectively (Kincheloe et al. 1979). The authors also
found morphological abnormalities in some surviving fry, however details were not
provided. Although this study clearly demonstrated sensitivity of eggs and early
salmonid life stages to nitrate, additional egg mortalities caused by Saprolegnia
fungal infestations could not be segregated from the data by the authors. In a study
looking at the effects of eutrophication on carp reproduction, Bieniarz et al. (1996)
exposed fertilized eggs to sodium nitrate concentrations of 15, 150 and 500 mg NO3·L-1. The percentage of eggs hatching was significantly lower (p < 0.01) than that in
the control at all experimental concentrations, suggesting that levels of nitrate
normally found in the environment may lower the reproductive effort in carp (Bieniarz
et al. 1996). It should be noted, however, that even within the control group there was
a very low hatch rate of approximately 48%.
Fingerling and juvenile stages of fish are significantly more resistant to nitrate
exposure than egg stages. Fingerlings of chinook salmon and rainbow trout were
exposed in fresh water to NaNO3 for 10 days to a maximum concentration of
6500 mg NO3-·L-1, with renewal of the test solutions after 4 days (Westin 1974).
Median lethal tolerance limits (7-d TLm) for these older salmonids are 4800 and
4700 mg NO3-·L-1, respectively (Westin 1974). Behavioural responses to nitrate
exposure for the fish in this study included an inability to swim upright, laboured
respiration, reduced movement with erratic swimming, yawning, and accelerated
opercular movements. For all exposure concentrations, no abnormalities were found
in tissues examined histopathologically (Westin 1974). [Note: Westin also conducted
toxicity tests with these two fish species in saline water. Those results are discussed
in section 6.4.3.]
Channel catfish (Ictalurus punctatus) juveniles are similarly tolerant to nitrate. In an
observational study on increasing catfish populations in a closed, recirculating
system, Knepp and Arkin (1973) found that ambient nitrate concentrations allowed to
reach 400 mg NO3-·L-1 over 170 days did not have an impact on individual growth or
behaviour (e.g., lethargy). The 96-h LC50 for fingerling channel catfish (50 to 76 mm
total length) exposed to sodium nitrate using static bioassays at 30°C was
6200 mg NO3-·L-1 (Colt and Tchobanoglous 1976). Although survival times of catfish
exposed to nitrate generally decreased with increasing temperatures, the incipient
LC50 values were independent of experimental temperatures (22°, 26° and 30°) (Colt
and Tchobanoglous 1976). In a ten-week study of the humoral immune response of
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channel catfish exposed to low (558 mg NO3-·L-1) and high (1280 mg NO3-·L-1) nitrate
levels, Collins et al. (1976) did not find a consistent effect on antibody levels of the
fish, suggesting that those levels of nitrate stress did not significantly increase
immunosuppression in I. punctatus.
Concentrations of nitrate which affect larval and juvenile stages of common bluegill
(Lepomis macrochirus) and fathead minnows (Pimephales promelas) are comparable
to those that are toxic to channel catfish. Trama (1954) determined the acute toxicity
(96-h LC50) of sodium nitrate to juveniles (5 to 9 cm total length) of the common
bluegill in relatively soft water (up to 50 mg CaCO3·L-1; pH 7.4 to 8.8) to be
8753 mg NO3-·L-1. Similarly, the 96-h LC50 for larval fathead minnows exposed to
sodium nitrate was 5941 mg NO3-·L-1 (Scott and Crunkilton 2000). In contrast to
Kincheloe et al. (1979), Scott and Crunkilton (2000) found significant failures of
hatching for fertilized P. promelas eggs only at 6353 mg NO3-·L-1. The difference in
susceptibility of the fertilized eggs could be species-specific, as P. promelas
incubation time is only 4 days, compared to over 30 days for the salmonids (Scott
and Crunkilton 2000). Chronic nitrate exposure to fathead minnows produced 7-d
larval and 11-d embryo-larval LOECs (with growth as the endpoint) of
3176 mg NO3-·L-1 (Scott and Crunkilton 2000). At this exposure level, larvae were
lethargic and exhibited bent spines before death (Scott and Crunkilton 2000).
The following studies also suggest that juvenile stages of fish are not acutely
susceptible to nitrate levels commonly found in the environment. Goldfish
(Carassius carassius) and bluegills exposed to NaNO3 had very similar 24-hour
median tolerance limits (TLm) in standard reference water at 8870 and
9344 mg NO3-·L-1, respectively (Dowden and Bennett 1965). Juvenile Guadalupe
bass (Micropterus treculi), a species native to streams and rivers of central Texas,
USA, exhibited acute toxicity (96-h LC50) at 5586 mg NO3-·L-1 in hard water
(310 mg NO3-·L-1) (Tomasso and Carmichael 1986). The lethal concentration limits
for
sticklebacks
(Gasterosteus aculeatus)
30 - 50 mm
in
length
were
- -1
- -1
1348 mg NO3 ·L for exposure to NaNO3 for 10 days, and 79 mg NO3 ·L for
exposure to KNO3 (Jones 1939). Exposing guppies (Poecilia reticulatus) to KNO3,
Rubin and Elmaraghy (1977) determined that acute mortality increased with
exposure time. The median lethal concentration estimates of nitrate for the guppy fry
reared in tap water for 24 and 96 hours were 1181- and 847 mg NO3-·L-1, respectively
(Rubin and Elmaraghy 1977).
Sub-lethal, physiological endpoints in the perch (Perca fluviatilis) and the Crusian
carp (Cyprinus carassius), were also not significantly altered at environmental nitrate
concentrations. Lahti et al. (1985) found no clear relationship between nitrate levels
up to 11.0 mg NO3-·L-1 and radioiodine accumulation in organs, suggesting that
uptake of iodide (a trace element required for normal physiological functioning in fish)
(Heath 1995), is not affected at environmental levels of nitrate.
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6.3.4 Amphibians
Amphibians are susceptible to water pollution as they have permeable skin and rely
on aquatic habitats for reproduction, larval development and hibernation (Hecnar
1995). Observed toxic responses to nitrate exposure for amphibian species include
reductions in egg hatching success, increases in embryo and larval (tadpole)
mortality and developmental impacts including decreased length and weight and the
appearance of deformities (Appendix A). Amphibians are particularly sensitive
ecological receptors because they often inhabit surface waters that collect
agricultural drainage. As breeding season in the spring tends to coincide with fertilizer
application, developing eggs and embryos are placed in contact with potentially
elevated nitrate pulses (Hecnar 1995).
Primary studies on amphibian embryos and larvae (including tadpoles) exposed to
NaNO3 showed acute toxic responses (4 to 16-d LC50s) ranging from 1179 to
7752 mg NO3-·L-1, and chronic responses (10 to 56-d LOECs) from 129 to
2190 mg NO3-·L-1 (Appendix A). Of the primary studies available, red-legged frog
embryos (Rana aurora) collected from the Cascade mountains of western Oregon,
USA, were the most susceptible with a LOEC of < 29.1 mg NO3--N·L-1
(129 mg NO3-·L-1) significantly reducing overall length after 16 days exposure in soft
well water (Schuytema and Nebeker 1999b). Growth of the common northern leopard
frog (Rana pipiens) was also significantly reduced (p = 0.019) by ~2 mm over a 9
week period from exposure to 133 mg NO3-·L-1 in hard water (324 mg CaCO3·L-1)
(Allran and Karasov 2000).
Schuytema and Nebeker (1999a,c) demonstrated that younger (embryonic)
amphibian life stages can be more sensitive to nitrates than more developed larval
forms. Ninety-six hour LC50 estimates for the Pacific tree frog (Pseudacris regilla)
were 2849 and 7752 mg NO3-·L-1 for embryos and larvae, respectively (Schuytema
and Nebeker 1999a, 1999c; Appendix A). Length of developing embryos of P. regilla
was also restricted at lower nitrate levels (10-d LOEC = 492 mg NO3-·L-1) than
tadpoles (10-d LOEC = 1148 mg NO3-·L-1) (Schuytema and Nebeker 1999a,c;
Appendix A).
The African clawed frog (Xenopus laevis), a common laboratory test organism
showed toxic responses in a similar range to native North American frog species.
Five-day LOECs for X. laevis embryos were 251, 492 and 1021 mg NO3-·L-1 for
changes in weight, length and deformities, respectively (Schuytema and Nebeker
1999a). Physical deformities noted for X. laevis and P. regilla at concentrations from
492 to 4338 mg NO3-·L-1 included cardiac and abdominal edemas and lordosis
(curvature of the spine) (Schuytema and Nebeker 1999a). The chronic mortality
estimate for P. regilla larvae (10-d LC50 = 1179 mg NO3-·L-1) was ~15% of the acute
value (96-h LC50 = 7752 mg NO3-·L-1) (Schuytema and Nebeker 1999c).
Larvae of tree frogs (Litoria caerulea), and the common toad (Bufo bufo) were highly
sensitive to NaNO3 exposure in distilled water. Following 13 days of exposure to
40 mg NO3-·L-1, the mean length of exposed larvae was significantly reduced
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(p < 0.05) from approximately 25 to 17 mm and survival was reduced from 92 to 15%
(p < 0.05) (Baker and Waights 1993). Baker and Waights (1994) found no difference
in tadpole growth between 40 and 100 mg NO3-·L-1 treatments, but growth in these
treatment groups was reduced relative to controls, from approximately 43 to 20 mm
(p < 0.05). Survival was also significantly reduced from 77% in controls to 46% in
treatments (p < 0.05), and of the remaining larvae, significantly fewer had attained
the developmental Gosner stage 27 (9%) than in controls (76%; p < 0.001).
Underdeveloped larvae can be more susceptible to predation, be less able to escape
unfavourable environmental conditions, or have reduced adult body size; all of which
can ultimately reduce survival (Baker and Waights 1994).
In contrast to the studies of Baker and Waights (1993, 1994), there were no
significant effects on the proportion of eggs hatching or of deformed larvae in two
species of salamander (Ambystoma jeffersonianum and A. maculatum), the American
toad (B. americanus) or the wood frog (Rana sylvatica) when exposed to
40 mg NO3-·L-1 for a maximum of 44 days in irrigated pond water (Laposata and
Dunson 1998). A deformity that involved substantial curling of the spines to a
crescent shape, resulting in reduced swimming speeds and swimming in helical
patterns, was observed in the wood frog larvae. However, there was no statistically
significant difference in the frequency at which this physical deformity occurred
among the control and treatment groups.
Synergistic effects from other environmental stressors on amphibian egg survival are
possible, and potential interactive effects with nitrate should not be ruled out
(Laposata and Dunson 1998). Survival and activity levels in larval Cascades frogs
(Rana cascadae) from Oregon have been shown to be significantly reduced in the
presence of high levels of nitrate (20 mg NO3-·L-1), ultraviolet radiation (UV-B; 280 315 nm) and low pH (pH 5), while not being significantly affected by high nitrate
levels alone (Hatch and Blaustein 2000).
Amphibian responses to exposure from KNO3 resulted in acute toxicity at lower
concentrations than NaNO3, with 15-d LC50 estimates of 73 mg NO3-·L-1 for the
Oregon spotted frog (Rana pretiosa) and 104 mg NO3-·L-1 for the northwestern
salamander (Ambystoma gracile) (Marco et al. 1999). At higher nitrate exposures (up
to 111 mg NO3-·L-1), Marco et al. (1999) found evidence of reduced feeding activity
and swimming vigor, disequilibrium, physical abnormalities (mainly edemas and bent
tails), paralysis, and death in R. pretiosa and A. gracile. In contrast, other species
tested, namely the Western toad (Bufo boreas) and Pacific tree frog experienced very
little mortality or sub-lethal effects at all concentrations, suggesting differential
responses to nitrate exposure between amphibian species (Marco et al. 1999).
6.4 Toxicity to Marine Life
6.4.1 Algae and Plants
In a review of inhibitory concentrations of nitrogen compounds for marine and
freshwater algae, none were reported for nitrate (Admiraal 1977). The growth of ten
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species of marine benthic diatoms (expressed as a percent increase in chlorophyll a)
under varying nitrate concentrations (as KNO3) was either not inhibited, or only
slightly inhibited, even at the highest concentration tested, at 1048 mg NO3-·L-1
(Admiraal 1977). No inhibition was seen in marine diatom cultures (Nitzschia
pungens) grown at 13.6 to 54.6 mg NO3-·L-1 (Bates et al. 1993). Naidoo (1990) found
that not only did sodium nitrate have no adverse effects on the growth of the tropical
marine mangrove (Bruguiera gymnorrhiza), it actually increased total propagule
biomass, with maximum growth occurring at 44 mg NO3-·L-1.
High nitrate levels might indirectly lead to metal toxicity in marine plants and algae.
Wang and Dei (2000, 2001) found that nitrate additions to marine phytoplankton
cultures increased concentration factors for selected metals (Cd, Se, Zn) in
phytoplankton cells. Addition of ammonium nitrate fertilizer has also been observed
to cause increased cadmium accumulation in terrestrial plants such as flax (Grant et
al. 2000) Nutrient enrichment may therefore influence trace metal uptake at the base
of the food chain.
6.4.2 Invertebrates
The tropical prawn Penaeus monodon was the marine species most sensitive to
nitrate exposure (Appendix B). Penaeid larvae were exposed to potassium and
sodium nitrate salts at 1, 10 and 100 mg NO3-·L-1 in 40-h static tests (Muir et al.
1991). Significant mortality (p < 0.01) was observed at 1 mg NO3-·L-1 for both
potassium (37% mortality) and sodium (31% mortality) salts. Histological examination
of surviving larvae revealed vacuolation and shrinkage of the ganglionic neuropiles,
and minor muscle fragmentation and shrinkage. At 10 and 100 mg NO3-·L-1, effects
also included the splitting of the hypodermis from the cuticle and cytoplasmic
vacuolation of cells in the midgut and proventriculus (Muir et al. 1991).
The prawn larvae in the Muir et al. (1991) study moulted from Protozoea I to
Protozoea II stage during the trials. As crustaceans are reportedly more susceptible
to toxins during the sensitive ecdysis stage (moulting), the increased susceptibility to
nitrate found by Muir et al. (1991) is likely due to developmental sensitivity. This level
of sensitivity to nitrate exposure is not seen in older penaeid shrimp. Wickins (1976),
found that the growth of juvenile P. monodon (0.5 - 1.5 g live wt.) was not affected
after 3 to 5 weeks exposure to concentrations over 886 mg NO3-·L-1, and the
48-h LC50 for five species of penaeids (pooled) was 15 062 mg NO3-·L-1. Adult
penaeid shrimp (Penaeus paulensis) were similarly tolerant to high nitrate exposure,
with a 96-h LC50 of 9621 mg NO3-·L-1 (Cavalli et al. 1996).
Polychaetous annelids collected from the vicinity of a domestic sewage outfall in
California were exposed to KNO3 in a static 28-day test (Reish 1970). Median lethal
mortalities (28-d TLm) for the semi-healthy zone indicator species Neanthes
arenaceodentata and Dorvillea articulata were 496 and 880 mg NO3-·L-1, respectively,
and 329 mg NO3-·L-1 for Nereis grubei which are found in healthy zones surrounding
the outfalls.
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Basuyaux and Mathieu (1999) tested growth (as daily % increase in mass) and
feeding rate (g·kg-1·d-1) in sea urchin (Paracentrotus lividus) and abalone
(Haliotis tuberculata) in response to increasing nitrate concentrations (0 to
1108 mg NO3-·L-1). Safe levels resulting in 1% mortality, were determined to be
around 443 mg NO3-·L-1 for P. lividus, and between 443 and 1108 mg NO3-·L-1 for
H. tuberculata. At 1108 mg NO3-·L-1, statistically significant decreases in growth
relative to controls of 76% and 71% were seen for sea urchins and abalone,
respectively (p < 0.001). A concentration of 1108 mg NO3-·L-1 also resulted in a
statistically significant decrease in feeding rate for sea urchins of 46% (p < 0.001).
For abalone, a slight (but not statistically significant) increase in growth was seen up
to 222 mg NO3-·L-1, suggesting that this taxa may benefit from typical environmental
levels of nitrate in the sea water (Basuyaux and Mathieu 1999).
Juvenile and adult hard clams (Mercenaria mercenaria) and American oysters
(Crassostrea virginica) from the U.S. east coast (Delaware) were found to be
extremely tolerant to nitrate (Epifanio and Srna 1975). For both species, sublethal
responses (20-h ECs for reduced feeding) and acute 96-h LC50s ranged from 2480 to
> 19 840 mg NO3-·L-1 as NaNO3, suggesting that these species are insensitive to
acute exposures of environmentally relevant levels (Epifanio and Srna 1975).
Likewise, juvenile Australian crayfish (Cherax quadricarinatus) exposed to NaNO3
concentrations up to 4430 mg NO3-·L-1 in 120-h renewal tests did not exhibit any
significant differences in oxygen consumption rates or mortality during the exposure
period (Meade and Watts 1995).
6.4.3 Fish
There are few studies available on nitrate toxicity to marine fish. Frakes and Hoff Jr.
(1982) found that survival of larval anemonefish (Amphiprion ocellaris), reared in
high-nitrate conditions (~443 mg NO3-·L-1) for 72 days, was 25% lower than larvae
reared in low-nitrate treatments (~71 mg NO3-·L-1). The mean total length of juvenile
anemonefish was 8% lower under high nitrate levels and these fish had noticeably
faded coloration, decreasing their commercial marketability (Frakes and Hoff Jr.
1982).
Pierce et al. (1993) tested the responses of five tropical and sub-tropical marine fish
to increasing sodium nitrate levels in response to concern over elevated nitrate levels
in recirculating aquarium systems. All five species were tolerant to nitrate in 32‰
salinity seawater with 96-h LC50 values ranging from 2538 mg NO3-·L-1 for the
planehead filefish (Monacanthus hispidus) to > 13 290 mg NO3-·L-1 for beaugregory
(Pomacentrus leucostictus) (Pierce et al. 1993).
The nitrate concentration required to reduce first-feeding incidence by 50% after a
24-h exposure (24-h first feeding EC50) in marine fish larvae was assessed for four
species of sub-tropical fish from South Africa (Brownell 1980). Again, all four species
were found to be very tolerant to nitrate, with EC50 values ranging from 2658 to
4582 mg NO3-·L-1. A shorter exposure time (24-h) to nitrate was used in this study to
avoid potential complications with the sensitive timing to first feeding event, as
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68
prolonged toxicant exposure to marine teleost eggs and larvae can delay
development (Brownell 1980). Acute mortality (24-h LC50) values of up to
22 372 mg NO3-·L-1 were observed (Appendix B), but Brownell (1980) demonstrated
that mortality at these high levels of NaNO3 were just as likely due to the elevated
salinity of the treatment waters.
The only data located on nitrate toxicity to temperate marine fish species were for
chinook salmon and rainbow trout reared in 15‰ salinity reconstituted seawater
(Westin 1974). The salmonids were exposed to NaNO3 for 7 days, with renewal of
the test solution after 4 days, to a maximum concentration of 6500 mg NO3-·L-1,
resulting in a 7-d TLm of 4000 mg NO3-·L-1, for both species. All trout exhibited acute
signs of toxic stress after 2 days of exposure; however, chinook exposed at ≤
4400 mg NO3-·L-1 did not exhibit toxic stress symptoms until after 5 to 8 days.
Symptoms included an inability to swim upright, laboured respiration, and reduced
movement with erratic swimming. Other behavioural signs of stress included
yawning, or gulping and accelerated opercular movements, with some fish breaking
the surface of the water (Westin 1974). None of these behavioural modifications were
observed in fish from control tanks. Westin (1974) also proposed safe concentrations
of 25 to 35 mg NO3-·L-1 for hatchery-reared salmonids based on 1/100th of the
7-d LC10 at 15‰ salinity (not reported in Appendix B).
6.5 Genotoxicity of Nitrate
The carcinogenicity of the nitrate ion, nitric acid, ammonium nitrate, sodium nitrate or
potassium nitrate is not classified under the International Agency for Research on
Cancer (IARC) system (WHO 2001), or by the U.S. National Toxicology Program
(NRC 1978; NTP 2001).
Although nitrate and its associated salts are unlikely to be carcinogenic themselves,
they may be indirectly involved in mutagenesis. Suzuki et al. (1982) found that the
photolysis of aromatic compounds in the presence of an aqueous nitrate solution
(73 mg NO3-·L-1) resulted in products that were mutagenic to Salmonella typhimurium
in Ames assays, whereas no mutagenicity was found when a non-nitrate aqueous
solution was used. By carrying out these experiments in wavelengths from 250 to
577 nm and in > 300 nm, Suzuki et al. (1982) found that the majority of the
mutagenicity was induced in exposure to ultraviolet light (i.e., < 300 nm wavelength).
It is also suspected that elevated gastric pH levels (i.e., pH > 4) in mammals
(including humans) may lead to the proliferation of denitrifying bacteria that would
break down nitrate to nitrite which may ultimately form N-nitroso compounds (Packer
1995) through the following pathway:
A) nitrite is converted to nitrous acid:
NO2- + H+ ↔ HNO2
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69
B) 2 molecules of nitrous acid reversibly form one molecule of nitrous acid
anhydride:
2 HNO2 ↔ N2O3 + H2O
C) which then reacts with non-ionized secondary amines to form Nnitrosamines:
R,R’NH + N2O3 → R,R’N2O + HNO2
(from NRC 1978)
Most N-nitroso compounds are carcinogens and nitrosamines have induced cancer in
every species of animal tested, including zebra fish (Brachydanio rerio), rainbow trout
(O. mykiss), and guppy (P. reticulata); however, as little or no information exists on
environmental exposure levels or uptake and metabolic fate, any assessment of
ecological hazards will remain highly uncertain (NRC 1978; Russo 1985). In a study
of nitrosating agents present in water, levels of sodium nitrate up to 8000 mg·L-1
(= 5840 mg NO3-·L-1) were found not to induce clastogenic responses (i.e., the
induction of micronuclei in red blood cells) in newt larvae (Pleurodeles waltl), under
varying environmental factors such as pH and lighting conditions (L'Haridon et al.
1993).
6.6 Toxicity to Semi-Aquatic Animals
No studies were located on the effects of ambient nitrate concentrations on marine or
freshwater mammals or birds.
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7. CANADIAN WATER QUALITY GUIDELINES
7.1 Protection of Aquatic Life
In accordance with the CCME protocol for the derivation of water quality guidelines
for the protection of aquatic life, toxicity studies were classified as either primary,
secondary or ancillary (CCME 1991). Because the nitrate ion is non-volatile and
tends to remain in solution (NRC 1978), and studies that monitored nitrate levels over
time did not report any significant losses from their experimental systems (Muir et al.
1991; Scott and Crunkilton 2000), some studies using static test conditions were still
classified as primary. Primary and secondary studies were considered for guideline
development. As there have been no conclusive relationships drawn between nitrate
toxicity and ambient levels of various water quality variables (Scott and Crunkilton
2000), studies that did not report some variables, but had adequate survivorship in
controls, were included. Studies using distilled and/or deionized water to hold test
organisms were not included due to potential ionic influences on survival (Anderson
1944). Only studies using species resident to Canadian waters were included in the
freshwater guideline derivation. Marine species included non-native temperatedwelling organisms as per the CCME (1991) protocol. Only toxicity data for sodium
nitrate were used in deriving the freshwater guidelines, while toxicity data for both
sodium nitrate and potassium nitrate were used in deriving the marine guidelines.
The rationale for this decision was discussed in Section 6.1.
7.2 Freshwater Guideline Derivation
Nitrate is toxic to sensitive early life-stages of freshwater invertebrates, amphibians,
and fish. For members of each group, nitrate can affect embryonic or larval survival,
growth, or behaviour. Amphibian larvae and invertebrates tend to be more
susceptible to nitrate exposure than larval fish (Appendix A).
Key primary studies used in guideline derivation with environmentally relevant
endpoints included mortality, growth, physical deformities and reproduction. A series
of primary studies by Schuytema and Nebeker (1999a-c) examined the impact of
sodium nitrate addition on embryo and tadpole mortality and growth parameters for
several species of frogs endemic to North America. Exposure period had a
substantial effect on mortality for the pacific tree frog (Pseudacris regilla), with a
10-d LC50 value of 1180 mg NO3-·L-1, compared to a 4-d LC50 value of 7752 mg NO3·L-1 (Schuytema and Nebeker 1999c). NOECs and LOECs based on frog weight were
observed at < 133 mg NO3-·L-1 (Schuytema and Nebeker 1999c). Another primary
study on northern leopard frog larvae (Rana pipiens) found that 133 mg NO3-·L-1
significantly slowed the growth of larvae (F2,213 = 4.04, p = 0.019), which could have a
significant detrimental impact on the frog’s size at maturity, rate of sexual maturation,
mate selection, rate of locomotion for predator evasion and overall probability of
survival (Allran and Karasov 2000). Invertebrates were found to be similarly sensitive
to nitrate exposure. The common freshwater cladoceran, Ceriodaphnia dubia,
experienced significantly reduced neonate production (7-d LOEC) at 189 mg NO3-·L-1
(Scott and Crunkilton 2000), and the 120-h LC50 value for the caddisfly, Hydropsyche
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71
occidentalis, was 290 mg NO3-·L-1 (Camargo and Ward 1992). Although there are
demonstrable nitrate effects on fish well below 100 mg NO3-·L-1 (see Appendix A), the
lowest effect acceptable for guideline development was a 7-d LOEC of 3176 mg NO3·L-1 for fathead minnows (Pimephales promelas) (Scott and Crunkilton 2000).
The critical study used to determine the freshwater guideline for the protection of
aquatic life from nitrate was Schuytema and Nebeker (1999c). This 10-day chronic
study examined the toxicity of sodium nitrate to the Pacific treefrog (Pseudacris
regilla). Tests followed standard procedures from ASTM (1997 a,b) and solutions
were renewed daily. The following water quality parameters were monitored
throughout the tests: temperature = 22 ± 1°C, dissolved oxygen = 7.2 ± 0.1 mg·L-1,
total hardness = 58.4 ± 9.5 mg·L-1 as CaCO3, total alkalinity = 52.0 ± 7.0 mg·L-1 as
CaCO3, conductivity = 156.0 ± 15.1 µS·cm-1, and pH = 7.0-7.6. Statistically significant
decreases in weight and length (p ≤ 0.05) were seen at concentrations as low as
133 mg NO3-·L-1 and 1148 mg NO3-·L-1, respectively (Schuytema and Nebeker
1999c). The former LOEC was used in developing the guideline. The test organisms
exposed to 133 mg NO3-·L-1 experienced a mean decrease in weight of 15% when
compared with the control group. This effect is likely to have ecological significance
as predation on amphibian larvae is size-dependent (Licht 1974; Caldwell et al. 1980;
Travis 1983; Wilbur 1984; Carey and Bryant 1995; Werner 1986). Other authors have
reported that amphibian larval size decreases of 11 and 17% can affect fitness, with
observed effects including decreased juvenile survival, decreased size at maturity,
and longer time to first reproduction (Smith 1987; Berven 1990).
Several studies exposing freshwater organisms to NaNO3 had effect concentrations
below the critical study, but were not considered for guideline derivation. Reasons for
excluding each of these studies are described below.
Significant mortality was found for chinook salmon fry at 20 mg NO3-·L-1 and
Lahontan cutthroat trout eggs and fry at 20 and 30 mg NO3-·L-1, respectively
(Kincheloe et al. 1979). The authors also found morphological abnormalities in some
surviving fry, however details were not provided. Although this study clearly
demonstrated egg sensitivity to early salmonid life stages, additional egg mortalities
caused by Saprolegnia fungal infestations could not be segregated from the data by
the authors and were therefore not useable for guideline development.
Effects on growth rate, as well as size and age at metamorphosis for larvae of the
European common frog (Rana temporaria) were observed at a concentration of
22 mg NO3-·L-1 (Johansson et al. 2001). These effects were marginal, however, and
were observed with frogs from one region, but not from another. A clear doseresponse relationship was not demonstrated, as effects were only observed at the
highest concentration tested. Also, this species is not native to Canada. Due to these
various factors, the data could not be used in deriving the guideline.
Methaemoglobin in the blood of rainbow trout, which occurred at 1% in control
treatments, reached elevated levels of 21 and 27% when the fish were exposed for
11 weeks to 26 mg NO3-·L-1 [as Ca(NO3)2] and 31 mg NO3-·L-1 (as KNO3),
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respectively (Grabda et al. 1974). These increased rates of methaemoglobin
formation corresponded to a dramatic decline in hepatic tissue respiration rates (up to
48%) which, according to the authors, would result in extreme physiological stress
(Grabda et al. 1974). This study, however, was not used to derive guideline values as
there was a large range in water oxygen levels among the test aquaria (3.1 to
7.8 mg·L-1), which may have promoted the reduction of nitrate to nitrite by anaerobic
bacteria in the surrounding water for some treatments. All 20 experimental fish for
each nitrate salt treatment were held in the same aquarium, resulting in insufficient
replication. As only one test concentration was administered for each salt, it was also
not possible to determine a dose-response relationship for methaemoglobin
formation in the trout.
In a study looking at the effects of eutrophication on carp reproduction, Bieniarz et al.
(1996) exposed fertilized eggs to sodium nitrate concentrations of 15, 150 and
500 mg NO3-·L-1. The percentage of eggs hatching was significantly lower (p < 0.01)
than that in the control at all experimental concentrations, suggesting that levels
normally found in the environment may reduce the reproductive effort in carp
(Bieniarz et al. 1996). This study was not used for deriving the freshwater guideline,
however, for two reasons. First, only nominal concentrations were reported, with no
analytical confirmation of the nitrate levels in the test vessels. Second, the hatching
success in the control group was quite low, at approximately 48%. Various other
authors report hatch rates of greater than 90% for carp eggs under control conditions
(Huckabee and Griffith 1974; Mattice et al. 1981; Oyen et al. 1991; Kaur et al. 1993).
This suggests that there may have been some problem with the experimental
conditions or the condition of the test organisms in the study by Bieniarz et al. (1996).
Tadpoles of the common toad (B. bufo) and the tree frog (L. caerulea) showed
significant reductions (p < 0.05) in growth when exposed to 40 mg NO3-·L-1 for
16 days (Baker and Waights 1993, 1994, respectively). At this concentration,
significantly fewer (p < 0.05) of the surviving L. caerulea reached the Gosner
developmental stage 27 than those in controls (Baker and Waights 1994). However,
these studies were not considered for guideline development because: neither of
these species is native to Canada; distilled water was used as the test medium,
which may have placed the tadpoles under additional ionoregulatory stress; and
nitrate levels in some chambers of the 1994 study decreased by as much as 50%.
Population levels of the freshwater hydra (Hydra attenuata) declined with increasing
nitrate concentrations (up to 150 mg NO3-·L-1) after 5 days exposure, and individuals
in the highest concentration exhibited clubbed tentacles and rapid mortality (Tesh et
al. 1990). This study was not selected for guideline development as no water quality
conditions were reported for the hydra-specific growth media used to test the
organisms, and no statistical interpretations were made on differences in survival
between treatments.
Mosquito fish (Gambusia affinis) exposed to 29 mg NO3-·L-1 were shown to
significantly increase enzyme activity levels which may have been indicative of a
physiological response to combat nitrate and nitrite stress (Nagaraju and Ramana
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Rao 1983, 1985). Although physiological endpoints may be considered for secondary
data sources (CCME 1991), these studies did not provide adequate information on
experimental conditions or control mortalities, and, being tropical freshwater fish, may
not be applicable to Canadian fish physiology.
Lahti et al. (1985) also found that nitrate levels between 0.88 and 1.5 mg NO3-·L-1
were sufficient to inhibit iodine uptake in the thyroids of Crusian carp
(Cyprinus carassius), rainbow trout, and perch (Perca fluviatilis). However, when
these fish were subjected to higher nitrate levels (up to 11 mg NO3-·L-1), iodine
uptake in the thyroid appeared to be activated; therefore a clear dose-response
relationship was not established.
In a 16-day chronic study, Schuytema and Nebeker (1999b) examined the toxicity of
sodium nitrate to embryos and larvae of the red-legged frog (Rana aurora).
Statistically significant decreases in length (p ≤ 0.05) were seen at concentrations as
low as 129 mg NO3-·L-1 (Schuytema and Nebeker 1999b). The difference in length
between the control group and the test organisms exposed at 129 mg NO3-·L-1 was
only 3%. This study was not selected for guideline development because although
the LOEC was statistically significant, the effects were small enough that their
ecological significance was questionable.
A similar rationale was employed in the decision not to derive the guideline based on
the study by Allran and Karasov (2000). In this chronic study, larvae of the northern
leopard frog (Rana pipiens) were exposed to sodium nitrate. Statistically significant
reductions in length (p = 0.02) were observed at 133 mg NO3-·L-1, but the amount of
reduction, compared to control organisms, was only 6%. The authors noted that
these effects may not be ecologically significant as many other natural environmental
variables can affect amphibian larval growth to a greater degree (Allran and Karasov
2000).
7.2.1 Recommended Freshwater Guideline
The recommended interim freshwater guideline for the nitrate ion is 13 mg NO3-·L-1.
This value is based on the lowest observable adverse effect concentration of
133 mg NO3-·L-1 reported for the Pacific treefrog (Pseudacris regilla) (Schuytema and
Nebeker 1999c). A safety factor of 0.1 was applied to the LOEC in accordance with
the CCME (1991) protocol, and the final result was rounded up.
Support for this guideline can be drawn from the fact that three other studies reported
LOECs within a similar range. Decreased length was observed in larvae of the redlegged frog and the northern leopard frog at LOECs of 129 mg NO3-·L-1 (Schuytema
and Nebeker 1999a) and 133 mg NO3-·L-1 (Allran and Karasov 2000), respectively.
The water flea Ceriodaphnia dubia was similarly susceptible, with a 7-d LOEC for
reduced reproduction at 189 mg NO3-·L-1 (Scott and Crunkilton 2000). The guideline
is also comparable with estimates that have been made for safe nitrate
concentrations for invertebrates. By converting 72-, 96- and 120-h mortality data to
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probit values and then to LC0.01s, Camargo and Ward 1995) calculated lifetime safe
concentrations for hydropsychid larvae (= 8760-h LC0.01) of 6.2 to 15.5 mg NO3-·L-1.
Temperate streams and rivers draining regions that support intensive agricultural
production and human populations can often exceed 13 mg NO3-·L-1 in Canada, the
U.S., and Europe (Rouse et al. 1999; Van Herpe and Troch 2000; Goolsby et al.
2001; OMOE 2001). If surface water nitrate concentrations approach guideline levels,
site-specific monitoring of ecological effects is recommended.
The Canadian Water Quality Guideline for the Protection of Aquatic Life for nitrate
was derived solely from data on direct toxic responses to freshwater organisms and
is not intended to protect against potential indirect toxic effects. Nitrate is only one of
the forms of inorganic nitrogen that can be taken up by primary producers, and
therefore other forms of nitrogen may also contribute to eutrophication effects. An
examination of the role of nitrogen and nitrogen-to-phosphorus ratios in
eutrophication processes in freshwater is presented in a separate discussion paper
(CCME 2002).
7.2.2 Data Gaps / Research Recommendations
The current freshwater data set of acceptable primary and secondary acute nitrate
toxicity studies contains five species of fish, three species of amphibians and four
species of invertebrates (Appendix A). Chronic studies include three species of fish,
four species of amphibians and two species of invertebrates (all planktonic)
(Appendix A).
To fulfill the requirements for full freshwater guideline status, one chronic invertebrate
study on a non-planktonic organism is required. As nitrate is a required nutrient to
stimulate plant growth, the plant toxicity requirements of the CCME protocol (CCME
1991) have been waived. This rationale has also been adopted by Australia and
New Zealand (Environment Australia 2000b).
To ensure that the nitrate freshwater guideline is sufficiently protective of all aquatic
organisms in Canada, it is recommended that additional toxicity tests be conducted
for fish and invertebrate species that are known to be highly sensitive. For example,
although toxicity data was available for caddisflies, generally mayflies and stonesflies
are considered more sensitive to contaminants; therefore, nitrate toxicity tests with
these other invertebrates would be useful. Effects of nitrate on brook trout,
particularly the egg and juvenile stages, should be studied as the spawning habits of
this species could make it particularly susceptible. Further investigation of nitrate
toxicity to fish eggs, in general, is also needed as this may be a particularly sensitive
life stage. For example, two ancillary studies (Kincheloe et al. 1979; Bieniarz et al.
1996) reported adverse effects on fish eggs at concentrations lower than the critical
study on which the guideline was based.
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Potential influences of other parameters on the toxicity of nitrate are currently not well
understood. Further research is needed on the interactions of nitrate with potassium,
ammonia, UV and low pH.
As discussed in Section 6.1, toxicity tests using potassium nitrate and ammonium
nitrate were excluded from derivation of the freshwater guidelines because the
greater toxicity observed in these studies was likely due to the K+ and NH4+ ions,
rather than the NO3-. Where the main inputs of nitrate to a freshwater system are in
the form of KNO3 and/or NH4NO3, adhering to the nitrate guidelines alone may not
protect against adverse effects. The Canadian Water Quality Guidelines for ammonia
should be followed to protect against effects from NH4NO3 (CCME 2000). We would
also recommend that a Canadian Water Quality Guideline be developed to protect
freshwater aquatic life from the adverse effects of potassium.
An interesting area for future research would be to pursue field validation of the
guideline. Such validation would need to be conducted in areas where nitrate does
not co-occur with other contaminants, such as those found in sewage or animal
wastes, as these could have an additional effect on the aquatic community beyond
the effect attributable to nitrate.
7.2.3 Summary of Existing Guidelines
The previous Canadian water quality guidelines for nitrate, developed in 1987,
consisted of a narrative statement that nitrate concentrations which will stimulate
weed growth should be avoided (CCREM 1987). British Columbia is the only
Canadian jurisdiction to have developed guidelines for the protection of aquatic life
from nitrate toxicity, with a maximum exposure of 200 mg NO3--N·L-1
(= 886 mg NO3-·L-1) and a 30-day average exposure of 40 mg NO3--N·L-1
(= 177 mg NO3-·L-1) (Nordin and Pommen 1986). These values were based on 50%
and 10%, respectively of the lowest 96-h LC50 reported in the literature (Nordin and
Pommen 1986). Québec has also adopted these values for provincial guideline use
(MEF 1998). Alberta’s surface water quality guidelines have a maximum allowable
concentration for total nitrogen (total inorganic plus total organic) of 1.0 mg N·L-1,
however this nitrate concentration is not considered directly toxic, rather the guideline
is to protect against deleterious influences of nitrate on conditions that affect aquatic
life (AEP 1999).
The freshwater CWQG for the protection of aquatic life greatly exceeds the moderate
reliability trigger value developed by Australia and New Zealand (0.70 mg NO3-·L-1)
(Environment Australia 2000b). The Australian/New Zealand guideline was derived
by applying a default acute-to-chronic ratio (ACR) to the 95% distribution of toxicity
data for potassium and sodium nitrate salts which included native Australian fish and
invertebrates (Environment Australia 2000b). Australian/New Zealand guidelines also
propose that a nitrate concentration of 17 mg NO3-·L-1, similar to the proposed
Canadian guidelines, would be protective of 80% of the freshwater species
(Environment Australia 2000a). This level however, exceeds a 9-d NOEC of
14 mg NO3-·L-1 for the Australian freshwater fish Mogurnda adspersa, and therefore
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was deemed not to be protective for some key Australian species (Environment
Australia 2000b).
The US EPA does not currently have a formal guideline for nitrate for the protection
of aquatic life. Based on observations by Knepp and Arkin (1973), however, the
US EPA
has
suggested
that
nitrate
levels
below
90 mg NO3--N·L-1
- -1
(= 399 mg NO3 ·L ) will be protective of warmwater fish (US EPA 1986).
Currently the European Union has no nitrate guideline for the protection of aquatic
life. The Netherlands, however has proposed a maximum allowable concentration for
nitrate of 2.0 mg NO3-·L-1 in eutrophic waters to protect against direct toxicity
(Speijers et al. 1989). In addition, the Netherlands recommends a maximum
allowable nitrate concentration of 0.04 mg NO3-·L-1 in oligotrophic waters to protect
against eutrophication impacts (Speijers et al. 1989).
7.3 Marine Guideline Development
There are relatively few studies available on nitrate toxicity to marine fish. With the
exception of Westin (1974), those which do exist are on tropical or subtropical
species (Brownell 1980; Frakes and Hoff Jr. 1982; Pierce et al. 1993). There appears
to be a greater body of information on nitrate toxicity responses from commercially
important marine invertebrates in aquaculture operations, such as prawns, crayfish
and bivalves (Epifanio and Srna 1975; Wickins 1976; Muir et al. 1991; Meade and
Watts 1995). Toxic responses to marine organisms include mortality, reductions in
feeding and growth, and physiological responses such as respiration and cellular
changes. As with freshwater animals, invertebrates, especially during larval stages,
tend to be more sensitive to nitrate than fish (see Appendix B).
Mortality responses in marine fish tend to be similar to that of freshwater fish, with
acute and chronic LC50 values ranging from approximately 2500 to
10 600 mg NO3-·L-1 (Appendix B). In a direct comparison between freshwater and
marine conditions, Westin (1974), exposed rainbow trout and chinook salmon to
sodium nitrate at concentrations ranging between 3500 and 6500 mg NO3-·L-1 in
freshwater and 15‰ salinity reconstituted seawater for 96 hours. It was found that
nitrate was 1.3 times more toxic in saltwater for both species; however, no
explanation was given for the increase in toxicity with increasing salinity.
The following three primary studies were acceptable for guideline derivation. Growth
(as % daily weight gain) of the purple sea urchin (Paracentrotus lividus) and the
abalone (Haliotis tuberculata) was statistically significantly reduced at an exposure to
1108 mg NO3-·L-1 (p < 0.001) for 15 days, but was also substantially reduced for
P. lividus at 443 mg NO3-·L-1 (Basuyaux and Mathieu 1999). The larval tiger shrimp
(Penaeus monodon) was the most sensitive marine test organism found, with
statistically significant increases in mortality occurring at 1 mg NO3-·L-1 when exposed
to either sodium or potassium salts (Muir et al. 1991). At this level of exposure, Muir
et al. (1991) also found various sublethal histopathological changes that could
decrease the fitness of larvae, thereby decreasing their chances of survival. Although
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P. monodon was the most sensitive marine receptor species, the Muir et al. (1991)
study was not used to derive the guideline value as this is a non-resident, tropical
prawn species. The only primary vertebrate study examined the ability of tropical and
subtropical fish to withstand high nitrate concentrations in captive, recirculating water
(Pierce et al. 1993). Acute mortalities (96-h LC50s) for fish exposed to NaNO3 ranged
from 2538 mg NO3-·L-1 to > 13 290 mg NO3-·L-1 (see Appendix B; Pierce et al. 1993).
The following secondary studies were considered for guideline development. Four
species of subtropical fish larvae reared from ocean-collected pelagic eggs and
exposed to nitrate showed delayed development to first feeding after yolk sac
absorption, but only at very high concentrations, with 24-h EC50s ranging from 2658
to 4582 mg NO3-·L-1 (Brownell 1980). In addition, Brownell (1980) demonstrated that
at nitrate concentrations required to elicit mortality in 24h LC50s (15 280 to
22 370 mg NO3-·L-1), the toxic response could just as easily be attributed to elevated
salinity, rather than the nitrate ion.
Chinook salmon (O. tshawytscha) and rainbow trout (O. mykiss) were the only
temperate fish species endemic to Canadian marine or brackish waters, for which
toxicity data were located. Fingerlings were exposed to NaNO3 in static (96-h) or
static-renewal (7-d) tests in brackish waters (15‰ salinity) (Westin 1974). Both
salmonid species had very similar median tolerance limits (96-h and 7-d TLm),
ranging from 4000 to 4650 mg NO3-·L-1.
Interim guidelines for the protection of marine life were derived, as both primary and
secondary data were included in the minimum dataset requirements (CCME 1991).
The key secondary study used for guideline development exposed temperate marine
adult polychaetes (Phylum: Annelida) to potassium nitrate as part of an effort to
determine their susceptibility to inorganic factors present at marine sewage outfalls
(Reish 1970). This static test was conducted in seawater with 19.2‰ chlorinity,
5.9 ppm dissolved oxygen, and a temperature range of 22° to 25°C. Of the three
species of polychaetes with acceptable control mortality, the lowest 28-d TLm (= LC50)
was 5.3 mg-at·L-1 (329 mg NO3-·L-1) for Nereis grubei (Reish 1970). This species is
also indicative of healthy zones surrounding sewage outfalls and is generally not
found directly beneath the outfall zone (Reish 1970).
7.3.1 Recommended Marine Guideline
The recommended interim marine guideline for the nitrate ion is 16 mg NO3-·L-1. This
value is based on a 28-day median lethal concentration of 329 mg NO3-·L-1 for the
temperate marine polychaete Nereis grubei (Reish 1970). A safety factor of 0.05 was
applied to the LC50. The CCME (1991) protocol for deriving water quality guidelines
recommends a safety factor of 0.1 for guidelines derived from a chronic study, and a
safety factor of 0.01 for guidelines derived from an acute study. An intermediate
safety factor of 0.05 was chosen for this guideline because, although it is based on a
chronic study, the endpoint was an LC50; therefore, low levels of mortality would have
been observed at concentrations less than 329 mg NO3-·L-1, and sublethal effects
may have occurred at even lower concentrations. The authors of the critical study
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noted that the test organisms used were of adult size and, as the early larval stage is
the most sensitive phase in the life history of marine invertebrates (Thorson 1956),
therefore may not have represented the most conservative estimates of toxicity
(Reish 1970). Further support for a conservative safety factor comes from the fact
that Muir et al. (1991) observed mortality effects for juvenile tropical prawns at
1 mg NO3-·L-1. Although these sensitive tropical prawns are not found in Canadian
marine waters, this study flags the possibility that there may be temperate species
with similarly high sensitivity to nitrate for which toxicity tests have not yet been
conducted.
The Canadian Water Quality Guideline for the Protection of Aquatic Life for nitrate
was derived solely from data on direct toxic responses to marine organisms and is
not intended to protect against potential indirect toxic effects. Nitrate is only one of
the forms of inorganic nitrogen that can be taken up by primary producers, and
therefore other forms of nitrogen may also contribute to eutrophication effects. An
examination of the role of nitrogen and nitrogen-to-phosphorus ratios in
eutrophication processes in marine waters is presented in a separate discussion
paper (CCME 2002).
7.3.2 Data Gaps / Research Recommendations
The current marine dataset of acceptable primary
studies contains eleven species of fish from three
invertebrates from one study (Appendix B). Chronic
species of fish from one study, and four species of
(Appendix B).
and secondary acute toxicity
studies, and three species of
studies include two temperate
invertebrates from two studies
To fulfill the requirements for full marine guideline status, at least one primary chronic
fish study on one other temperate species is required. The requirements for
invertebrates have been met, however, as the distributions of both the purple sea
urchin (P. lividus) and the abalone (H. tuberculata) lie within European waters, and
the three polychaete species were obtained from the Californian coast, it would be
preferable to include at least one study on a marine invertebrate endemic to
Canadian waters. As nitrate is a required nutrient for plant growth, no marine plant
toxicity studies were required for guideline development.
7.3.3 Summary of Existing Guidelines
Australia and New Zealand have adopted their moderate reliability freshwater
guideline of 0.70 mg NO3-·L-1 as the marine low reliability trigger value (Environment
Australia 2000b). Although a low reliability trigger level of 13 mg NO3-·L-1 for marine
animals was derived using an uncertainty factor of 200, the more conservative
moderate reliability freshwater value was adopted according to protocol (Environment
Australia 2000b). The Netherlands have proposed a maximum acceptable
concentration of 0.4 mg NO3-·L-1 (Speijers et al. 1989). This value is based on a
recommended limit of 0.1 mg N·L-1 to prevent eutrophication impacts, and the
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assumption that all nitrogen present is in the form of nitrate. This level is also deemed
protective against direct toxicity to marine organisms (Speijers et al. 1989).
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8. GUIDANCE ON APPLICATION OF THE GUIDELINES
8.1 General Guidance on the Use of Guidelines
Canadian Water Quality Guidelines (CWQGs) are numerical concentrations or
narrative statements that are recommended as levels that should result in negligible
risk of adverse effects to aquatic biota. As recommendations, the CWQGs are not
legally enforceable limits, though they may form the scientific basis for legislation or
regulation at the provincial, territorial, or municipal level. CWQGs may also be used
as benchmarks or targets in the assessment and remediation of contaminated sites,
as tools to evaluate the effectiveness of point-source controls, or as “alert levels” to
identify potential risks.
CWQG values are calculated conservatively, such that they protect the most
sensitive life stage of the most sensitive aquatic life species over the long term.
Hence, concentrations of a parameter that are less than the applicable CWQG are
not expected to cause any adverse effect on aquatic life. Concentrations that exceed
the CWQGs, however, do not necessarily imply that aquatic biota will be adversely
affected, or that the water body is impaired; the concentration at which such effects
occur may differ depending on site-specific conditions. Where the CWQGs are
exceeded, professional advice should be sought in interpreting such results. As with
other CWQGs, the guidelines for nitrate are intended to be applied towards
concentrations in ambient surface waters, rather than immediately adjacent to point
sources such as municipal or industrial effluent outfalls. Various jurisdictions provide
guidance on determining the limits of mixing zones when sampling downstream from
a point source (see, for example, BC MELP 1986 and MEQ 1991), though
Environment Canada and the CCME do not necessarily endorse these methods.
8.2 Monitoring and Analysis of Nitrate Levels
In comparing field measurements of nitrate to the Canadian water quality guidelines,
it is important to be aware of potential seasonal and meteorological impacts at the
time of sampling. Nitrate concentrations in surface waters can peak for short periods
of time during storm events and spring melt. As these pulses often occur in the spring
when the most sensitive live stages (e.g., larvae) for many organisms are present,
their relationship to the guideline should be considered. A stream may normally have
a low baseline concentration of nitrate, but during and immediately following (1-2
days) one of these events, the nitrate concentrations could exceed the guideline
value. The exceedance could result from one of two scenarios. First, the increase in
nitrate could occur as a result of a natural increase in background levels, for example
due to snow melt in a pristine area. Second, the source of the nitrate in storm- or
meltwater may not be natural; for example, it could be due to runoff from agricultural
fields where nitrate fertilizer has been applied, or due to greater inputs from
combined sewer overflows. In the former case the guidelines do not strictly apply
(because a guideline cannot be set lower than natural background levels for a
naturally occurring substance). Nonetheless, we recommend that if nitrate levels are
found to exceed the recommended interim guideline values, that data on the
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frequency and severity of the exceedances should be evaluated on a site-specific
basis to determine whether they warrant any preventative or remedial actions.
For monitoring long-term temporal trends in nitrate levels, an undue weighting should
be not be given to samples that were collected during, or immediately following a
storm event, or during the spring thaw. Due to seasonal variability in nitrate levels,
comparison of long-term trend data should occur between standardized collection
intervals over similar time periods (i.e, spring, summer, fall, winter).
Depending on the analytical methods used, water samples are sometimes analysed
for the total concentration of nitrate plus nitrite. In most cases, these measured
nitrate + nitrite concentrations consist almost entirely of nitrate, and therefore may be
directly compared to the guidelines recommended in this document, which are given
for concentrations of nitrate. Most natural ambient waters are sufficiently aerobic that
nitrite concentrations are negligible, with the nitrite being readily oxidised to nitrate by
nitrifying bacteria (NRC 1978; Halling-Sorensen and Jorgensen 1993). Where direct
comparison might not be appropriate, due to the possibility of elevated levels of
nitrite, is with water samples obtained from highly reducing environments. Low redox
potentials (Eh) which would promote nitrite formation are associated with elevated pH
and waters nearing anoxia (Figures 8.1 a,b). These conditions are often found at the
sediment-water interface, at the bottom of permanently stratified meromictic lakes, or
in bogs and bog lakes with very high levels of reducing humic acids (Wetzel 2001).
8.3 Developing Site-Specific Objectives
There are some cases in which the development of site-specific objectives for nitrate
should be considered. The guidelines were derived to be protective of the most
sensitive species for which toxicity data is available, with an extra margin of safety to
account for the unknown, such as the possibility of an unstudied species that is more
sensitive. Nonetheless, in spite of the safety factor, in locations where highly
sensitive or endangered species occur, managers may wish to consider the use of a
more conservative site-specific objective. Conversely, where certain sensitive
species are historically absent, the use of less conservative site-specific objectives
for those particular areas could be justified. For example, the critical study used to
derive the freshwater guideline is based on the toxicity of nitrate to amphibians. In
water bodies where amphibians are known not to occur naturally, and which are not
likely to become amphibian habitat, managers might consider developing a sitespecific objective that is based on the most sensitive fish or invertebrate species
instead.
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a)
Eh
b)
pE
Oligotrophic
+1.0
+15
NO3-
+10
+5
NO2-
Eutrophic
0
O2
Depth
+0.5
Eh
0
Eh
O2
NH4+
NH3
-5
-0.5
4
7
10
pH
Eh, O2
Figure 8.1 a) Redox potential (Eh) and electron potential (pE) for various species of
inorganic nitrogen, as a function of pH (note: N2 is treated as a redox inert
compound). b) Generalized vertical distribution of redox potential and
dissolved oxygen in stratified lakes of very low and very high productivity.
[from a) Stumm and Morgan 1981; b) Wetzel 2001]
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Managers of surface water bodies where there are groundwater upwellings should
note that elevated levels of nitrate (i.e., above the recommended guideline values) in
the immediate vicinity of the upwelling could pose a potential risk to some aquatic life.
In particular, brook trout, and other fish species that seek out groundwater upwelling
areas for spawning may be at risk. At present there are no existing nitrate toxicity
data available for brook trout, so comments cannot be made about the sensitivity of
this species. It is possible that brook trout eggs are more susceptible to nitrate toxicity
than other fish eggs discussed in this document (e.g., fathead minnow, rainbow trout,
salmon), because they have a longer incubation period (Morris 2001). Also, hatching
of brook trout eggs occurs in March and April when groundwater levels of nitrate
peak. In brook trout spawning areas, managers may want to consider setting more
conservative site-specific nitrate objectives.
General guidance on the site-specific application of Canadian Water Quality
Guidelines is currently being drafted for the CCME (MacDonald et al. 2002).
8.4 Trophic Status Management
The nitrate WQGs developed in this document are intended to protect aquatic life
from direct toxic effects. Nitrate concentrations that are below these levels, however,
may still contribute to increased primary production within a waterbody, and could
therefore result in indirect toxic effects that are associated with eutrophication. Due to
the wide range in responses seen in algal biomass and species composition as a
result of increased nitrate supply, and the simultaneous influence of other factors in
regulating primary production (e.g., phosphorus levels, light availability, water
retention times), it may not be feasible to propose threshold levels for inorganic nitrogen
in fresh waters which will protect against nuisance algal growth (CCME 2002). To
assess the role of nitrate in regulating production in a specific waterbody, nitrogen-tophosphorus ratios could be used to first determine potential nutrient limitation,
followed by nutrient bioassays with resident water sources to determine the impact
from increased nitrate levels (see CCME 2002).
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Van Herpe, Y. and P. A. Troch. 2000. Spatial and temporal variations in surface water nitrate
concentrations in a mixed land use catchment under humid termperate climatic conditions.
Hydrol. Process. 14: 2439-2455.
Verstraete, W. and M. Alexander. 1973. Heterotrophic nitrification in samples of natural
ecosystems. Environ. Sci. Technol. 7(1): 39-42.
Viets, F. G., Jr. 1965. The plant's need for and use of nitrogen. In: Soil Nitrogen.
Bartholomew, W. V. and F. E. Clark (eds.). American Society of Agronomy, Inc. Madison,
WI. pp. 503-548.
Wang, W.-X. and R. C. H. Dei. 2000. Metal uptake in a coastal diatom influenced by major
nutrients (N, P, and Si). Wat. Res. 35(1): 315-321.
Wang, W.-X. and R. C. H. Dei. 2001. Effects of major nutrient additions on metal uptake in
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Wassenaar, L. I. 1994. Evaluation of the origin and fate of nitrate in the Abbotsford aquifer
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Watson, S. W., F. W. Valois, and J. B. Waterbury. 1981. The family nitrobacteraceae. In:
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Berlin. pp. 1005-1022.
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Webb, J. R., B. J. Cosby, F. A. Deviney, Jr., K. N. Eshleman, and J. N. Galloway. 1995.
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103
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Science-based Solutions No. 1-6
104
APPENDIX A. SUMMARY OF FRESHWATER TOXICITY STUDIES.
Organism
INVERTEBRATES
Ceriodaphnia dubia
(water flea)
Cheumatopsyche pettiti
(caddisfly)
Life Stage
Neonates
Na
+
Neonates
Na
+
Neonates
Na
+
Early Instar
Na
Last Instar
Endpoint
7-d NOEC
(reproduction)
7-d LOEC
(reproduction)
48-h LC50
+
Na
+
Early Instar
Na
Last Instar
Effect
Test
concentration Type*
- -1
(mg NO3 ·L )
94
R
189
R
1657
R
8760-h LC0.01
11
S
8760-h LC0.01
16
S
+
120-h LC0.01
30
S
Na
+
120-h LC0.01
43
S
Early instar
Na
+
120-h LC50
472
S
Early instar
Na
+
96-h LC50
503
S
Last instar
Na
+
120-h LC50
527
S
Last instar
Na
+
96-h LC50
733
S
Early instar
Na
+
72-h LC50
846
S
Na
+
72-h LC50
930
S
Last instar
Daphnia magna
(water flea)
Cation
K
+
96-h TLm
24
S
ND
K
+
48-h TLm
299
S
ND
Na
96-h TLm
485
S
ND
K
96-h TLm
549
S
7-d NOEC
(reproduction)
48-h LC50
1586
R
2047
R
ND
+
+
Neonates
Na
+
Neonates
Na
+
Temp
(°C)
Hardness Alkalinity
DO
-1
-1
-1
(mg·L )
(mg·L ) (mg·L )
pH
Reference
25 ± 1.0 7.9 - 8.3 156 - 172 140 - 170 7.9 - 8.3 Scott and
Crunkilton 2000
25 ± 1.0 7.9 - 8.3 156 - 172 140 - 170 7.9 - 8.3 Scott and
Crunkilton 2000
25 ± 1.0 7.9 - 8.3 156 - 172 140 - 170 7.9 - 8.3 Scott and
Crunkilton 2000
18
9.6
42.7
35
7.9
Camargo and
Ward 1995
18
9.6
42.7
35
7.9
Camargo and
Ward 1995
18
9.6
42.7
35
7.9
Camargo and
Ward 1995
18
9.6
42.7
35
7.9
Camargo and
Ward 1995
18
9.6
42.7
35
7.9
Camargo and
Ward 1992
18
9.6
42.7
35
7.9
Camargo and
Ward 1992
18
9.6
42.7
35
7.9
Camargo and
Ward 1992
18
9.6
42.7
35
7.9
Camargo and
Ward 1992
18
9.6
42.7
35
7.9
Camargo and
Ward 1992
18
9.6
42.7
35
7.9
Camargo and
Ward 1992
ND
ND
ND
ND
ND
Dowden and
Bennett 1965
ND
ND
ND
ND
ND
Dowden and
Bennett 1965
ND
ND
ND
ND
ND
Dowden and
Bennett 1965
ND
ND
ND
ND
ND
Dowden and
Bennett 1965
25 ± 1.0 7.9 - 8.3 156 - 172 140 - 170 7.9 - 8.3 Scott and
Crunkilton 2000
25 ± 1.0 7.9 - 8.3 156 - 172 140 - 170 7.9 - 8.3 Scott and
Crunkilton 2000
Science-based Solutions No. 1-6
Ranking** Notes
1
1
1
1
a
1
a
1
a
1
a
1
1
1
1
1
1
A
a,b,c
A
a,b,c
A
a,b,c
A
b,c
1
1
105
Organism
Life Stage
Hydropsyche occidentalis
(caddisfly)
Lymnea spp.
(snail)
Endpoint
Effect
Test
concentration Type*
- -1
(mg NO3 ·L )
2614
S
Temp
(°C)
Hardness Alkalinity
DO
-1
-1
-1
(mg·L )
(mg·L ) (mg·L )
pH
Na
+
48-h TLm
ND
Na
+
96-h TLm
3070
S
Neonates
Na
+
R
Early Instar
Na
3650
S
Early Instar
Na
+
6205
S
Adult
Na
+
50
S
ND
ND
ND
ND
ND
Adult
+
7-d LOEC
(reproduction)
48-h EC
(immobilization)
16-h EC
(immobilization)
12-d LOEC
(mortality)
13-d NOEC
(mortality)
8760-h LC0.01
3176
+
150 - 250
S
ND
ND
ND
ND
6
S
18
9.6
42.7
ND
Hydra attenuata
(hydra)
Cation
K
ND
ND
ND
ND
ND
Dowden and
Bennett 1965
ND
ND
ND
ND
ND
Dowden and
Bennett 1965
25 ± 1.0 7.9 - 8.3 156 - 172 140 - 170 7.9 - 8.3 Scott and
Crunkilton 2000
2+
25
ND
[Ca ] = 31 97 - 100
ND
Anderson 1946
-1
mg·L
25
ND
ND
97 - 100
ND
Anderson 1944
A
b,c
1
c
Tesh et al. 1990
A
c,d
ND
Tesh et al. 1990
A
c,d
35
7.9
Camargo and
Ward 1995
Camargo and
Ward 1995
Camargo and
Ward 1995
Camargo and
Ward 1995
Camargo and
Ward 1992
Camargo and
Ward 1992
Camargo and
Ward 1992
Camargo and
Ward 1992
Camargo and
Ward 1992
Camargo and
Ward 1992
Dowden and
Bennett 1965
Dowden and
Bennett 1965
Dowden and
Bennett 1965
Dowden and
Bennett 1965
1
a
1
a
1
a
1
a
Last Instar
Na
8760-h LC0.01
10
S
18
9.6
42.7
35
7.9
Early Instar
Na
+
120-h LC0.01
20
S
18
9.6
42.7
35
7.9
Last Instar
Na
+
120-h LC0.01
29
S
18
9.6
42.7
35
7.9
Early instar
Na
+
120-h LC50
290
S
18
9.6
42.7
35
7.9
Last instar
Na
+
120-h LC50
342
S
18
9.6
42.7
35
7.9
Early instar
Na
+
96-h LC50
431
S
18
9.6
42.7
35
7.9
Last instar
Na
+
96-h LC50
483
S
18
9.6
42.7
35
7.9
Early instar
Na
+
72-h LC50
658
S
18
9.6
42.7
35
7.9
Last instar
Na
+
72-h LC50
813
S
18
9.6
42.7
35
7.9
K
+
96-h TLm
671
S
ND
ND
ND
ND
ND
Eggs
K
+
48-h TLm
910
S
ND
ND
ND
ND
ND
Eggs
Na
+
96-h TLm
2373
S
ND
ND
ND
ND
ND
Na
+
48-h TLm
4716
S
ND
ND
ND
ND
ND
Science-based Solutions No. 1-6
a,c
A
+
Eggs
A
c
+
Eggs
Ranking** Notes
A
Na
Early Instar
Reference
1
1
1
1
1
1
A
a,b,c
A
a,b,c
A
a,b,c
A
a,b,c
106
Organism
Macrobrachium rosenbergii
(prawn)
Polycelis nigra
(planaria)
FISH
Carassius carassius
(crucian carp)
Cyprinus carpio
(common carp)
Gambusia affinis
(mosquito fish)
Gasterosteus aculeatus
(stickleback)
Ictalurus punctatus
(channel catfish)
Lepomis macrochirus
(bluegill)
Life Stage
Cation
Endpoint
Effect
Test
concentration Type*
- -1
(mg NO3 ·L )
709
S
Na
+
21-d LC50
Juvenile
Na
+
775
ND
Na
+
21-d EC50
(growth)
48-h LC50
ND
Na
+
Juvenile
Na
+
ND
Na
+
Egg
Na
+
Sperm
Na
+
Juvenile
Na
+
Juvenile
Na
+
Juvenile
+
ND
K
ND
Na
+
Juvenile
none
Juvenile
none
fingerlings
Na
Juvenile
ND
Juvenile
+
pH
Reference
Ranking** Notes
Sat
ND
ND
ND
Wickins 1976
2
e
S
28.0
Sat
ND
ND
ND
Wickins 1976
2
e
2666
S
15 - 18
ND
ND
ND
6.4
Jones 1941
A
c,f
48-h LC50
2697
S (R?)
15 -18
ND
ND
ND
6.4
Jones 1940
A
c,f
64-d LOEC
(iodine uptake
inhibition)
24-h TLm
0.9
ND
5-6
ND
ND
ND
ND
Lahti et al. 1985
A
c,g
8870
S
ND
Sat
ND
ND
7.9
A
b,c
15
R
ND
8-9
300 - 310
ND
7.5
Dowden and
Bennett 1965
Bieniarz et al.
1996
A
,j
8860
S
4
ND
ND
ND
ND
2
n
29
S
ND
Sat
ND
ND
ND
A
c,e
29
S
ND
Sat
ND
ND
79
R
15 - 18
ND
ND
1348
R
15 - 18
ND
> 400
R
ND
>1280
R
6200
5-d LOEC
(hatching
success)
2-h LOEC
(reduced
motility)
96-h LOEC
(enzyme
induction)
96-h LOEC
(enzyme
induction)
10-d NOEC
(mortality)
10-d NOEC
(mortality)
164-d LOEC
(growth, feeding)
10-wk LOEC
(physiological)
96-h LC50
Epler et al.
2000
A
c,e
ND
Nagaraju and
Ramana Rao
1985
ND
Nagaraju and
Ramana Rao
1983
6.0 - 6.8 Jones 1939
A
h,i
ND
ND
6.0 - 6.8 Jones 1939
A
h,i
ND
ND
ND
A
i
26
6.1 - 6.8
ND
ND
A
g
S
22,26,30
Sat
102
220
22 ± 1.0 4.8 - 8.3
46 - 49
50 - 58
ND
ND
45 - 50
51 - 56
+
96-h LC50
1840
S
K
+
24-h TLm
3355
ND
96-h LC50
8753
S
+
Hardness Alkalinity
DO
-1
-1
-1
(mg·L )
(mg·L ) (mg·L )
28.0
K
Na
Temp
(°C)
ND
ND
22 ± 1.0 4.6 - 6.6
Science-based Solutions No. 1-6
ND
Knepp and
Arkin 1973
6.4- 6.7 Collins et al.
1976
8.6 - 8.8 Colt and
Tchobanoglous
1976
7.5 - 8.4 Trama 1954
ND
Dowden and
Bennett 1965
7.4 - 8.8 Trama 1954
2
2
h
A
b,c
2
107
Organism
Life Stage
ND
Cation
Na
+
Micropterus salmoides
(largemouth bass)
Micropterus treculi
(Guadalupe bass)
Juvenile
none
Juvenile
Na
+
Oncorhynchus kisutch
(coho salmon)
Egg
Na
+
Fry
Na
+
Egg
Na
+
Fry
Na
+
Juvenile
Na
+
Egg
Na
+
Fry
Na
+
Oncorhynchus mykiss
(steelhead trout)
Oncorhynchus mykiss
(rainbow trout)
Oncorhynchus tshawytscha
(chinook salmon)
Perca fluviatilis
(perch)
Pimephales promelas
(fathead minnow)
2-yr olds
Ca
2-yr olds
K
fingerlings
fingerlings
Fry
2+
+
+
Na
+
Na
+
Na
+
Egg
Na
fingerlings
fingerlings
Juvenile
Na
+
Na
+
Na
Larvae
Na
+
Larvae
Na
+
Larvae
Na
+
+
Endpoint
24-h TLm
Effect
Test
concentration Type*
- -1
(mg NO3 ·L )
9344
ND
Temp
(°C)
Hardness Alkalinity
DO
-1
-1
-1
(mg·L )
(mg·L ) (mg·L )
ND
ND
ND
ND
pH
Reference
Dowden and
Bennett 1965
ND
ND
Knepp and
Arkin 1973
183 -163 7.9 - 8.4 Tomasso and
Carmichael
1986
25
6.2
Kincheloe et al.
1979
25
6.2
Kincheloe et al.
1979
25
6.2
Kincheloe et al.
1979
25
6.2
Kincheloe et al.
1979
ND
ND
Lahti et al. 1985
164-d LOEC
(growth, feeding)
96-h LC50
> 400
R
ND
ND
ND
5586
S
22
Sat
222 - 203
> 30-d LOEC
(survivorship)
> 30-d LOEC
(survivorship)
> 30-d LOEC
(survivorship)
> 30-d LOEC
(survivorship)
64-d LOEC
(iodine uptake
inhibition)
> 30-d LOEC
(survivorship)
> 30-d LOEC
(survivorship)
77-d EC
(physiological)
77-d EC
(physiological)
7-d LC50
96-h LC50
> 30-d LOEC
(survivorship)
> 30-d LOEC
(survivorship)
7-d LC50
96-h LC50
64-d LOEC
(iodine uptake
inhibition)
7-d NOEC
(growth)
7-d LOEC
(growth)
7-d NOEC
(mortality)
>20
F
10
ND
8 - 10
>20
F
10
ND
8 - 10
5
F
10
ND
8 - 10
>20
F
10
ND
8 - 10
1.5
ND
5-6
ND
ND
10
F
10
ND
8 - 10
25
10
F
10
ND
8 - 10
25
26
R
11 - 15
3.1 - 7.8
ND
ND
31
R
11 - 15
3.1 - 7.8
ND
ND
4700
6000
20
R
S
F
13 - 16.8
13 - 16.8
10
Sat
Sat
ND
ND
ND
8 - 10
ND
ND
25
>20
F
10
ND
8 - 10
25
4800
5800
1.5
R
S
ND
13 - 16.8
13 - 16.8
5-6
Sat
Sat
ND
ND
ND
ND
ND
ND
ND
1586
R
3176
R
3176
R
ND
6.2
Kincheloe et al.
1979
6.2
Kincheloe et al.
1979
6.8 - 7.0 Grabda et al.
1974
6.8 - 7.0 Grabda et al.
1974
ND
Westin 1974
ND
Westin 1974
6.2
Kincheloe et al.
1979
6.2
Kincheloe et al.
1979
ND
Westin 1974
Westin 1974
ND
Lahti et al. 1985
25 ± 1.0 7.9 - 8.3 156 - 172 140 - 170 7.9 - 8.3 Scott and
Crunkilton 2000
25 ± 1.0 7.9 - 8.3 156 - 172 140 - 170 7.9 - 8.3 Scott and
Crunkilton 2000
25 ± 1.0 7.9 - 8.3 156 - 172 140 - 170 7.9 - 8.3 Scott and
Crunkilton 2000
Science-based Solutions No. 1-6
Ranking** Notes
A
b,c
A
i
A
c
A
i
A
i
A
i
A
i
A
c,g
A
i,j
A
i
A
d,g,i
A
d,g,h,i
2
2
A
i
A
i,j
2
2
A
c,g
1
1
1
108
Organism
Life Stage
Larvae
Na
+
Larvae
Na
+
Na
+
Larvae
Poecilia reticulatus
(guppy)
Salmo clarki
(cutthroat trout)
AMPHIBIANS
Ambystoma gracile
(northwestern salamander)
Cation
Endpoint
7-d NOEC
(spawning
success)
96-h LC50
Effect
Test
concentration Type*
- -1
(mg NO3 ·L )
3176
R
5941
R
7-d LOEC
(mortality)
96-h LC50
6353
R
847
S
Temp
(°C)
Hardness Alkalinity
DO
-1
-1
-1
(mg·L )
(mg·L ) (mg·L )
pH
1
25 ± 1.0 7.9 - 8.3 156 - 172 140 - 170 7.9 - 8.3 Scott and
Crunkilton 2000
25 ± 1.0 7.9 - 8.3 156 - 172 140 - 170 7.9 - 8.3 Scott and
Crunkilton 2000
77 F
>6.0
117 - 126 25.2 - 43.8 7.4 - 7.7 Rubin and
Elmaraghy
1977
77 F
>6.0
117 - 126 25.2 - 43.8 7.4 - 7.7 Rubin and
Elmaraghy
1977
77 F
>6.0
117 - 126 25.2 - 43.8 7.4 - 7.7 Rubin and
Elmaraghy
1977
77 F
>6.0
117 - 126 25.2 - 43.8 7.4 - 7.7 Rubin and
Elmaraghy
1977
13
ND
6-9
39
7.6
Kincheloe et al.
1979
13
ND
6-9
39
7.6
Kincheloe et al.
1979
1
K
+
Fry
K
+
72-h LC50
882
S
Fry
K
+
48-h LC50
969
S
Fry
K
+
24-h LC50
1181
S
Egg
Na
20
F
Fry
Na
> 30-d LOEC
(survivorship)
> 30-d LOEC
(survivorship)
30
F
Larvae
K
+
55
R
15
ND
ND
ND
7?
K
+
104
R
15
ND
ND
ND
7?
> 41
S
5 - 10
ND
ND
ND
6.5
> 41
S
5 - 10
ND
ND
ND
> 41
S
5 - 10
ND
ND
>111
R
15
ND
ND
Larvae
Ambystoma jeffersonianum
(Jefferson salamander)
Egg
Na
+
Ambystoma maculatum
(spotted salamander)
Egg
Na
+
Bufo americanus
(American toad)
Egg
Na
+
Bufo boreas
(western toad)
Larvae
K
+
25-d LOEC
(hatching
success,
deformities)
44-d LOEC
(hatching
success,
deformities)
23-d LOEC
(hatching
success,
deformities)
15-d LOEC
(mortality)
Ranking** Notes
25 ± 1.0 7.9 - 8.3 156 - 172 140 - 170 7.9 - 8.3 Scott and
Crunkilton 2000
Fry
15-d LOEC
(mortality)
15-d LC50
Reference
Science-based Solutions No. 1-6
1
1
e,h
1
e,h
1
e,h
1
e,h
A
i,j
A
i
Marco et al.
1999
Marco et al.
1999
Laposata and
Dunson 1998
1
h
1
h
2
l
6.5
Laposata and
Dunson 1998
2
l
ND
6.5
Laposata and
Dunson 1998
2
l
ND
7?
Marco et al.
1999
1
h
109
Organism
Bufo bufo
(common toad)
Litoria caerulea
(tree frog)
Pseudacris regilla
(formerly Hyla regilla)
(Pacific treefrog)
Rana aurora
(red-legged frog)
Life Stage
Cation
Na
+
Tadpole
Na
+
Tadpole
Na
+
Tadpole
Na
+
Larvae
K
Tadpole
+
Endpoint
16-d LOEC
(mortality)
16-d LOEC
(length)
16-d LOEC
(length)
16-d LOEC
(mortality)
15-d LOEC
(mortality)
Effect
Test
concentration Type*
- -1
(mg NO3 ·L )
40
R
Temp
(°C)
Hardness Alkalinity
DO
-1
-1
-1
(mg·L )
(mg·L ) (mg·L )
19 - 24
ND
ND
ND
40
R
19 - 24
ND
ND
ND
40
R
22.5 - 26
ND
ND
ND
40
R
22.5 - 26
ND
ND
ND
>111
R
15
ND
ND
ND
pH
Reference
5.6 - 7.5 Baker and
Waights 1993
5.6 - 7.5 Baker and
Waights 1993
5.6-7.6 Baker and
Waights 1994
5.6-7.6 Baker and
Waights 1994
7?
Marco et al.
1999
Ranking** Notes
A
e,f
A
e,f
A
e,f,h
A
e,f,h
1
h
Tadpole
Na
+
10-d LOEC
(weight)
133
R
Embryo
Na
+
251
R
Embryo
Na
+
492
R
22 ± 0.1 7.6 ± 0.1 75.0 ± 4.6 54.0 ± 1.2
Schuytema and
Nebeker 1999a
1
Tadpole
Na
+
560
R
Tadpole
Na
1148
R
Tadpole
Na
+
1179
R
Embryo
Na
+
10-d LC50
2561
R
Embryo
Na
+
96-h LC50
2849
R
Tadpole
Na
+
96-h LC50
7752
R
Embryo
Na
+
129
R
Embryo
Na
+
517
R
Embryo
Na
+
1041
R
Embryo
Na
+
16-d LOEC
(length)
16-d NOEC
(weight)
16-d LOEC
(weight)
16-d LC50
2819
R
Embryo
Na
+
4067
R
22 ± 0.1 7.2 ± 0.1 58.4 ± 9.5 52.0 ± 7.0 7.0 - 7.6 Schuytema and
Nebeker 1999c
22 ± 0.1 7.2 ± 0.1 58.4 ± 9.5 52.0 ± 7.0 7.0 - 7.6 Schuytema and
Nebeker 1999c
22 ± 1.0 7.2 ± 0.1 58.4 ± 9.5 52.0 ± 7.0 7.0 - 7.6 Schuytema and
Nebeker 1999c
22 ± 0.1 7.6 ± 0.1 75.0 ± 4.6 54.0 ± 1.2
6.7
Schuytema and
Nebeker 1999a
22 ± 0.1 7.6 ± 0.1 75.0 ± 4.6 54.0 ± 1.2
6.7
Schuytema and
Nebeker 1999a
22 ± 1.0 7.2 ± 0.1 58.4 ± 9.5 52.0 ± 7.0 7.0 - 7.6 Schuytema and
Nebeker 1999c
15 ± 1 8.7 ± 0.2 25.5 ± 1.7 24.2 ± 1.6
6.8
Schuytema and
Nebeker 1999b
15 ± 1 8.7 ± 0.2 25.5 ± 1.7 24.2 ± 1.6
6.8
Schuytema and
Nebeker 1999b
15 ± 1 8.7 ± 0.2 25.5 ± 1.7 24.2 ± 1.6
6.8
Schuytema and
Nebeker 1999b
15 ± 1 8.7 ± 0.2 25.5 ± 1.7 24.2 ± 1.6
6.8
Schuytema and
Nebeker 1999b
15 ± 1 8.7 ± 0.2 25.5 ± 1.7 24.2 ± 1.6
6.8
Schuytema and
Nebeker 1999b
1
+
10-d NOEC
(weight and
length)
10-d LOEC
(weight and
length)
10-d NOEC
(length)
10-d LOEC
(length)
10-d LC50
16-d EC100
(mortality)
22 ± 0.1 7.2 ± 0.1 58.4 ± 9.5 52.0 ± 7.0 7.0 - 7.6 Schuytema
and Nebeker
1999c
22 ± 0.1 7.6 ± 0.1 75.0 ± 4.6 54.0 ± 1.2
6.7
Schuytema and
Nebeker 1999a
Science-based Solutions No. 1-6
6.7
1
1
1
1
1
1
1
1
1
1
1
1
110
Organism
Rana cascadae
(Cascades frog)
Rana pipiens
(northern leopard frog)
Rana pretiosa
(Oregon spotted frog)
Life Stage
Cation
Na
+
Larvae
Na
+
Larvae
Na
+
Larvae
K
+
Larvae
K
+
Larvae
Rana sylvatica
(wood frog)
Egg
Na
+
Rana temporaria
(European common frog)
Larvae
Na
+
Larvae
Na
+
Embryo
Na
+
Embryo
Na
+
Embryo
Na
+
Tadpole
Na
+
Embryo
Na
+
Embryo
Na
+
Tadpole
Na
+
Embryo
Na
+
Tadpole
Na
+
Embryo
Na
+
Tadpole
Na
+
Embryo
Na
+
Xenopus laevis
(African tree frog)
Endpoint
21-d LOEC
(mortality)
21-d LOEC
(activity)
56-d LOEC
(length)
15-d LOEC
(mortality)
15-d LC50
(mortality)
23-d LOEC
(hatching
success,
deformities)
35 to 48-d
(growth and
maturation)
72-h LOEC
(mortality)
120-h NOEC
(weight)
120-h LOEC
(weight)
120-h NOEC
(length)
10-d NOEC
(weight)
120-h LOEC
(length)
120-h NOEC
(deformities)
10-d LOEC
(weight)
120-h LOEC
(deformities)
10-d NOEC
(length)
120-h LC50
10-d LOEC
(length)
120-h EC50
(deformities)
Effect
Test
concentration Type*
- -1
(mg NO3 ·L )
> 20
R
Temp
(°C)
Hardness Alkalinity
DO
-1
-1
-1
(mg·L )
(mg·L ) (mg·L )
pH
Reference
Hatch and
Blaustein 2000
Hatch and
Blaustein 2000
Allran and
Karasov 2000
Marco et al.
1999
Marco et al.
1999
Laposata and
Dunson 1998
12 - 17
ND
32 - 48
15 - 20
5 or 7
> 20
R
12 - 17
ND
32 - 48
15 - 20
5 or 7
133
R
22
11.5
324
ND
8
55
R
15
ND
ND
ND
7?
73
R
15
ND
ND
ND
7?
> 41
S
5 - 10
ND
ND
ND
6.5
22
R
17.4
ND
ND
ND
> 4425
R
ND
ND
ND
ND
110
R
22 ± 0.1 7.6 ± 0.1 36.2 ± 6.5 34.5 ± 4.1
7
251
R
22 ± 0.1 7.6 ± 0.1 36.2 ± 6.5 34.5 ± 4.1
7
251
R
22 ± 0.1 7.6 ± 0.1 36.2 ± 6.5 34.5 ± 4.1
7
291
R
22 ± 0.1 7.2 ± 0.1 20.6 ± 0.2 26.0 ± 0.9 6.7 - 7.6
492
R
22 ± 0.1 7.6 ± 0.1 36.2 ± 6.5 34.5 ± 4.1
7
492
R
22 ± 0.1 7.6 ± 0.1 36.2 ± 6.5 34.5 ± 4.1
7
560
R
22 ± 0.1 7.2 ± 0.1 20.6 ± 0.2 26.0 ± 0.9 6.7 - 7.6
1021
R
22 ± 0.1 7.6 ± 0.1 36.2 ± 6.5 34.5 ± 4.1
1148
R
22 ± 0.1 7.2 ± 0.1 20.6 ± 0.2 26.0 ± 0.9 6.7 - 7.6
1942
R
22 ± 0.1 7.6 ± 0.1 36.2 ± 6.5 34.5 ± 4.1
2190
R
22 ± 0.1 7.2 ± 0.1 20.6 ± 0.2 26.0 ± 0.9 6.7 - 7.6
2311
R
22 ± 0.1 7.6 ± 0.1 36.2 ± 6.5 34.5 ± 4.1
Science-based Solutions No. 1-6
7.7-7.9 Johansson et
al. 2001
7.5
7
7
7
Johansson et
al. 2001
Schuytema and
Nebeker 1999a
Schuytema and
Nebeker 1999a
Schuytema and
Nebeker 1999a
Schuytema and
Nebeker 1999c
Schuytema and
Nebeker 1999a
Schuytema and
Nebeker 1999a
Schuytema and
Nebeker 1999c
Schuytema and
Nebeker 1999a
Schuytema and
Nebeker 1999c
Schuytema and
Nebeker 1999a
Schuytema and
Nebeker 1999c
Schuytema and
Nebeker 1999a
Ranking** Notes
A
l,m
A
l,m
1
1
h
1
h
2
l
A
e,g
A
e,l
1
1
1
1
1
1
1
1
1
1
1
1
111
Organism
Life Stage
Tadpole
Tadpole
Cation
Endpoint
Na
+
10-d LC50
Na
+
96-h LC50
Effect
Test
concentration Type*
- -1
(mg NO3 ·L )
5476
R
7335
R
Temp
(°C)
Hardness Alkalinity
DO
-1
-1
-1
(mg·L )
(mg·L ) (mg·L )
pH
Reference
22 ± 1.0 7.2 ± 0.1 20.6 ± 0.2 26.0 ± 0.9 6.7 - 7.6 Schuytema and
Nebeker 1999c
22 ± 1.0 7.2 ± 0.1 20.6 ± 0.2 26.0 ± 0.9 6.7 - 7.6 Schuytema and
Nebeker 1999c
Ranking** Notes
1
1
Notes: ND = no data provided; Sat = saturation (O2)
* Test Types: R = renewal, S = static, F = flow-through
** Ranking Scheme: 1 = primary source, 2 = secondary source, A = ancillary source
a
LC0.01 extrapolated from Camargo and Ward (1992) LC50 data, therefore not used in guideline development
b
tests run with filtered local lake water
c
insufficient test details / water quality information provided
d
lack of statistical support
e
non-resident, or tropical species
f
distilled water used as test medium
g
lack of clear dose-response relationship
h
potassium salts not suitable for guideline derivation
i
inadequate test design or conditions
j
control mortality > 10%
k
organisms only exposed to one test concentration
l
lowest observable effect level beyond nitrate concentration range tested
m
>10% change in nitrate concentration in test containers
n
the ecological significance of this endoint is uncertain
Science-based Solutions No. 1-6
112
APPENDIX B. SUMMARY OF MARINE TOXICITY STUDIES.
Organism
INVERTEBRATES
Cherax quadricarinatus
(Australian crayfish)
Crassostrea virginica
(oyster)
Dorvillea articulata
(polychaete)
Haliotis tuberculata
(abalone)
Mercinaria mercinaria
(hard clam)
Life
Stage
Cation
Juvenile
Na
+
Juvenile
Na
+
Juvenile
Na
+
Adult
Na
+
Juvenile
Na
+
Adult
Na
+
ND
K
+
ND
Na
+
Juvenile
Na
+
Adult
Na
+
Juvenile
Na
+
Adult
Na
+
Endpoint
Effect
Test
concentration Type*
- -1
(mg NO3 ·L )
Temp
(°C)
Salinity Alkalinity
DO
-1
-1
(‰)
(mg·L )
(mg·L )
120-h LOEC
(mortality)
120-h LOEC
(respiration)
20-h LOEC
(feeding)
20-h LOEC
(feeding)
96-h LC50
> 4430
R
28.0
Sat
ND
70.5 ± 5
> 4430
R
28.0
Sat
ND
70.5 ± 5
9921
S
20 ± 2
7.0 - 8.2
35
ND
9921
S
20 ± 2
7.0 - 8.2
35
ND
11533
S
20 ± 2
7.0 - 8.2
35
ND
96-h LC50
16803
S
20 ± 2
7.0 - 8.2
35
ND
28-d LC50
880
S
22 - 25
5.9
34.7
ND
15-d LOEC
(growth)
20-h LOEC
(feeding)
20-h LOEC
(feeding)
96-h LC50
1108
R
Sat
34 ± 1
200 ± 25
2480
S
18.5 ±
0.5
20 ± 2
7.0 - 8.2
35
ND
9921
S
20 ± 2
7.0 - 8.2
35
ND
>19840
S
20 ± 2
7.0 - 8.2
35
ND
pH
7.5 ± 0.2 Meade and
Watts 1995
7.5 ± 0.2 Meade and
Watts 1995
7.7 - 8.2 Epifanio and
Srna 1975
7.7 - 8.2 Epifanio and
Srna 1975
7.7 - 8.2 Epifanio and
Srna 1975
7.7 - 8.2 Epifanio and
Srna 1975
ND
Reish 1970
8.1 ± 0.5 Basuyaux and
Mathieu 1999
7.7 - 8.2 Epifanio and
Srna 1975
7.7 - 8.2 Epifanio and
Srna 1975
7.7 - 8.2 Epifanio and
Srna 1975
7.7 - 8.2 Epifanio and
Srna 1975
ND
Reish 1970
96-h LC50
>19840
S
20 ± 2
7.0 - 8.2
35
ND
+
28-d LC50
496
S
22 - 25
5.9
34.7
ND
+
28-d LC50
15-d LOEC
(growth /
feeding)
40-h LOEC
(mortality)
40-h LOEC
(cellular
changes)
40-h LOEC
(mortality)
329
1108
S
R
22 - 25
18.5 ±
0.5
5.9
Sat
34.7
34 ± 1
ND
200 ± 25
1
S
28.0
Sat
NR
ND
8.2
1
S
28.0
Sat
NR
ND
8.2
1
S
28.0
Sat
NR
ND
8.2
Neanthes arenaceodentata
(polychaete)
Nereis grubei (polychaete)
Paracentrotus lividus
(purple sea urchin)
ND
K
ND
ND
K
+
Na
Penaeus monodon
(prawn)
Larvae
Na
+
Larvae
Na
+
Larvae
K
+
Science-based Solutions No. 1-6
Reference
ND
Reish 1970
8.1 ± 0.5 Basuyaux and
Mathieu 1999
Ranking**
Notes
2
a,f
2
a,f
A
b,c
A
b,c
A
b,c
A
b,c
2
1
A
b,c
A
b,c
A
b,c
A
b,c
2
2
1
Muir et al.
1991
Muir et al.
1991
1
d
1
d
Muir et al.
1991
1
d
113
Organism
Penaeus paulensis
(prawn)
Penaeus spp.
(prawn)
Porites compressa
(coral)
FISH
Amphiprion ocellaris
(anemonefish)
Centropristis striata
(Gulf black sea bass)
Diplodus sargus
(white seabream)
Gaidropsarus capensis
(cape rockling)
Heteromycteris capensis
(cape sole)
Lithognathus mormyrus
(striped seabream)
Monacanthus hispidus
(planehead filefish)
Oncorhynchus mykiss
(rainbow trout)
Oncorhynchus tshawytscha
(chinook salmon)
Life
Stage
Cation
+
Larvae
K
Early
Instar
Adult
Na
+
Na
+
Early
Instar
Nubbin
Na
+
Larvae
Na
+
ND
Na
+
Larvae
Na
+
Larvae
Na
+
Larvae
Na
+
Larvae
Na
+
Larvae
Na
+
Larvae
Na
+
Larvae
Na
+
Larvae
Na
+
ND
Na
fingerling
Na
fingerling
fingerling
fingerling
Na
K
+
Endpoint
40-h LOEC
(cellular
changes)
21-d NOEC
(growth)
96-h LC50
Effect
Test
concentration Type*
- -1
(mg NO3 ·L )
1
S
Temp
(°C)
Salinity Alkalinity
DO
-1
-1
(‰)
(mg·L )
(mg·L )
pH
Reference
Sat
NR
ND
8.2
Muir et al.
1991
1
d
ND
Wickins 1976
A
d
2
d
A
d
A
d,e
A
b,c,e
F
28
ND
30 - 34
ND
9621
R
Sat
32 to 41
NR
48-h LC50
15062
S
27.0 ±
0.2
26 - 28
ND
30 - 34
ND
7.7 ± 0.2 Cavalli et al.
1996
ND
Wickins 1976
35-d LOEC
(growth)
> 0.35
F
ND
Sat
ND
1.96
-1
meq·L
7.1 - 8.0 Marubini and
Atkinson 1999
72-d LOEC
(growth,
mortality)
96-h LC50
443
S
ND
ND
NR
ND
ND
10 632
R
20 - 24
ND
32 ± 2
ND
ND
3455
S
15.0
ND
15 771
S
15.0
ND
4582
S
15.0
ND
> 17 720
S
15.0
ND
3145.3
S
15.0
ND
22 372
S
15.0
ND
107.5 122.5
107.5 122.5
107.5 122.5
107.5 122.5
107.5 122.5
107.5 122.5
107.5 122.5
107.5 122.5
ND
24-h EC50
(feeding)
24-h LC50
2658
S
15.0
ND
15 284
S
15.0
ND
+
96-h LC50
2538
R
20 - 24
ND
34.4 35.7
34.4 35.7
34.4 35.7
34.4 35.7
34.4 35.7
34.4 35.7
34.4 35.7
34.4 35.7
32 ± 2
+
7-d LC50
4000
S
13 - 14
>7
15
ND
Na
+
Na
+
96-h LC50
7-d LC50
4650
4000
S
S
13 - 14
13 - 14
>7
>7
15
15
+
96-h LC50
4400
S
13 - 14
>7
15
24-h EC50
(feeding)
24-h LC50
24-h EC50
(feeding)
24-h LC50
Notes
28.0
886
24-h EC50
(feeding)
24-h LC50
Ranking**
Science-based Solutions No. 1-6
Frakes and
Hoff Jr. 1982
Pierce et al.
1993
7.8 - 7.9 Brownell 1980
2
7.8 - 7.9 Brownell 1980
A
7.8 - 7.9 Brownell 1980
2
7.8 - 7.9 Brownell 1980
A
7.8 - 7.9 Brownell 1980
2
7.8 - 7.9 Brownell 1980
A
7.8 - 7.9 Brownell 1980
2
7.8 - 7.9 Brownell 1980
A
ND
1
ND
Pierce et al.
1993
Westin 1974
2
ND
ND
ND
ND
Westin 1974
Westin 1974
2
2
ND
ND
Westin 1974
2
f
f
f
f
1
114
Organism
Life
Stage
Cation
+
Endpoint
Effect
Test
concentration Type*
- -1
(mg NO3 ·L )
> 13 290
R
Temp
(°C)
Pomacentrus leucostictus
ND
Na
96-h LC50
20 - 24
(beaugregory)
+
Raja eglanteria
ND
Na
96-h LC50
> 4253
R
20 - 24
(clearnose skate)
+
96-h LC50
4430
R
20 - 24
Trachinotus carolinus
ND
Na
(Florida pompano)
Notes: ND = no data provided; NR = variable measured but not reported; Sat = saturation (O2)
* Test Types: R = renewal, S = static, F = flow-through
** Ranking Scheme: 1 = primary source, 2 = secondary source, A = ancillary source
a
lowest observable effect level beyond nitrate concentration range tested
b
insufficient test details / water quality information provided
c
lack of statistical support
d
tropical species
e
lack of clear dose-response relationship
f
toxicity could be due to increased salinity levels
Salinity Alkalinity
DO
-1
-1
(‰)
(mg·L )
(mg·L )
pH
ND
32 ± 2
ND
ND
ND
32 ± 2
ND
ND
ND
32 ± 2
ND
ND
Science-based Solutions No. 1-6
Reference
Pierce et al.
1993
Pierce et al.
1993
Pierce et al.
1993
Ranking**
Notes
1
a
1
a
1
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