Review and Identification of Research Needs to Address Key Issues Related to
Reactive Nitrogen (RN) Deposition and Eutrophication in a Canadian Context
(Final Report)
Prepared for:
Acid Rain Task Group
Canadian Council of Ministers of the Environment
Prepared by:
Judi Krzyzanowski
Krzyzanowski Consulting
Stirling-Rawdon ON
June 5, 2010
PN 1450
This report was prepared by Krzyzanowski Consulting, under contract to the
Canadian Council of Ministers of the Environment (CCME) and is a working
paper only. It contains information which has been prepared for, but not
approved by, CCME. CCME is not responsible for the accuracy of the
information contained herein and does not warrant, or necessarily share or
affirm, in any way, any opinions expressed therein.
© Canadian Council of Ministers of the Environment 2010
Executive Summary
This report consolidates what is known about the deposition and effects of reactive nitrogen (RN)
in a Canadian context and offers recommendations for representative monitoring and
management in Canada. It is based on a review of existing knowledge conducted in 2009 and
expands, synthesizes and updates the original review into a source of information for the public
and policy-makers on the threat and assessment of reactive nitrogen effects in Canada. While
molecular nitrogen (N 2 ) comprises about 78% of Earth’s atmosphere, anthropogenic emissions of
reactive (biologically available) nitrogen have been increasing in Canada in recent decades. The
emission of these compounds to air, soil and water arise primarily from combustion (NO x ) and
agriculture (NH x ); together can lead to ecosystem eutrophication or acidification in Canadian
ecosystems. The focus of this report is atmospheric nitrogen and the eutrophication that can occur
from its deposition. However due to links between non-atmospheric nitrogen, other nutrients such
as phosphorus, and processes such as acidification, the breadth of this report is expanded to
include a (lesser) discussion of these elements.
It general it was found that:
(1) Compared to parts of Europe and the United States, atmospheric RN deposition is low in
most of Canada, but is highest in the Great Lakes basin, and in Alberta’s Calgary corridor
and oil sands regions. Other oil and gas producing areas are also experiencing increasing
emissions of NO x - such as northeastern British Columbia and parts of Saskatchewan and
Manitoba. In British Columbia’s Georgia Basin increases in agriculturally produced
forms of reduced nitrogen (NH x ), and transport sector produced oxidised forms (NO x ),
are also of interest.
(2) Ecosystem sensitivity to acid deposition is determined by the buffering capacities of soil
and water, whereas sensitivity to eutrophication depends on the nutrient status and
nutrient requirements of ecosystems and their species.
(3) Nitrogen eutrophication occurs prior to nitrogen acidification, however the occurrence of
these two states also depends largely on nutrient status and sulphate availability,
respectively.
(4) Critical loads and their resulting exceedances are among the most useful tools for the
assessment and management of ecosystem eutrophication and acidification. Critical loads
are used to set reduction targets for emissions so that undesirable ecosystem effects do
not occur, and have been successful in emissions reduction programs throughout Europe
and eastern North America.
(5) Data show that increasing RN emissions are increasing the threat of eutrophication and
acidification to ecosystems in western Canada.
(6) Current deposition and air quality monitoring stations should be expanded to represent a
greater portion of Canada’s varied landscapes and emission sources.
(7) Existing water quality monitoring can be used to aid in eutrophication sensitivity
analysis.
(8) Forest and ecosystem monitoring should accompany atmospheric monitoring stations so
as to develop Critical Loads specific to Canadian ecosystems.
(9) The use of integrated or cumulative effects management principles can greatly aid in
approaching the inherently multidisciplinary and complex nature of RN in Canadian
ecosystems.
(10) Emission zoning is recommended as a tool for the management of atmospheric RN loads
1
While the sensitivity of Canadian ecosystems to eutrophication is still largely unknown, other
than what has been inferred from European studies, use of the appropriate multidisciplinary
and integrated approaches can be used to monitor and manage ecosystem effects across
Canada.
Acknowledgements
This work was carried out with funding from CCME who are acknowledged with thanks. In
addition to the helpful comments and survey responses of various anonymous researchers, this
report was aided with data, maps and helpful discussions from the following individuals: Julian
Aherne (Trent University), Robert Vet (Environment Canada), Cèline Audette (Environment
Canada), Sagar Krupa (University of Minnesota), Daniel McCarthy (Brock University), Jennifer
Grixti (Ontario Environment), Duncan Boyd (Ontario Environment), Darrell Taylor (Nova Scotia
Environment), Kyla Brake (Newfoundland Environment), Craig Nelda (New Brunswick
Environment), Nicole Armstrong (Manitoba Water Stewardship), Shanti Berryman (Stantec),
Mark Fenn (US Forest Service), Martha Guy (Environment Canada), Tom Clair (Environment
Canada), Bill Sukloff (Environment Canada), and Witold Fraczek (ESRI) – all of whom are
gratefully acknowledged. Additional thanks goes to the Acid Rain Task Group (ARTG) and
ARTG’s Nitrogen Sub-group for their valuable commentary on this manuscript.
2
Table of Contents
Executive Summary ........................................................................................................................ 1
Acknowledgements ......................................................................................................................... 2
Table of Contents ............................................................................................................................ 3
List of Tables................................................................................................................................... 4
List of Figures ................................................................................................................................. 5
Glossary of Abbreviations and Terms............................................................................................. 6
1.0 Introduction ............................................................................................................................. 11
2.0 Atmospheric RN Deposition ................................................................................................... 12
2.1 RN Chemistry...................................................................................................................... 13
2.2 Atmospheric RN Deposition ............................................................................................... 14
2.3 Sources of Atmospheric RN................................................................................................ 16
2.4 Sources of Non-Atmospheric RN ....................................................................................... 19
3.0 Effects of RN Deposition ........................................................................................................ 20
3.1 Eutrophication ..................................................................................................................... 20
3.1.2 Chemical effects........................................................................................................... 24
3.1.4 Whole ecosystem effects.............................................................................................. 25
3.1.5 Societal effects ............................................................................................................. 28
3.2 Acidification........................................................................................................................ 28
3.2.1 Chemical effects........................................................................................................... 29
3.2.2 Biological effects.......................................................................................................... 30
3.3 RN Output and Sinks........................................................................................................... 30
3.4 Integrating Eutrophication, Acidification Effects ............................................................... 32
4.0 Tools for Monitoring and Management in Canada ................................................................. 35
4.1 Tools for Quantifying Deposition ....................................................................................... 36
4.1.1 Measured Deposition.................................................................................................... 36
4.1.2 Modelled Deposition .................................................................................................... 42
4.2 Nitrogen Loading ................................................................................................................ 46
4.3 Tools for Eutrophication ..................................................................................................... 47
4.3.1 Critical Loads of Eutrophication (Nutrient Nitrogen).................................................. 48
4.3.2 Alternative tools, models and indices .......................................................................... 53
4.4 Tools for Acidification ........................................................................................................ 56
4.4.1 Critical Loads of Acidification..................................................................................... 56
5.0 Directions for RN Effects Assessment and Management in Canada ...................................... 57
5.1 Measured and Modelled Deposition ................................................................................... 57
5.2 Physical, Chemical and Biological Processes ..................................................................... 60
5.3 A Comprehensive Monitoring Program.............................................................................. 61
5.4 Eutrophication or Acidification........................................................................................... 65
5.5 Eutrophication Critical Loads ............................................................................................. 67
5.6 Management Strategies ....................................................................................................... 68
6.0 Conclusion............................................................................................................................... 71
7.0 References ............................................................................................................................... 73
3
List of Tables
Table 1. Examples of different RN compounds by state and depositional form. Developed using
information from Chambers et al. (2001); Russow and Böhme (2005); Acker et al. (2005);
Moran et al. (2009). This is not an inclusive list. Our knowledge and measurement of RN chemistry
is constantly evolving and RN may still be found in Canada’s atmosphere in different phases and
forms. ..................................................................................................................................... 15
Table 2. Relationships among the units for atmospheric RN deposition particularly in terms of
assessing ecosystem sensitivity to eutrophication and acidification (adapted from CCME
2005)...................................................................................................................................... 16
Table 3. Total year 2007 atmospheric emissions of RN included in the NPRI by
province/territory (in alphabetical order) and RN chemical species. All values are in t/y. .. 18
Table 4. National 2008 NPRI facility reported RN (NO x and NH 3 ) releases to land, water and air
(tonnes) in Canada from Environment Canada (2010). ........................................................ 46
Table 5. Empirical critical loads of nutrient nitrogen for protection of overall ecosystem health
and productivity in selected ecosystem. Adapted from Bobbink et al. (2002). The high end
of the range should be used if phosphorus is limited and the lower range if phosphorus is
available. ............................................................................................................................... 49
Table 6. Suggestions to use lower, middle or upper part of critical load range (Table 5) based on
ecosystem characteristics. Adapted from Achermann and Bobbink (2003). ........................ 49
Table 7. Indicators for the ecosystem sensitivity to nitrogen deposition in the Mountain Hemlock
Zone of the Georgia Basin (adapted from Zhong, 2004b). ................................................... 52
Table 8. Summary of N and P monitoring at sites reporting WQI through CESI. A = all sites
measure the parameter, S = some sites measure a parameter................................................ 62
4
List of Figures
Figure 1. Artistic rendition showing sources of reactive nitrogen focusing on atmospheric
deposition pathways. Atmospheric sources are shown with up arrows, sources of deposition
(after chemical transformation) are shown with down-arrows, sources that can be both from
the surface and though atmospheric volatisation (such as manure, fertiliser and municipal
sources of NH 3 ) are underlined, and the soil processes of nitrification, nitrogen fixation and
mineralization are given by 1), 2) and 3) respectively. The production of ozone via
photolysis and the formation of aqueous ammonium are also shown. This does not represent
an exhaustive representation of RN sources or chemistry. ................................................... 13
Figure 2. Contributions to NPRI-included federal emissions of NO x and NH 3 by source grouping
(a) and province or territory (b)............................................................................................. 17
Figure 3. Documented sites of nutrient enrichment in 1998 (from Chambers et al. 2001). Green
areas represent impacts from agriculture, red from industry, yellow from municipal effluent,
grey from other and natural factors, and the light green tinge highlights agricultural areas. 21
Figure 4. Conceptualisation of the effects of increased reactive nitrogen deposition on forest
ecosystems adapted from Gundersen (1999). The green forest growth curve represents
positive growth, rather than impact, in the y-direction. ........................................................ 33
Figure 5. Plot of N and S emissions and reductions required to meet priorities. Source: UNECECLRTAP (2004).................................................................................................................... 34
Figure 6. CAPMoN stations and what they measured in 2009. “Air” refers to filter-pack
measurements of daily gaseous HNO 3 ; and daily particulate NH 4 + and NO 3 -. Source: Robert
Vet Personal Communication February 2010....................................................................... 37
Figure 7. Map of precipitation monitoring stations (circles) in Canada past (white) and present
(black) – as of June 2006. CAPMoN sites, of particular relevance to RN deposition, are
represented by red outlines. Source: Canadian National Atmospheric Chemistry
Precipitation Database (2006). .............................................................................................. 38
Figure 8. Location of NAPS Network sites that measure concentrations of NO 2 . Source:
National Air Pollution Surveillance Monitoring Program, Environment Canada, Cèline
Audette, Personal Communication March 2010................................................................... 41
Figure 9. Changes in the spatial distribution of measured wet nitrate deposition in a) 1990, b)
1995, c) 2000, d) 2005. Source: International Joint Commission (2008), with data from
NAtChem (Canadian National Atmospheric Chemistry Precipitation Database (2007)) using
measurements from CAPMoN and NADP (National Atmospheric Deposition Program of
the US)................................................................................................................................... 42
Figure 10. AURAMS simulated total N deposition across (western) Canada using year 2002
emissions. Source: Mike Moran............................................................................................ 44
Figure 11. Critical loads for nutrient N and their exceedances using approaches in Bobbink et al
(2002). Source: Aherne (2007).............................................................................................. 51
Figure 12. Residual Soil Nitrogen under 2001 agricultural management practices. From: Drury et
al. (2005). The classes are from green to red: 0-9.9, 10-19.9, 20-29.9, 30-39.9, >40 kgN/ha.
............................................................................................................................................... 54
5
Glossary of Abbreviations and Terms
AERMOD – American Meteorological Society (AMS) / United States Environmental Protection
Agency
(EPA) Regulatory Model Improvement Committee’s Regulatory Model
Anion – a molecule (ion) with a negative charge i.e. more electrons than there are protons;
examples include nitrate (NO 3 -) and chloride (Cl-)
ANC – Acid Neutralising Capacity; the difference between base cation and strong-acid anion
concentrations
AURAMS – A Unified Regional Air-quality Modelling System developed by Environment
Canada
BC – base cations; include Ca2+, Na2+, K+ and Mg+
BCPCSN – British Columbia Precipitation Chemistry Sampling Network
Biosolids – treated sewage
Bryophytes – non-vascular plants comprised of mosses and liverworts
Buffering Capacity – the ability of a system to neutralise incoming acidity (H+); often measured
in water as acid Neutralising Capacity
CALPUFF – California Puff Model
CAPMoN – The Canadian Air and Precipitation Monitoring Network
Cation – a molecule (ion) with a positive charge i.e. more protons than there are electrons;
examples include ammonium (NH 4 +) and calcium (Ca2+)
CCME – Canadian Council of Ministers of the Environment
CEA – Cumulative Effects Assessment
CEMA – Cumulative Effects Management Association in Wood Buffalo, Alberta
CESI – Canadian Environmental Sustainability Indicators
Chromatography – a broad range of methods used to separate and analyse mixtures for their
constituents/
Colorimetry - measurement of the wavelength of radiation (light, colour) emitted by a substance
so that its make-up can be described
6
CL – Critical Load(s)
CL(EX) – Critical Load Exceedance; the difference between atmospheric deposition and the CL
CMAQ – The Community Multiscale Air Quality Model developed by the US-EPA
Critical Load – A quantitative estimate of an exposure to one or more pollutants below which
significant harmful effects on specified sensitive elements of the environment do not occur
according to present knowledge (Nilsson and Grennfelt 1988)
DFO – Canadian Department of Fisheries and Oceans
Ecozone – a biogeographic region of Canada; principally a system of naming regions based on
their geological and biological distributions
EMAN – Ecological Monitoring and Assessment Network (Canada)
EMEP – European Monitoring and Evaluation Programme - a scientifically based and policy
driven programme under the Convention on Long-range Transboundary Air Pollution for
international co-operation to solve transboundary air pollution problems.
eq/ha/yr – hydronium ion (H+) equivalents per hectare per year
Equilibrium – a condition in a system in which opposing forces (or chemical compounds) are
balanced
Heathland – ecosystems dominated by dwarf shrub communities from the heath or heather
(Ericaceae) family
HNO 3 – nitric acid
HNO x – represents an acid of oxidised nitrogen; either nitric acid (HNO 3 ) or nitrous acid
(HNO 2 )
Hydronium Ion – H+, a single proton and neutron, or hydrogen that is missing an electron, the
ion associated with acids and acidification in soil or water
IER – Ion Exchange Resin
Ion Exchange Resin – IER; used in a column for collecting constituents of precipitation
Inorganic – used here in the chemical sense to mean a non-carbon based compound
Ion – any molecule with chemical charge i.e. that is not chemically neutral. Examples are nitrate
(NO 3 -) and ammonium (NH 4 +); see entries for anion and cation
7
IROWC-N – Indicator of the Risk of Water Contamination by Nitrogen
ISCT3 – Industrial Source Complex Short Term Model (3)
kgN/ha/yr – kilograms of nitrogen per hectare per year; a measure of total deposition where Nnitrogen represents the weight fraction of N in a molecule. For example NO 3 weighs
approximately 14 + (3 x 16) = 62g, but only 14 of that is N. So 62kg of NO 3 would be equal to
14kg of N and so on
Kriging – a geostatistical technique of interpolating a the value of a field at an unmeasured
location, through values measured at other nearby locations
Macrophyte – (large) aquatic plant whose form can be seen with the naked eye; unlike algae,
many of which are single-celled (microphytes)
MAGIC - Model of Acidification of Groundwaters In Catchments
Mass Balance – the sum of inputs minus outputs from a system; if inputs and outputs are equal
the system is said to be at “Steady State”
Mole – a measure used in chemistry to represent the weight of a compound in grams; 1 mole is
always equal to 6.022 x 1023 molecules of that compound (known as Avogadro’s number), for
example 1 mole of H+ weighs approximately 1 gram, 1 mole of N weighs approximately 14
grams and so on
N – nitrogen the element in the context of any of its molecular forms
N 2 – molecular nitrogen
NADP – National Atmospheric Deposition Program of the US
NAESI – National Agri-Environmental Standard Initiative
NAPS – National Air Pollution Surveillance Network
NatChem – National Atmospheric Chemistry (Database)
NEG/ECP – Conference of New England Governors and Eastern Canadian Premiers
NH x – reduced forms of reactive nitrogen; namely ionised or unionised ammonia
NH 3 – unionised ammonia
NH 4 + – ionised ammonia, also called ammonium
Nitrification – the conversion (oxidation) of NH 3 to NO 2 and NO 2 to NO 3 by nitrifying bacteria
in soil or water systems
8
Nitrophilous – nitrogen-loving species or plants that thrive in high nitrogen conditions
NLPMN – Newfoundland Environment Precipitation Monitoring Network
NO – nitrogen oxide
NO 2 – nitrogen dioxide
NO 3 - – nitrate ion
NO x – nitrogen oxides or oxidized forms of reactive nitrogen; namely NO, NO 2 or NO 3 in
neutral or ionic state
O 3 – ozone; in this case referring to the tropospheric kind i.e. that of the lower atmosphere
OH- – hydroxyl ion; a free radical responsible for dry oxidation of nitrogen in the atmosphere
Oligotrophic – soil and water systems that are nutrient-poor
Ombrotrophic – a wetland (bog, moor or fen) that is rainwater fed rather than being fed by
streams or groundwater
Organic – used here in the chemical sense to mean a carbon-based compound
P – phosphorus
PAN – peroxyacyl nitrate – the most common of a class of atmospheric pollutants known as
“peroxy nitrates” and peroxyacetyl nitrate is the most common form of PAN
Peroxy nitrates – a class of secondary pollutants of the general form RO 2 NO 2 where “R”
represents an organic compound
pH – the inverse logarithm of hydronium ion (H+) concentration; used as a measure of acidity in
a solution
Plankton – small single-celled organisms in aquatic environments; they include the
photosynthetic phytoplankton, and zooplankton that require food from external sources
PM – particulate matter
ppm – parts per million; a unit used to measure pollutant concentration
RN – reactive nitrogen; including NO x , NH y and many nitrogen containing organic molecules.
Scrubland – an ecosystem dominated by woody shrubs
9
The Strategy – The Canada-Wide Acid Rain Strategy for Post-2000
Throughfall – precipitation that falls to the surface after being intercepted by vegetation
Toxic Aluminium – inorganic aluminium in the form of free Al3+ ions; at higher soil or water pH
aluminium is bound, usually as Al(OH) x (aluminium hydroxides) where the subscript x denotes
variable molecular structures.
Trophic Level – akin to steps in the food chain but begins with primary producers i.e. plants, that
don’t eat anything
UNECE – United Nations Economic Commission for Europe
UNEP – United Nations Environmental Programme
US – United States
US-EPA – United States Environmental Protection Agency
VEC – valued ecosystem component(s)
VOC – volatile organic compound(s)
Volatilisation – conversion of a compound from solid or liquid into gaseous state
WQI – Water Quality Index
g/m3 – micrograms per cubic metre; a measure of air pollutant concentration
10
1.0 Introduction
The Canadian Council of Ministers of the Environment (CCME) is “comprised of the
environment ministers from the federal, provincial and territorial governments. These 14
ministers meet to discuss national environmental priorities and determine work to be carried out
under the auspices of CCME. The Council seeks to achieve positive environmental results,
focusing on issues that are national in scope and that require collective attention by a number of
governments” (CCME 2010). Under guidance of CCME, the Acid Rain Task Group (ARTG) is a
multi-stakeholder team that assists in implementation of The Canada-Wide Acid Rain Strategy
for Post-2000. The Strategy, signed in 1998 by all 26 Energy and Environment Ministers aims to
meet environmental thresholds for acid deposition across the country through various emission
reduction approaches. One of The Strategy’s long-term goals is assessing the role of atmospheric
nitrogen in acidification and acidification-related issues such as eutrophication.
Reactive nitrogen (RN) refers to nitrogen species in biologically active (or available) form.
Atmospheric RN includes nitrogen oxides (NO x ), ammonia (NH x ), and nitrous oxide (N 2 O). As
primary pollutants or as chemically transformed reaction products, these substances may enter
natural systems through atmospheric, aquatic or terrestrial pathways; and arrive at ecological
receptors in solid (particulate matter), liquid (dissolved), or gaseous states. The potential effects
of RN are obscured by its ability to act as both a growth promoter and growth inhibiter at
different levels of exposure. This report focuses on RN arriving at ecosystems through
atmospheric pathways. However, as will be shown throughout this report, the outcomes of
atmospheric RN are inextricably tied to aquatic and terrestrial sources and sinks.
Two major effects of RN deposition are eutrophication and acidification of aquatic and terrestrial
ecosystems. Eutrophication is defined here as the gradual increase and enrichment of nutrients
(nitrogen) to levels exceeding those that can be utilised (or absorbed) by a system and undesirable
changes begin to occur. Likewise acidification, which generally occurs after eutrophication, is
defined here as the reduction in water or soil pH from baseline conditions to a point at which
undesirable ecosystem changes occur. The implication of these definitions, particularly use of the
term “undesirable” is discussed throughout the review that follows.
In Canada, the majority of anthropogenic nitrogen is emitted to the atmosphere in the form of
(NO x ) (Niemi et al. 2009) by transportation, industrial, upstream oil and gas, and utility sectors
(CCME 2008a). Natural sources of NO x include combustion (forest or grass fires), lightening,
microbial and vegetative processes. Atmospheric NH 3 and N 2 O emissions are primarily from
agricultural sectors (Niemi et al. 2009). In addition to the RN that ecosystems receive from
atmospheric deposition, sources such as wastewater discharge and land application of N
fertilisers, manure and biosolids also contribute to biologically available nitrogen pools
(Bouwman et al. 2002).
Biological response to RN deposition varies between ecosystems and species. Being the second
largest country by landmass in the world, Canada is blessed with ecosystem diversity from boreal
and temperate forest, grassland, tundra, coastal communities and an abundance of freshwater.
Despite the complexity of nitrogen deposition and the diversity of its receptors, this report
11
summarises and reviews existing knowledge pertaining to the deposition and effects of RN in a
Canadian context. Additionally this report aims to answer the following questions:
1. What additional programs and/or research are necessary and/or desirable to assist decision
makers to improve the measured and modelled N deposition data in Canada?
2. What types of studies and/or research are necessary to better understand the physical,
chemical and biological effects associated with nitrogen deposition to major ecosystem
types in Canada?
3. What information and/or studies are necessary to determine how, when and where N
deposition should be considered eutrophying, acidifying and/or both?
4. What additional programs and/or research are necessary, or desirable, to develop critical
load values for eutrophication loads of nitrogen for sensitive ecosystems in Canada and to
develop an N deposition sensitivity map for Canada?
5. What are the eutrophication management strategies and/or (e.g. use of critical loads,
intensity or volume caps) or other tools for managing N-deposition that might be most
applicable or relevant to the Canadian context and from a scientific standpoint what
would be needed to develop such tools?
6. What would constitute a comprehensive national N deposition and effects monitoring
program and what additional monitoring would be required to implement such a national
monitoring program?
Reactive nitrogen deposition and effects in Canada require an inherently integrated and
interdisciplinary approach. The following three sections (2.0 – 4.0) provide the background and
research necessary to approach these questions. The questions themselves are answered in section
5.0 and in section 6.0 conclusions and recommendations are given.
2.0 Atmospheric RN Deposition
Earth’s atmosphere contains 78% gaseous N (as N 2 ), which is inert and cannot be utilised by
most organisms. Only a small portion of this N 2 is fixed by N-fixing soil bacteria to become
reactive or biologically active nitrogen (RN). RN can also be synthesised by lightning and
manufactured as fertilisers. Inorganic RN includes oxidized forms (NO x , HNO 3 , HNO 2 , N 2 O)
and reduced forms (NH x ). Organic RN compounds (OrgN in Figure 1) such as urea ((NH 2 ) 2 CO),
amino acids, amines, proteins, and nucleic acids (Cowling et al. 2002, Nielsen 2005) are not of
atmospheric origin, but may be transported to some degree as dust, liquid aerosols and particulate
matter (Neff et al. 2002; Calderón et al. 2007), and converted into RN by soil bacteria through the
process of mineralization. Here the focus is on inorganic oxidised (NO x , HNO x , N 2 O) and
reduced (NH 4 + and NH 3 ) forms of RN of atmospheric origin. However, due to linkages between
forms of RN in the biosphere (Figure 1), organic and non-atmospheric forms of RN must also be
considered.
12
Figure 1. Artistic rendition showing sources of reactive nitrogen focusing on atmospheric deposition
pathways. Atmospheric sources are shown with up arrows, sources of deposition (after chemical
transformation) are shown with down-arrows, sources that can be both from the surface and though
atmospheric volatisation (such as manure, fertiliser and municipal sources of NH 3 ) are underlined, and the
soil processes of nitrification, nitrogen fixation and mineralization are given by 1), 2) and 3) respectively.
The production of ozone via photolysis and the formation of aqueous ammonium are also shown. This
does not represent an exhaustive representation of RN sources or chemistry.
2.1 RN Chemistry
Nitrogen oxide (NO), and to a lesser extent nitrogen dioxide (NO 2 ) are formed from the
combustion, and hence oxidation, of anything in a nitrogen dominated atmosphere (Figure 1). NO
becomes further oxidised to NO 2 in the atmosphere, which after further interaction with water
and atmospheric gases, forms nitrate (NO 3 -). These compounds are referred to collectively as
(NO x ). In the free atmosphere NO x combines with the hydroxyl radical (OH-) to form nitric acid
(HNO 3 ). A very reactive and therefore short-live molecule, OH- is formed in the atmosphere
through reactions between sunlight, oxygen and water and its availability depends on the
availability of these reactants. NO 2 also combines with water (H 2 O) to form both nitric and
nitrous acids (HNO 3 and HNO 2 , respectively). Both of these acids contribute to the total RN pool
arriving at Earth’s surface (Bynerowicz et al. 1995) and HNO 2 is a major source of OH- in the
lower atmosphere (Acker et al. 2005).
Atmospheric NO x is also a precursor of tropospheric (ground level) ozone (O 3 ) and peroxy
nitrates (such as PAN) (Jenkin and Clementshaw 2002), both known for their ability to cause
direct damage to vegetation (Karnosky and Thakur 2004), and human health (Curtis et al. 2006;
Horvath et al. 1986). In the lower atmosphere O 3 , particulate matter (PM) and HNO 3 vapour
(Bytnerowicz et al. 2005) make up what is known as ‘smog’, named as a mixture of smoke and
13
fog, and known for its negative impacts on human health and visibility (Environment Canada
2004). Atmospheric NO 3 - is also rapidly neutralised by alkaline compounds such as NH 3 , Ca and
Mg to form coarse particles of biologically available (reactive) ammonium nitrate (NH 4 NO 3 ),
calcium nitrate (Ca(NO 3 ) 2 ), and magnesium nitrate (Mg(NO 3 ) 2 ) and/or their hydrated forms
(Sagar Krupa Persona. Communication January 2010). The formation of this PM from NO 3 - is
determined by the availability of alkaline compounds in the atmosphere and climatic conditions
such as temperature and humidity (Wang et al. 2009). Other RN reactions such as O 3 and HNO 3
formation are “photochemical” and depend (among other things) on the availability of sunlight;
and therefore time of day or year.
Ammonia (NH 3 ) is proving important in the formation of acidic particulate matter (PM)
(Bytnerowicz et al. 2007; Lillyman et al. 2009). When ammonia (NH 3 ) dissolves in water some
of it becomes ammonium (NH 4 +) in an equilibrium reaction in which the amount of either
molecule depends on the pH of the solution. Ammonia acts as a base in the atmosphere, meaning
it has the ability to neutralise acids including nitric and sulphuric acids produced by nitrogen and
sulphur oxides, respectively. The products of this neutralisation are ammonium salts (ammonium
nitrate (NH 4 NO 3 ) and ammonium sulphate ((NH 4 ) 2 SO 4 )) that exist as PM (Summers et al.
1986). However, in soil and groundwater NH 3 acts an as acid releasing acidic H+ as it is oxidised
into nitrate (NO 3 -) by soil bacteria (Kowalchuk and Stephen 2001; Avila-Segura et al. 2002). In
addition, ammonia has the ability to form acidic particles in the atmosphere (Lillyman et al.
2009)
In the lower atmosphere N 2 O is chemically stable i.e. it does not react with other atmospheric
constituents to make new compounds (Sagar Krupa Personal Communication, January 2010) and
therefore has a long atmospheric lifetime (>100 years). This long lifetime increases N 2 O’s
potency as a greenhouse gas – 310 times more so than CO 2 (Government of Alberta 2004) – and
in the upper atmosphere (stratosphere) N 2 O produces NO and NO 2 that destroy atmospheric
ozone (Crutzzen 1970). The stratosphere is where most N 2 O removal occurs, however some N 2 O
may also be taken up by agricultural (Majumdar 2009) and natural systems such as peatlands
through microbial reduction to inert N 2 (Roobroeck et al. 2009). The atmospheric chemistry of
RN and related compounds is complex and includes additional intermediate compounds (such as
N 2 O 5 ) and complex reactions not discussed here. Instead this report focuses on those compounds
that reach the surface and interact with living systems causing eutrophication and related effects.
2.2 Atmospheric RN Deposition
The RN chemical species that deposit on the surface of soil, water or vegetation may be deposited
in wet, dry, or particulate form and received as a solid, liquid, or gas (Table 1). The form of RN
present in the atmosphere, soil or surface water changes with moisture, oxygen availability,
biological processes, sunlight and the availability of free radicals (such as OH-). Deposition via
fog or low cloud (sometimes termed “occult precipitation”) is an important input to high
elevation forests (Summers et al. 1986).
While atmospheric pollutants such as O 3 , SO 2 , NO 2 , and PM are often measured as a
concentration in ambient air (in ppm or g/m3), deposition is quantified as a surface flux (i.e. in
g/m2/yr or eq/ha/yr) – the rate of delivery of a chemical species to a (soil, plant, monitor filter,
14
etc.) surface. The quantification of fluxes requires different tools than that of gaseous
concentrations. While the rate of deposition to a surface is related to the concentration of a
compound in the overlying air, the actual fallout rate (or “deposition velocity”) is determined by
a number of factors, particularly the chemical species of interest, the time of day and year, and
the surface itself.
Table 1. Examples of different RN compounds by state and depositional form. Developed using
information from Chambers et al. (2001); Russow and Böhme (2005); Acker et al. (2005); Moran et al.
(2009). This is not an inclusive list. Our knowledge and measurement of RN chemistry is constantly
evolving and RN may still be found in Canada’s atmosphere in different phases and forms.
State / Deposition
Particulate / Aerosol
Gas
a
Wet
HNO 2 , NO 3 -,
NH 4 +, NH 3 ,
NH 4 NO 3
a
NH 3 , NH 4 +, NO 3 -,
HNO 3
Dry
HNO 3 , NH 3 , NO 2 ,
NH 4 NO 3
HNO 3 , HNO 2 ,
NO 2 , NO, NH 3 ,
N2O
wet gaseous deposition would occur through water vapour i.e. fog or cloud and is referred to as “occult
precipitation”.
For instance it has been found that NO 2 deposition to forest canopies is much higher in broadleaf
tree species than conifers (Hanson et al. 1989) and that increased wind speeds increase transport
and deposition to a surface. Open water surfaces (e.g. lakes, wet vegetation canopies) have a
negligible resistance to (and therefore a relatively high deposition flux of) soluble RN species
(Wesely 2007) such as NH 3 , HNO 3 and HNO 2 . However, less soluble RN forms, such as NO
and NO 2 , deposit less freely to wet surfaces and surface resistance is an important component of
their deposition calculations. In contrast, the resistance of any surface to all particulate forms of
RN is considered negligible. Atmospheric (or aerodynamic) resistance to a compound is variable
and is an important component in the deposition flux calculations for all RN forms (Grennfelt
1987). How actual measurements of deposition differ from those of concentration is discussed in
section 5.3.1.
A number of units may be used to quantify atmospheric nitrogen deposition; and conversion
between units is a function of the molecular weight of the compound of interest (Table 2). For the
critical loads of nutrient N or eutrophication, units of kg/ha/yr are used, while for the critical
loads for acidity, units of eq/ha/yr or meq/ha/yr are used. Units of mol c /ha/yr are sometimes
employed for determining “critical loads” in an aquatic ecosystem and are interchangeable with
eq/ha/yr (Table 2). The deposition of both NO x and NH x may be quantified in kg/ha/yr, however
the molecular weight of the compound of interest determines the contribution of that substance to
total nitrogen, measured in kg N/ha/yr. For example NO 2 (what NO x is frequently represented as)
has a molar mass of 46.05g, 30.4% of which is N. Similarly, NH 3 has a molar mass of 17.30g,
82.2% of which is nitrogen. These percentages (ratios) are used to convert the relative
contributions of different RN compounds to total N (in kg N/ha/yr). So by mass, NH 3 contributes
more total N than NO 2 due to the relative mass contribution of both hydrogen (H) and oxygen
(O), respectively.
15
Table 2. Relationships among the units for atmospheric RN deposition particularly in terms of assessing
ecosystem sensitivity to eutrophication and acidification (adapted from CCME 2005).
Chemical species
NO 3 NH 4 +
N
kg/ha/yr
1.00
1.00
1.00
eq/ha/yr
16.1
55.4
71.4
meq/ha/yr
1.61
5.54
7.14
mol c /ha/yr
16.1
55.4
71.4
Although atmospheric pollutants are able to travel some distance from their emission source,
most surface deposition occurs near large sources or in areas of high emission density. The
Canadian regions receiving the highest NOx deposition are the densely populated regions of
southern Québec and Ontario, Alberta’s south, and Fort McMurray and the Georgia Basin of
southwestern British Columbia. However, estimates show that between 69-76% of Canada’s
deposited NOx originates in the southern Ontario and the eastern United States (Vet et al. 2005).
The highest NH 3 deposition occurs in the agricultural regions of southern Quebec and Ontario,
and cattle-rich Alberta (Niemi 2004). Vet et al. (2005) reported that 50-65% of eastern Canada’s
deposited NH 3 was of US origin, whereas in the west NH 3 originates from local Canadian
sources.
2.3 Sources of Atmospheric RN
The main component of atmospheric NO x is biologically produced NO occurring quite uniformly
throughout Earth’s atmosphere. However, man-made sources of NO x occur in localised parts of
the world and Canada enhancing RN in these areas (Bytnerowicz et al. 1998). Anthropogenic
nitrogen oxides (NO x ) originate primarily from combustion processes. In Canada in 2007 most
NO x was emitted from the burning of fossil fuels in the industrial and transportation (mobile)
sectors, accounting for 30% and 49% of nationally reported 2007 NO x emissions, respectively
(Figure 2a). Total annual NO x emissions in 2007 were 2.47 million tonnes. In eastern Canada
mobile sources dominate NO x emissions (64%), whereas in western Canada industrial sources are
more significant (40%) making mobile emissions less so (39%) (Environment Canada 2009a).
Despite total cross-Canada NO x emission decreases of 13.9% from 1995-2007, NO x emissions in
western Canada have increased significantly due to an expanded upstream oil and gas sector
whose NPRI reported emissions increased by 52.3% from 1995 to 2007 (Environment Canada
2009b). Marine transport emissions on both the east and west coasts of Canada are also gaining
significance in their contribution to both off-shore and in-shore atmospheric NO x loads
(Transport Canada 2009).
Major NO x emission regions in Canada include the Toronto-Ottawa-Quebec corridor, CalgaryEdmonton corridor, Fort McMurray oilsands region and Winnipeg. Areas surrounding the
population centres of Vancouver, Regina, Windsor and Halifax also contribute substantially to
federal NO x emissions (Niemi 2004). Alberta is the largest emitter of NO x in Canada, accounting
for 34.1% of national NO x emissions in 2007 (Figure 2b). Petroleum sector NO x emissions from
Alberta – including upstream extraction and processing, downstream refining, and transport or
distribution – reached 436,642 tonnes in 2007, accounting for 51.7% of total NO x emissions in
the province (Table 3).
16
Figure 2. Contributions to NPRI-included federal emissions of NO x and NH 3 by source grouping (a) and
province or territory (b).
Ammonia (NH 3 ) is emitted by biological processes in the decay of organic matter. It evaporates
from fertilised fields and animal waste where it enters the free atmosphere. Global NH 3
emissions reached 39 Tg/yr in the mid-1990s and are responsible for the production of fine
particles, atmospheric haze and acid deposition (Galloway 2003). In Canada total NH 3 emissions
increased by 4.0% from 1995 to 2007 reaching 503,137 tonnes (Table 3) down roughly 10%
from 2005 reported emissions of 558,290 (Environment Canada 2009b). Although NH 3 emission
rates have been relatively constant in eastern Canada, consistently increasing emissions in
western Canada since 1995 are due to the expansion of transportation, animal husbandry, and
agriculture leading to increased application of synthetic fertilisers, manures, biosolids, and
pesticides (Niemi 2004, Lillyman et al. 2008). In 2007 88.5% Canada’s atmospheric NH 3 was
emitted by the agricultural sector (Figure 2a). Total federally reported NH 3 emissions from
agriculture in 2007 were 445,540 tonnes, accounting for 98.8% of all open sources (Figure 2a).
Ammonia emissions in Canada are generally highest in areas of intensive agriculture such as the
Lower Fraser Valley of British Columbia, southern Alberta, southern Manitoba, southwestern
17
Ontario, and southern Quebec (Niemi 2004). In the agricultural regions of western Canada, NH 3
emissions are expected to increase by approximately 50% from 2000-2020 (Chambers et al.
2001) or 54% between 1995 and 2015 (CCME 2008a).
Table 3. Total year 2007 atmospheric emissions of RN included in the NPRI by province/territory (in
alphabetical order) and RN chemical species. All values are in t/y.
Province / Territory
Alberta
British Columbia
Manitoba
New Brunswick
Newfoundland Labrador
Nova Scotia
Northwest Territories
Nunavut
Ontario
Prince Edward Island
Quebec
Saskatchewan
Yukon
National total
NO x
844,622
279,355
90,233
55,894
54,591
77,666
20,988
11,887
479,395
6,469
287,650
256,404
8,129
2,471,283
NH 3
125,377
21,547
57,194
4,495
2,083
6,659
180
7
111,092
4,198
78,492
91,537
276
503,137
Nitrous oxide (N 2 O), and some NO x originate from microbial denitrification and nitrification,
respectively. N2O may be emitted to the atmosphere from agricultural fields, livestock operations,
nitrogen-based fertilisers and to a lesser extent from combustion processes – particularly engines
with catalytic converters. Although designed to reduce NO x and hydrocrabon emissions, these
converters increase emissions of other pollutants such as N 2 O, NH 3 and CO 2 (Colvile et al.
2002). N 2 O is transported through atmospheric deposition, municipal wastewater, and surface
runoff, and into ground- or surface-water (lakes, rivers and oceans). Once in surface water
additional N 2 O is produced from biological processes (Mosier and Kroeze 2000). The N 2 O
produced by the combining of oxygen and nitrogen in lightning storms is very small compared to
that emitted from anthropogenic and biologic sources. Although included in Canada’s NPRI
substance lists, there are no sources of N 2 O included in available NPRI emissions datasets from
2000-2008 (see Environment Canada 2009). However N 2 O emissions reported to Canada’s
Greenhouse Gas (GHG) inventory were 150,000 t in 2007 down from 180,000 in 1995 (Dupre
2009).
The data from Canada’s National Pollutant Release Inventory (NPRI) were used to summarise
year 2007 emissions of NO x and NH 3 from all provinces/territories including both anthropogenic
and natural sources (Table 3 and Figure 2). While the NPRI was developed as a tool for public
rather than scientific enquiry, as of 2002 the NPRI constitutes Canada’s primary tool for
emissions inventorying. In 2007 emissions reporting was required for facilities operating with 20
000 employee hours per reporting year or more (Canada 2007) – meaning that for industrial
18
emission sources with less than 10 full-time employees, reporting was not required. This
threshold is particularly significant when considering many of the small upstream oil and gas
facilities in Western Canada (Krzyzanowski 2009) and will be discussed further in the
recommendations section of this report.
2.4 Sources of Non-Atmospheric RN
Atmospheric deposition is only one of the pathways by which RN may enter ecosystems and
contribute to the total RN pool. In order to fully understand the fate of atmospheric RN in an
ecosystem, it is necessary to be able to quantify the existing RN stores in, as well as fluxes into,
that ecosystem. Like atmospheric RN discussed above, anthropogenic sources of ammonia,
nitrate, etc. that are released to land and water are also reported to the NPRI. However, like their
atmospheric counterparts, the reporting of these emissions is limited by exemptions and reporting
thresholds (Chambers et al. 2001). Non-atmospheric RN is released as biologically active NO 3 from chemical and manure-based crop and forest fertilisation. In both British Columbia and Nova
Scotia, water concentrations of NO 3 - related to forest fertiliser application have approached levels
of concern to human and ecosystem health (Chambers et al. 2001). Urea is sometimes used as a
fertiliser for crops and forests to enhance growth; which, like ammonia, dissociates into
eutrophying NO 3 - and acidifying H+ (also producing water and carbon dioxide) (Avila-Segura et
al. 2002).
Nitrogen fixation occurs in the soil via fixation by leguminous crops (such as beans and peas) as
well as wild legumes (Chambers et al. 2001) and some non-leguminous trees such as alder (Alnus
spp.) (Borman et al. 1993; Brockley and Sanborn 2003). While most of this fixed nitrogen is
utilised by plants for growth and the assimilation of protein or chlorophyll, excess N can lead to
nutrient imbalances in other species (Brockley and Sanborn 2003). Disturbance in watersheds,
from activities such as deforestation and road building also leads to an export of dissolved RN
(Eshleman et al. 2009). Nitrogen mineralization is the process by which organic nitrogen (in the
form of manure, decaying plant and animal matter etc.) is transformed in biologically usable
forms (primarily NO 3 - and NH 4 +) by bacteria as part of the process of decomposition. The rate of
RN released through mineralisation fluctuates with the availability of organic N, soil temperature
and moisture availability (Crohn 2004; Rixen et al. 2008); and is therefore closely and complexly
tied to both land-use and climate.
Other anthropogenic sources of RN include wastewater discharge to surface and groundwater, the
disposal of biosolids, sewage and waste from aquaculture (Chambers et al. 2001), which produce
atmospheric NH 3 and N 2 O through bacterial processes. Chemicals used in fire retardants,
explosives, pesticides, plastics and by-products of industrial processes (Domene and Ayres 2000)
also contribute to ecosystem RN inputs through waste and volatisation. The movement towards
nitrogen and phosphorus based pesticides and away from organochlorides has further increased
nutrient exports from agricultural areas (MacDonald et al. 2000). High inputs of RN to Earth’s
surface from non-atmospheric sources, combined with atmospheric deposition, can have additive
and interactive effects on ecosystems, resulting in excessive nutrient and H+ concentrations in
soils and surface-water, instigating a myriad of potential biological and chemical ecosystem
responses.
19
3.0 Effects of RN Deposition
All plants depend on reactive nitrogen (RN) for growth and for the creation of amino acids,
complex proteins and chlorophyll. Henceforth, RN is a moderator of ecosystem development, but
may interact with plants, soil and animals either beneficially or detrimentally, depending on the
dosage. Excess RN may significantly change biogeochemical N cycling and is linked to
ecosystem eutrophication, acidification, health and productivity (Nielsen 2005). The scale, degree
and speed of ecosystem eutrophication and acidification are closely linked with ecosystem
characteristics.
3.1 Eutrophication
Eutrophication occurs from the perpetual input of the nutrient elements (primarily N or P)
through transport pathways such as atmospheric N deposition, N and P released from fertilisation
and wastewater discharge, and symbiotic N fixation by N-fixing plants. Eutrophication occurs in
both terrestrial (e.g., forests, grasslands, tundra) and aquatic ecosystems (e.g., rivers, lakes,
oceans), by different processes. In terrestrial ecosystems eutrophication is made possible by
nitrogen enrichment – N usually limits growth. Terrestrial ecosystems that are naturally nitrogen
limited are the most susceptible to changes caused nitrogen enrichment. Globally, temperate and
boreal ecosystems are particularly N-limited (UNECE-CLRTAP 2009), meaning that Canada’s
terrestrial ecosystems are especially vulnerable to eutrophication. In aquatic ecosystems, N is a
supplementary nutrient for eutrophication – phosphorus (P) usually limits growth (Chambers et
al. 2001, Schindler et al. 2006; Schindler et al. 2009). Most inland waters of Canada are
intrinsically P limited and thus additions of P will accelerate, and increase the risk of,
eutrophication (Chambers et al. 2001). In coastal marine waters however, growth is significantly
limited by nutrient nitrogen even in conditions of elevated P (Fenn et al. 2003). In coastal waters,
like terrestrial ecosystesms, N is the limiting nutrient (rather than P) and atmospheric RN
deposition can induce eutrophication more easily in coastal, rather than inland, aquatic systems.
Although much of the research on eutrophication in Canada over the past three decades has
focused on phosphorus (P), there is increased evidence of RN’s involvement in eutrophication
(CCME 2008a). Nitrogen mediated eutrophication is the focus of this section. By 2001 nutrient
(N and P) enrichment had been found in every Canadian province (Chambers et al. 2001). While
the territories were yet to show signs of enrichment, it is unknown whether this is due to a lack of
population centres, agriculture and industry, or simply a lack of monitoring in the north. More
recent studies show that algal growth in northern Canadian lakes is co-limited by the availability
of both N and P (Ogbebo et al. 2009).
In Canada nutrient enrichment occurs in watersheds where there are naturally high concentrations
of N and P, large nutrient-loadings from air, land and water, or small and shallow lakes where
nutrients accumulate (Government of Canada 1996; Hall et al. 1999). Large rivers, like the
Saskatchewan, Bow, or St. Lawrence that are naturally low in nutrients experience a water
quality decline, or nutrient enrichment, downstream of urban centres (Government of Canada
1996). The same is true of smaller rivers like the Yamaska in southern Quebec (Chambers et al.
2001). Lakes are also subject to local nutrient (RN) inputs. For instance the Great Lakes, or
Ontario’s Simcoe and Rice lakes, are influenced by a myriad of municipal, agricultural and
industrial sources of RN. Inland waters throughout Nova Scotia and New Brunswick are
experiencing a similar fate (Figure 3). Areas with intensive agriculture such as the Windsor –
20
Quebec corridor, the Lower Fraser Valley in British Columbia, and Southern Manitoba and
Saskatchewan, also make large contributions to Canada’s RN pool through surface and
atmospheric releases of NH 3 .
Figure 3. Documented sites of nutrient enrichment in 1998 (from Chambers et al. 2001). Green areas
represent impacts from agriculture, red from industry, yellow from municipal effluent, grey from other
and natural factors, and the light green tinge highlights agricultural areas.
There are (and have been) some signs of eutrophication in southern Ontario, the oil sands regions
of Alberta and the Georgia Basin of British Columbia due to increases in local sources of RN
(Schindler et al. 2006). On the west coast, agricultural and urban centres in the lower Fraser
Valley, along with some industrial nutrient sources in the upper Fraser and lower Columbia
rivers, are the largest sources of nutrient enrichment. An estimated 90% of British Columbia’s
municipal wastewater is discharged into the lower Fraser or its tributaries (Government of
Canada 1996) providing sources of both NH 3 and NO 3 -. Although, there is currently little
21
evidence of nutrient enrichment in Pacific coastal marine waters, likely due to high rates of
flushing, the 1960s saw fish-killing toxic algal blooms in this region (Anderson et al 2008).
These blooms are one of the symptoms of eutrophication discussed below, that Canada’s east
coast experienced in the 1980s (Martin et al. 1990). These blooms may become more common on
Canada’s Pacific coast, as a result of the high nutrient outputs that are likely associated with
British Columbia’s aquaculture industry (Islam 2005).
In Canada’s northern boreal forest, stretching from British Columbia across to Newfoundland
Labrador, nutrient enrichment (in the 1990s) occurred only in localised areas downstream of
municipal or industrial outflows on the Athabasca and Wapiti rivers in northern Alberta
(Chambers 1996; Wrona et al. 1996). However, with increased development in the north –
especially from the upstream oil and gas sector in northern British Columbia and Alberta, RN
emissions are increasing in Canada’s high latitude regions. Local inputs increase the risk of
nutrient enrichment and eutrophication in northern regions. Most of Canada’s temperate
ecosystems are N-limited, and enhanced RN inputs initially stimulate plant growth (Chambers et
al. 2001). However, these beneficial effects diminish as other nutrients such as calcium (Ca) or
magnesium (Mg) become limiting (or used up) at enhanced growth rates. Eutrophication can only
occur when all other nutrients are in abundance and therefore not limiting to growth (Garham et
al. 1986).
3.1.2 Biological effects
The form of N and type of deposition (wet or dry) influences the response of vegetation and soil
or water systems to that deposition. In terrestrial ecosystems the direct deposition of RN
compounds such as NO, NO 3 -, NH 3 to plant foliage can cause direct injury to natural and seminatural vegetation (Bobbink et al. 2002; Krupa 2003; Bytnerowicz et al. 2007). Vegetation takes
up NO and NO 2 through the leaf cuticle where through a series of metabolic reactions these RN
compounds are transformed into proteins and amino acids and utilised for growth. However the
ability of a plant to reduce and utilise this RN is determined by the availability of sunlight and
moisture and the species’ ability to reduce foliar nitrate. Excess foliar nitrate can lead to
symptoms of visible injury (e.g. leaf necrosis, or abnormal needle elongation) and increased
susceptibility to other environmental stressors (Bytnerowicz et al. 1998). Similarly, HNO 3 ,
HNO 2 may also cause direct injury to vegetation at concentrations above which can be
metabolised by leaf tissue (Bytnerowicz et al. 1998; Bytnerowicz et al. 2005; Bytnerowicz et al.
2010); and secondary pollutants formed through chemical transformations of NO x in the
atmosphere, such as PAN and O 3 , are also known to cause damage to vegetation through foliar
deposition and uptake (Karnosky and Thakur 2004). Ammonium is directly toxic to plants
through foliar uptake (Britto and Konzruker, 2002; Bytnerowicz et al. 2101) and may cause acute
visible effects on vegetation (Krupa 2003).
When RN deposits on the soil surface it is first taken up by vegetation and has a fertilising effect
on plants, thereby increasing growth. Because N is necessary for a plant’s assimilation of
proteins, DNA, and chlorophyll, this enrichment can be beneficial to plants in environments that
are normally low in RN. However, there are limits to the benefits of this growth and not all
species or ecosystems respond similarly. One of the most difficult parts of RN effects assessment
is determining the stage at which this enrichment, or even enhanced growth, becomes detrimental
22
to the function of the ecosystem as a whole, or when effects become “undesirable”. In addition,
when RN arrives at soil in reduced form (i.e., NH 3 or NH 4 +) the uptake of other cations such as
K+ and Mg2+ may be suppressed (Britto and Konzruker, 2002).
Mosses and lichens are particularly adept at retaining and metabolising incoming RN (Carroll et
al. 2000, Heijmans et al. 2002) and due to direct uptake of nutrient via atmospheric pathways, are
particularly sensitive to direct RN exposure, especially agriculturally derived NH 3 . However, at
very high concentrations, as with vascular plants, NO 3 - builds up in tissue (Carroll et al. 2000)
and direct damage, or decline may occur (Cunha et al. 2002). Excessive N inputs have also been
shown to cause declines in the cover and density of other bryophytes (Lee and Caporn 1998;
Carroll et al. 2000). However, the physiological response of bryophytes and vascular plants to
atmospheric RN deposition depends on the species, and a number of environmental factors.
Similarly RN effects occur on higher woody plants such as trees; but whether RN impacts growth
in a positive or negative way is determined by soil nutrient status (Bobbink et al. 2002). RN
deposition has also been associated with biomass reduction in vascular plants; decreases in shoot
Ca and Mg concentrations (Baron et al 2009); increases in shoot/root ratios; and reductions in
root colonisation by fungal mycorrhizae (Bobbink et al. 2002; Treseder 2004; DeVries et al.
2007) under conditions of enhanced soil RN. This response also appears to be species- and
environmental variable - specific.
Nitrate leaching is a typical symptom of eutrophication in terrestrial ecosystems that is closely
associated with N saturation. Nitrogen saturation occurs in a soil system in either wet or dry
states and is marked by more RN in the soil matrix than can be utilised by plants and
microorganisms. The unused RN is lost by leaching in the form of NO 3 -, or is emitted to the
atmosphere as N 2 O and NH 3 (Stoddard 1994). Nitrogen saturation is not only dependent on total
deposition, but also on land use, vegetation cover, and forest stand age and soil type (Fenn et al.
1998; Aber et al. 1997). Once a terrestrial system becomes N-saturated, NO 3 - is leached from the
system into surface and groundwater. The leaching of NO 3 - through a soil system is dependent on
water availability and if drought prevails RN will build up in the soil. Once leaching occurs,
NO 3 - has additional effects on the aquatic ecosystems it enters.
Elevated NO 3 - in aquatic systems primarily promotes the excessive growth of algae (Carpenter et
al. 1998), which like terrestrial plants, utilise RN in protein and amino acid synthesis and growth.
Cyanobacteria or “blue-green algae” are bacteria rather than algae, but are photosynthetic and
also experience increased growth under nutrient-enriched conditions. Some planktonic algae and
cyanobacteria produce toxins that render water unfit for consumption (Lawton and Codd 1991)
and pose threats to both human and livestock health through either the direct ingestion of water
(Chambers et al. 2001) or the ingestion of organisms that have fed on toxic algae (such as
mussels) (Smith et al. 1999; Kane et al. 2009). These toxic algal blooms have occurred with “red
tides” in coastal areas of Canada and around the globe (Mudie et al. 2002).
Marine environments are generally more sensitive to eutrophication than freshwater lake or river
systems because the ocean is thought to be nitrogen limited, unlike freshwater systems, which are
generally phosphorus limited. However, we are realising that marine systems may be able to fix
more nitrogen than originally believed (Brandes and Devol 2002) and that some harmful
cyanobacteria are capable of fixing their own nitrogen. This means that these salt-water systems
may not be as nitrogen limited as previously thought. The effects of nitrogen enrichment are
23
complex and depend on the availability of numerous micronutrients such as silica required for the
growth of diatoms, or particulate RN required for certain dinoflagellates (Heisler et al. 2008).
In estuaries the situation is further complicated by the fresh-salt water interface (Schindler et al.
2008). In lakes most nuisance algae species are capable of nitrogen fixation, however this ability
is hindered by high cellular NH 3 (Mandaville et al. 1999) that can occur from the contamination
of surface waters with agricultural runoff. NH 3 is also acutely toxic to aquatic life (Chambers et
al 2001) – a toxicity that increases with increased water temperature and pH (Saskatchewan
Environment and Public Safety 1988).
The most apparent and well-known effect of enhanced RN on aquatic systems is the increased
growth of algae and its related outcomes. However nitrogen enrichment can affect higher aquatic
plants and animals especially in oligotrophic systems. Certain benthic invertebrates (larval insects
forms etc.) in streams and ponds are known to be sensitive to enhanced nutrient levels (Butzler
and Chase 2009) and low oxygen levels (Gooday et al. 2009). Marine benthic invertebrates are
particularly sensitive to increased nitrate levels (CCME 2003). The growth of non-algal aquatic
plants (macrophytes), may respond positively to an initial addition of nutrients but begin to
decline (as do zooplankton) if additions become perpetual; unlike algae and phytoplankton whose
growth continues to respond positively to increased nutrient additions (Butzler and Chase 2009).
The rate (flux) and timing at which nutrients are added to marine (Harrison 1978; Heisler et al.
2002), freshwater (Butzler and Chase 2009), or terrestrial (Nellemann and Thomsen 2001)
ecosystems has been found to be more important in determining biological response than the total
amount of nutrient added. This has important consequences for how nutrient loads are managed
and assessed.
3.1.3 Chemical effects
The anthropogenic input of N into systems though atmospheric deposition, fertiliser application,
and the disposal of human or animal wastes, leads to a number of chemical changes in soil and
aquatic systems. Many of these changes either cause, or are caused by, the biological effects
described above. For instance, the chemical properties of soil are a result of bacterial, and
chemical exchange, processes. The input of excess RN to a soil system leads to the long-term
accumulation of N in forest soils (Burns 2004; Sullivan et al. 2004). Much of this N is utilised for
plant growth and is therefore retained by the ecosystem. As discussed above, this enhanced
growth is also associated with the enhanced uptake of, and therefore lower soil contents, of
carbon and base cations. However, NH 3 deposition (O’Neil and Wilkinson 1977) as well as urea
((NH 2 ) 2 CO) and NO 2 - (Hütsch 1998) reduce a soil’s ability to take up and oxidise methane
(CH 4 ) – an important greenhouse gas and source of soil carbon stores. Increased nitrification
(from increases in NH 3 and NO 2 ) leads to decreases in soil organic matter and reduced C:N
ratios, a condition that also seems exacerbated by increases in climatic variations (Aber et al.
1997).
Although classified as RN and taken up to some extent by micro-organisms (Majumdar 2009;
Roobroeck et al. 2009), N 2 O is considered to be more important as a greenhouse gas, destroyer
of stratospheric ozone (Crutzzen 1970) and a symptom, rather than source, of RN-mediated
eutrophication. Excess NO 3 - increases biological denitrification (the removal of nitrate of NO 3 -)
thereby increasing the emissions of N 2 O to the atmosphere (Mosier and Kroeze 2000; Berendse
et al. 2001; Roobroeck et al. 2009). Soil RN at levels beyond which can be utilised (or
24
metabolised) by plants and micro-organisms, eventually leaches from the soil as NO 3 -, leading to
base cation depletion, and increased toxic Al3+ availability (Murdoch 1998; Bobbink et al. 2002).
Although NH 4 + is bound to the soil column and usually accumulated instead of leaching like
NO 3 -, NH 4 + is reduced to NH 3 in soil, and is nitrified (oxidised) into usable NO 3 - (and acidifying
H+), which can then exist in excess and become leached from the system (Bobbink et al. 2002).
For every 1 mole of NO 3 - that is leached from the rooting zone, 1 mole of H+ remains (AvilaSegura 2002) meaning that NO 3 - also is acidifying. However this H+ release is only acidifying
when there is no NO 3 - uptake by plants or soil organisms i.e. when RN is in excess and the
system is saturated. This is because when plants take up NO 3 - they release OH- to the soil in
order to maintain ionic neutrality in the soil complex. OH- is alkaline and neutralises the H+
associated with NO 3 - leaching, or that which may be associated with RN that deposits on the soil
as HNO 3 ; which along with NH 4 NO 3 , is a typical depositional form of RN in Canada (Julian
Aherne Personal Communication, January 2010). Studies conducted in recent years show that no
Canadian forests soils are currently N saturated (Houle 2005, Jeffries et al. 2005) and while this
implies that NO 3 - leaching will not be occurring, it does not mean that there will be no biological
or ecosystem effects from increases in available RN.
Much of the atmospheric RN that arrives in surface water gets there after leaching through the
soil complex following N-saturation. However, both direct atmospheric RN deposition and
surface-borne nutrient loads also contribute to eutrophication. The role of atmospheric RN
deposition in aquatic eutrophication varies with the chemistry and nutrient status of a water body.
If P is the principal limiting nutrient, no biological impacts directly attributable to N deposition
are expected in freshwater lakes (INDITE 1994; Hessen et al. 1997a; Chambers et al. 2001).
However, the effects of atmospheric N deposition can be found in both N-limited marine
ecosystems, and those freshwaters with P such that it is no longer a limiting nutrient (Moss et al.
1997). Much of the effect of RN deposition has to do with the occurrence and efficiency of
nitrogen fixing organisms. If nitrogen fixers are active and plentiful in a soil or water system,
then the system is not N limited, and is consequently less likely to be affected by increases in RN.
In P-limited freshwater systems, increases in RN will simply mean increases in total water N.
However, if the system is N-limited and eutrophication occurs, oxygen depletion and inputs of
toxic organic substances are some of the potential water chemistry results of increased algal
growth. If N-saturation and leaching are occurring in the surrounding basin’s soil, surface water
may also receive inputs of H+, aluminium and other metals associated with soil acidification.
Changes in N dynamics may have various impacts on nutrient limitations, such as switches from
N limitation to P limitation and vice versa (Vitousek and Howarth, 1991; Erisman et al. 1998). Nsaturation may induce changes in N/P or N/C ratios (Kelly et al. 1990; Gahnstrom et al. 1993;
Hessen et al. 1997a,b). Changes in the N/P ratio affects species of algae and phytoplankton,
which in turn may bring a changes to entire ecosystem structure (Downing and McCauley 1992;
Gahnstrom et al. 1993).
3.1.4 Whole ecosystem effects
By affecting organisms that are sensitive to RN increases – such as algae, grasses, or benthic
invertebrates - RN deposition can also influence entire ecosystems by promoting changes in
structure, biodiversity and habitat quality. The deposition of atmospheric RN to terrestrial
ecosystems may cause shifts from native to non-native species, favouring plants with a high N
25
demand (Smith et al. 1999, Carroll et al. 2000, Stevens et al. 2004, Nordin et al. 2005). These
nitrophilous (nitrogen-loving) plants often qualify as nuisance or invasive species (Kennedy
2003) such as has been reported by shifts from native to non-native species of moss in Sphagnum
bogs (Berendse et al., 2001). Shifts in community structure, such as from wildflowers to sedges
and grasses in alpine tundra (Burns 2004; Sullivan et al. 2004), or from heathland to grasslands in
parts of Europe (UNECE-CLRTAP 2009), are a common result of terrestrial eutrophication as
nitrophilic species become favoured under high RN deposition. These species shifts may
substantially reduce plant species diversity. The diversity of ground-cover in European forests
was found to decrease dramatically at N depositions of greater than 14 kg/ha/yr (DeVries et al.
2003).
Grasslands with high N retention and C storage rates are particularly vulnerable to excess RN
deposition (Cunha et al. 2002). Shifts in C and N cycling due to increased nitrate mineralisation,
N losses and reductions in carbon storage are associated with losses of biodiversity and species
shifts (Wedin and Tilman 1996). In the Great Plains of southern Alberta, Saskatchewan and
Manitoba, increased nitrogen deposition has been related to the expansion of forests into native
temperate grasslands (Köchy and Wilson 2001) which will in turn affects large ungulate grazing
capabilities and alters species’ composition. This expansion of forests into grassland may be due
to the ability of invasive tree species to alter nitrogen cycling for their own benefit (Laungani and
Knops 2009).
Shifts in nutrients occur when increased growth from RN enrichment leads to the subsequent
enhanced uptake of other nutrients (Ca, Mg, K, P, etc.) required for growth, which subsequently
become limited if the demand exceeds the supply of those nutrients. These shifts can lead to
imbalanced nutrition of tree species (Burns 2004; Sullivan et al. 2004). When trees take up excess
RN, the deficiencies of other nutrients increase forest susceptibility to stresses such as insect
attack and drought (Bobbink et al. 1996; UNECE-CLRTAP 2009) by altering biochemical
pathways of defence.
Shifts in plant species, growth, and nutrition also affect the quantity and quality of available
animal forage and human food sources in those ecosystems. For instance while red raspberries
(Rubus ideaus) are a nitrophilous species whose colonisation and growth is enhanced by
increased RN (Nordin et al. 2001) other species such as bearberry (Arctostaphylos uva-ursi) show
negative responses to increased RN availability (Turkington et al 1998). In addition, nitrate
availability is closely linked with plant nutritional quality / quantity, and high plant tissue N is
associated with increased rates of herbivory due to increases in protein content (Jefferies and
Maron 1997; Smith et al. 1999). High nitrate concentrations, associated with foliar build-up when
proteins are not further synthesised, are also associated with human health effects and the
formation of carcinogens following ingestion (Gangolli et al. 1994). In this way, RN deposition
may have indirect effects on terrestrial wild animal species, forest dependent First Nations
communities, and other citizens relying on locally produced food sources, in addition to the
potential direct health effects caused by high ambient concentrations of atmospheric RN.
Although there is little evidence of direct links between RN deposition and effects on terrestrial
animal species, faunal diversity and floral diversity are inexplicably tied (DeVries et al. 2007)
through patterns of herbivory and predation.
26
Climate change will further influence species composition, nutrient dynamics and the sensitivity
or response of vegetative communities to acidification in Canada’s terrestrial ecosystems. For
instance, at higher temperatures plants’ detoxification abilities increase offering improved
resistance to damaging pollutants (Bytnerowicz et al 2010). Water is required for plant growth
even more so than nitrogen and is also a growth-limiting factor. Reductions in soil water from a
warming climate would favour drought resistant plant species, which also generally require less
nitrogen for growth (Fichtner and Schulze 1992). In this case, ecosystems would become
nitrogen saturated more quickly than if dominated by water-loving nitrogen-limited plant species.
On the other hand, if water availability increases from enhanced snowpack or glacier melt,
enhanced runoff would increase the rate at which excess nitrogen enters waterways. However, if
RN deposition does not exceed levels which can be utilised for growth, the resulting growth
increases equate to increased biomass production and therefore increased carbon sequestration by
plants – particularly forest trees (Jassal et al. 2010). In the boreal peatlands of Alberta deposition
of 1.45-3.76 kg-N/ha/yr caused significant increases in vertical peat accumulation from
Sphagnum spp. growth thereby increasing cumulative carbon storage (Turetsky et al. 2003;
Gunderson et al. 2009) and taking up some of the excess atmospheric CO 2 associated with
climate change. However at higher RN doses, rates of productivity and decomposition in the
Sphagnum mosses of peatlands decline, consequently reversing the carbon sink (Bobbink et al.
2002; Kivimaki et al. 2007).
Similar to the effects in terrestrial ecosystems, RN enrichment in aquatic ecosystems leads to
growth enhancements, which also lead to changes in species composition and biodiversity (Burns
2004). Oligotrophic (nutrient poor) aquatic ecosystems are the most sensitive (Mandaville et al.
1999; Baron et al. 2009; Ogbebo 2009) because they have adapted to function under conditions
of low nutrient nitrogen. In addition, marine and river environments with low-flow or flush are
also particularly sensitive (Mandaville et al. 1999). In aquatic ecosystems algal growth is
responsible for a number of negative biological, chemical and physical effects. Photosynthetic
algae utilise large amounts of oxygen during respiration, and after death oxygen is utilised for
decay processes. Due to their quick growth and short life span, RN-supplemented algae often
create such low oxygen conditions (anoxic) that ecosystems are unable to support fish
populations (Kane et al, 2009; US-EPA 2010) in freshwater and marine ecosystems. Most
organisms, other than some forms of anaerobic bacteria and yeasts, require oxygen for life and
will also die if water conditions become anoxic.
Enhanced growth of planktonic (free-floating single celled) algae, filamentous algae and aquatic
plants can eventually result in reduced water clarity (Smith et al. 1999) from high organic carbon
associated with dead and decaying organisms. This increased turbidity can influence the
predation success of fish by reducing their ability to see their prey (Mandaville et al. 1999).
Aquatic plants receive less sunlight when turbidity is increased, which leads to declines in
photosynthesis, further reducing oxygen in the water system. Increases in algal growth may
however promote colonisation by species of Daphnia, which can effectively clear the water
column of dead and live algal biomass through grazing (Edmonson 1994). Although an increase
in primary productivity will generally increase total aquatic herbivore numbers, “desirable” fish
species are lost first. These cold-water deep dwelling fish disappear when oxygen depletion
occurs in deep water due to biological decay of algae at the lake bottom (Mandaville et al. 1999).
As eutrophication progresses oxygen is depleted from further up the water column (i.e. closer to
the surface) and more species are lost in lake and coastal marine aquatic communities.
27
Increased algal growth has also been shown as the cause of increased snail populations that act as
hosts for a parasite responsible for deformities seen in amphibians throughout North America
(Johnson et al. 2007). Decline in amphibians in southern Ontario (Chambers et al. 2001; Rouse et
al. 1999) and British Columbia (De Solla et al. 2002) have been linked to long-term exposure to
elevated nitrate concentrations due to the sensitivity or their larval (aquatic) forms. Amphibians
are among the most sensitive organisms to nitrate exposure, and may show developmental or
behavioural changes at levels below human drinking water guidelines (CCME 2003).
3.1.5 Societal effects
Some of the effects of eutrophication have obvious societal implications – such as the complete
loss of sport-fish from an important fishing lake. Other recreation values are also lost in aquatic
areas from eutrophication. Swimmers and fisher-people may be deterred from turbid water, less
light penetration and deep organic lake sediments. Similarly the odour sometimes associated with
sulphur-reducing bacteria in RN-enriched lakes can be a deterrent for boaters, hikers and
picnickers in parks and other locales with firm tourism and recreation values. The excessive
growth of macrophyte weeds can cause water use impairments and the blockages of screens,
filters or engines. Commercial fisheries and sea-freight may be negatively impacted by these
symptoms of eutrophication in marine environments. Toxic algal blooms also lead to toxicity
and deaths higher up in the food chain through bioaccumulation, such as that associated with
various types of shellfish poisoning (Anderson et al. 2008).
Increases in RN inputs lead to chronic increase in surface water NO 3 - (Burns 2004; Sullivan et al.
2004), which may increase the frequency and spatial extent to which the drinking water guideline
for nitrate is exceeded in ground water (Chambers et al. 2001) and have detrimental
consequences for human health (CCME 2003). Increases in surface- and groundwater NO 3 - and
potentially toxic algal blooms, create an economic burden to Canadians in the form of healthcare
and water treatment, monitoring and remediation expenditures (Chambers et al. 2001). Additional
factors such as the taste, odour, clarity and colour of water also impact water usage and treatment
options.
Although not as apparent as freshwater effects, eutrophication effects on terrestrial ecosystems
also have societal consequence. Changes in forest structure and biodiversity alter the
appropriateness of forest and land use management strategies. Tourism and wildlife values can be
impacted through changes in wild animal diversity and numbers. Although there is little direct
evidence for the impacts of eutrophication on wild terrestrial animals, changes in floral diversity
are closely linked to faunal diversity (DeVries et al. 2007) through complex webs of herbivory
and predation. Changes in faunal species may negatively impact tourism, recreation, and
conservation. In addition, changes to either floral or faunal diversity will have direct and
potentially severe impacts on First Nations’ communities that exist as part of a natural or forestbased economy and rely on wild sources of food and wares. As in aquatic systems, the effects of
eutrophication are expected to cascade up the food chain and be felt at all levels. The loss of any
species, whether aquatic or terrestrial, has unmeasurable effects on Canada and its people.
3.2 Acidification
Nitrogen oxides (NO x ) (together with sulphur dioxide (SO 2 )) react with water, oxygen, and
28
sunlight in the atmosphere to produce nitric acid (HNO 3 ) and sulfuric acid (H 2 SO 4 ). These acidic
compounds are the primary agents of acid deposition (van Egmond et al. 2002, Erisman et al.
2002, Hanson 2008). Acids are in “equilibrium” when dissolved in water such as lakes and
rivers, clouds, rain droplets or soil water. Both sulphuric and nitric acids are “strong acids”
meaning that at equilibrium they are almost completely dissociated into their ions of SO 4 2- + H+
and NO 3 - + H+, respectively. For this reason, when discussing the role of RN in acidification, we
refer to concentrations and measurements of nitrate or NO 3 -. Unlike NO 3 -, SO 4 2- is biologically
inactive and is therefore a very mobile anion in ecosystems. This mobility is responsible for
SO 4 2- playing the primary role in acidification, whereas biologically active NO 3 - is largely
retained in catchments and does not become mobile until biological requirements are exceeded
(Driscoll et al. 2001, Jeffries and Ouimet 2005). When NO 3 - becomes a mobile anion it acts as an
acidifying agent akin to SO 4 2- (Aber et al. 1998). Although sulphate (SO 4 2-) deposition plays a
dominant role in Canadian ecosystem acidification and nitrate (NO 3 - ) deposition is generally less
significant, as RN sources increase and ecosystems become nitrogen saturated, atmospheric RN
may begin to play a role in ecosystem acidification in Canada.
Since RN input is rarely above and beyond what can be utilised for growth in an ecosystem, the
leaching of NO 3 - that is responsible for acidification is not commonly found in Canadian
terrestrial ecosystems (CCME 2008a). While water bodies in Canada have experienced
symptoms of eutrophication, and nitrate leaching has been noted in agricultural areas of Canada
(Farm Centre 1998), more recent studies suggest that no Canadian forests are currently N
saturated (Houle 2005, Jeffries and Ouimet 2005). However the capacity of forest ecosystems to
accumulate N is finite and regions experiencing elevated levels of N deposition will likely
become saturated at some time in the future without RN emissions reductions. Additionally, in
the winter or early spring when soils are frozen and plant life is dormant, NO 3 - may contribute to
soil and water acidification because it is not absorbed and retained by ecosystems (Jeffries 1990;
Laudon et al. 2002). Nitrate stored in snowpacks can be released quickly during melt causing
rapid and significant pH depressions in streams as it passes over frozen soils. This phenomenon
has been noted all over temperate North America, with the most significant Canadian event
causing a pH drop to 3.4 in the Mersey River of Kejimkujik National Park, Nova Scotia
(Wigington et al. 1992). Similar flushes may be expected from hard rain following a period of
drought.
3.2.1 Chemical effects
As mentioned above, RN is not acidifying until ecosystems reach a state of nitrogen saturation
and atmospheric inputs of RN occur at a greater rate than can be utilised biologically for growth,
protein assimilation, etc. Soil acidification is the loss of buffering capacity or acid neutralising
capacity of the soil (defined as the difference between cations and anions). Acid deposition to
calcareous soils (on limestone bedrock) causes the release of calcium (Ca2+) and carbonate
(HCO 3 -) ions, which are able to neutralise incoming acidity until used up (Bobbink et al. 2002).
In silicate (rather than carbonate) soils, RN saturation leads to NO 3 - and H+ leaching through the
soil matrix, disrupting base cations (Ca2+, Na2+, Mg+, K+) from soil particle surfaces replacing
them with H+. This leads to a depletion of soil base cations and the input of H+ makes the system
more acidic and reduces soil pH (Galloway et al. 2003). The eventual depletion of base cations
leaves a nutrient deficient rooting layer. If deposition rates of K, Ca, Mg and Na in throughfall
are significant, this will offset soil cation leaching and counteract soil acidification caused by
atmospheric deposition (Watmough et al. 2005).
29
Similarly, the impact of acid deposition on lakes is strongly dependent on the buffering capacity
of the lake ecosystem. This ability to neutralise incoming acidity is a function of the surrounding
soil and parent material as well as the atmospheric inputs of neutralising base cations. For
example, lakes embedded in granite (silicate) basins will have a very low buffering capacity, and
be sensitive to acid deposition, whereas lakes in limestone parent material have a large buffering
capacity and are therefore less sensitive (UNECE-CLRTAP 2004), Much like their terrestrial
counterparts, freshwater systems utilise calcium carbonate buffering to neutralise incoming
acidity. Low-lying lakes can typically be thought to represent the average chemical and nutrient
status of their basins. Rivers are less often used in acidification research because they are
chemically variable along their reaches and in time. While research in recent years shows that
oceans may be at risk of acidification, it is a process caused by CO 2 rather than RN enrichment
and is therefore not discussed further.
When the pH of soils becomes very low, nitrification is hampered and ammonia build-up may
occur (Roelofs et al. 1985). Additionally, at a pH <5 silicate clay minerals begin to break down
(Bobbink et al. 2002). During this process aluminium (Al3+) – once bound to the clay minerals becomes mobilised within the soil solution. Some forms of Al are toxic, particularly those that
exist at low pH (<4.5) (DeVries et al. 2003). Other toxic metals can also be mobilised by
acidification, but Al3+ is usually referred to as it is the most abundant metal on Earth, and is
therefore usually in ample supply.
3.2.2 Biological effects
The forms of Al that are leached through soil complexes in times of acidification are toxic plants
and fish (Palmer and Driscoll 2002) and may be responsible for much of the biological injury
associated with soil and water acidification (Henriksen and Posch 2001; UNECE-CLRTAP
2004). Biological toxicity from other metals (such as Pb, Cu, Cd, Zn) can also occur as a result
of acidification (Bobbink et al. 2002). In addition direct toxicity of metals and H+ on root
systems and their nutrient exchange, the nutrient leaching (of Ca, Mg, K, Na) that results from
acidification is known to make forest trees more sensitive to disease and stress (Driscoll et al.
2001; Rabalais 2002). Calcium is an essential nutrient to animal life as well. Acidification can
lead to the dissolution of calcium carbonate shells leading to reductions in terrestrial and aquatic
snails and other molluscs. In Europe, snail decline from acid deposition is thought to cause thin
egg shells from resulting calcium deficiencies in nesting birds who consume them (Graveland et
al. 1994). Aquatic fauna are also sensitive to increased H+ in the water column. Molluscs such as
snails and clams are particularly sensitive to reductions in water pH (Environment Canada 2002)
because their calcium carbonate shells are sensitive to dissolution from reactions with acidic
protons. The biological impacts of acidification in both aquatic and terrestrial environments are
diverse and like eutrophication can lead to changes in the biodiversity, especially in pHdependent systems (Findlay 2003). However, since the focus of this report is RN deposition, and
there is not yet any evidence to support that Canadian ecosystems are nitrogen saturated or that
RN is contributing to acidification, the reader is referred to the existing literature, including
CCME publications, on the biological effects of acidification.
3.3 RN Output and Sinks
Nitrogen saturation can also be instigated by natural phenomena (Hessen et al. 1997b). For
30
example: RN retention is lower in old forest stands with decreasing N demand; defoliation by
insects has been reported to increase NO 3 - leaching from several watersheds in the United States
(Dow and Dewalle, 1997; Eshleman et al. 1998); and the leaching of N and other nutrients occurs
immediately following forest fires. On a more human level, clear-cut forest harvesting is known
to generate nitrogen leaching up to 4 years after cutting (Wiklander et al. 1991; Pardo et al.,
1995). All in all land use and environmental histories are major factors influencing N-cycling and
retention (Aber et al. 1997).
In order for RN to not be leached, RN must be retained – i.e. taken up as biomass by plants,
animals or micro-organisms. Until these organisms die and the organic nitrogen in their biomass
mineralised and returned to reactive forms, the organisms and their associated biomass act as a
store or sink of ecosystem nitrogen. Although nitrogen retention in Canadian ecosystems seems
higher than that found in European ecosystems (Kaste and Dillon 2003) there appears to be a
deposition threshold at which retention begins to fail and leaching occurs. Uptake and retention
of RN by soil and vegetation is – like most RN effects – determined by species and climatic or
chemical variables. For instance in the forests of Denmark it was found that RN uptake and
retention increased with temperature and was higher in deciduous than coniferous forests
(Gunderson et al. 2009). Similarly, Aber (1998) found that N-leaching was higher from
coniferous versus deciduous forests at comparable latitudes and climates. Rates of leaching and
retention are governed by a myriad of factors including moisture, growth rate and microbial soil
composition.
In eastern Canada, Watmough et al. (2005) found the majority of RN in deposition to be retained
in catchments receiving < 7 kg ha-1 per year bulk inorganic RN deposition. Less conservative
estimates of have found nitrogen saturation to occur at annual NO 3 - depositions exceeding 10
kg/ha/yr (Jeffries 1995; Dise and Wright, 1995; Sullivan et al. 1997) and 15 kg /ha/yr (Grennfelt
and Hultberg 1986); levels at which subsequent nitrate leaching through the soil profile and into
surface water may be expected. Studies have shown that less than 50% of atmospheric RN may
be retained by alpine tundra due to low biomass accumulation rates and short growing seasons
(Burns 2004).
Forest floors store (organic) nitrogen in the form of undecomposed organic matter. Forest plants
store nitrogen as biomass. When leaves are shed or plants die, decomposition and subsequent
mineralization return nitrogen to the soil in reactive form. While trees and most plants contain
little N in comparison to C or H, animal flesh is a good sink of RN due to its high protein and
amino acid content. While not discussed in detail here, the import and export of domesticated
animals for consumption can have a dramatic influence on national or local N budgets (Burke et
al. 2009). High nitrate or protein feed for livestock increases their protein content. If they are
exported, so is the associated nitrogen from their feed (e.g. soy). About 20% of the RN applied as
fertiliser to feed crops is taken up as animal protein (UNEP 2007). In contrast when meat is
imported, the protein is made of foreign nitrogen, and following consumption and digestion, the
associated RN that is not incorporated into human biomass, ends up in domestic municipal waste.
Municipal waste is a main source of RN to land and water in Canada (Chambers et al. 2001;
Environment Canada 2010).
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3.4 Integrating Eutrophication, Acidification Effects
In aquatic and terrestrial ecosystems the final result of perpetual N enrichment (eutrophication),
is water and soil acidification. Eutrophication occurs before acidification, and is the consequence
of stimulating plant and microbial growth by addition of growth-limiting nutrients (e.g., N and
P). Nitrogen saturation is reached when more N is available to plants and microbes than can be
utilised or stored, causing excess mobile nitrate to leach from soils into surface and ground
waters (Dupont et al. 2000). Eutrophication can be viewed as the entire period of enrichment
before complete N saturation occurs (Ojima and Baron 1999). This is followed by acidification.
For an N-limited terrestrial ecosystem, atmospheric RN deposition is much like synthetic
fertiliser is to agricultural crops, and results in increased plant growth, biomass accumulation and
ecosystem productivity (Tamm et al. 1995, Wilson and Emmett 1999). The ‘positive’ effect is
sustained as long as the demand for RN by vegetation is greater than the supply of soil available
RN. However, if there is excess RN being deposited, and only part of the N can be taken up by
vegetation and retained, RN begins to leach from the soil. In the from of NO 3 - this nitrogen takes
base cations with it, leading to water eutrophication, soil (and eventually water) acidification and
varied effects from ecosystem degradation and decline of ecosystem productivity (Thimonier et
al. 1994, Erisman and de Vries 2000, Vitousek et al. 2000, Emmett and Reynolds 2009) to shifts
in species composition (Bobbink et al. 2001; DeVries et al. 2007). But at what stage are the
effects of excess nitrogen “undesirable”? Maybe it’s desirable for tree-growth to be enhanced, but
not as desirable to increase the growth of noxious weeds or invasive species. Perhaps effects do
not become undesirable until surface water becomes enriched and experiences enhanced algal
growth. The answers to this question will depend on whom you ask, but will also differ by
location, ecosystem and local management strategy.
Terrestrial ecosystems that are susceptible to eutrophication are those with low RN demand, high
atmospheric RN deposition, and/or low buffering capacity (or ANC). Such ecosystems include
high-elevation alpine or subalpine ecosystems and tundra (Bobbink and Roelofs 2005), that are
slow-growing and have low net primary productivity; and nitrogen limited prairie grasslands
(Köchy and Wilson 2001; Stevens et al. 2009). Areas with high nitrogen deposition and a high
sensitivity to RN excess are most at risk of eutrophication such as northern peatlands (tundra);
reported by Turetsky et al. (2003) to have very high N deposition rates in Manitoba (1620
eq/ha/yr) and Alberta (135-7980 eq/ha/yr). While these relatively sensitive northern ecosystems
may be expected to reach N-saturation more easily due to low uptake, their risk of acidification
still depends on the buffering capacity of underlying mineralogy.
The effects of increasing RN deposition on forests can be illustrated schematically after
Gundersen (1999), cited in Kennedy (2003) (Figure 4). Initially atmospheric RN deposition is
used for vegetative growth. As deposited RN increases, changes in species composition of nonwoody herbaceous ground flora begin, due their short life cycles in comparison to trees.
Nitrophiles (N loving species) out-compete plants with lower N requirements, thus N deposition
can impact on biodiversity. Further increases in RN deposition can result in imbalances in tree
nutrition and forest growth may be detrimentally affected. Finally if RN deposition is perpetual
and reaches levels at which inputs to the forest ecosystem are in excess of both biological
demand and soil storage capacity, the nitrate will begin to “leak” (or leach) from the soil below
the rooting zone (DuPont et al. 2000). This state, which has been termed “N saturation” (Aber et
32
al. 1989) has two key detrimental effects: 1) NO 3 - leakage disrupts the ion balance in the soil
causing soil acidification, base cation depletion and increased aluminium (Al3+) toxicity to roots;
and 2) the migration of NO 3 - from the soil profile may enter surface and groundwater supplies,
resulting in problems of aquatic eutrophication and algal growth.
Figure 4. Conceptualisation of the effects of increased reactive nitrogen deposition on forest ecosystems
adapted from Gundersen (1999), cited in Kennedy (2003). The green forest growth curve represents
positive growth, rather than impact, in the y-direction.
The onset of change in forest ground vegetation composition may overlap with forest growth
improvements in times of RN enrichment. Similarly imbalances in forest tree nutrition continue
to occur as soil becomes acidified (Figure 4). These overlaps of forest response states have
important implications for the interpretation of national eutrophication sensitivity maps that seek
to identify the threshold of pollutant input that specific ecosystem elements can withstand. From
a management perspective this presents the need to prioritise the species, structures or processes
that we are trying to protect. Soil acidification comes as the last stage, and as a result of deep N
saturation and leaching from the soil matrix (Figure 4). Forest growth begins to decline as a result
of nutrient imbalances and continues to decline further as a result of acidification and
corresponding base cation leaching and toxic Al mobilisation.
The states of eutrophication and acidification are intricately related and bound by a common
element – nitrogen (in its reactive forms). Therefore in the science and management of
eutrophication it is necessary to include (at least some of) the science and management of
acidification as well. Since acidification can be an outcome of the atmospheric emission and
subsequent deposition of both reactive nitrogen and sulphur compounds, emissions reductions
programs have often approached both NO x and SO x in combination. Often acidification
33
objectives can be achieved by limiting the emissions of either NO x or SO x ; however most often,
reductions of both atmospheric pollutants are required (Figure 5).
Figure 5. Plot of N and S emissions and reductions required to meet priorities. Source: UNECE-CLRTAP
(2004).
The strategies for managing eutrophication inevitably need to include strategies for the control of
environmental P additions that eventually end up in waterways alleviating the typical P-limitation
associated with excess growth in freshwater lakes. In addition to interactions and synergies
amongst atmospheric constituents and plant nutrients, there are interactions between effects and
affected organisms – some of which were discussed in section 3.1.4. To complicate matters
further, the rate of compound influx and the complexity of changes over time, calls for an
understanding of past, present and future processes and response in Canadian ecosystems.
The concept and practice of Cumulative Effects Assessment (CEA) have evolved in order to
tackle similarly complex environmental problems or projects. While Cumulative Effects
Assessments in Canada occur on a project-by-project basis, the concept in principle deals with
the effects of multiple activities in shared space and time (Hegmann et al. 1999). Effects and their
interactions are made manageable through the use of: Valued Ecosystem Components (VEC) – a
sensitive biological receptor in need of protection; ecological indicators – measurable symptoms
of “undesirable” change; scoping – bringing an assessment down to a manageable size such as an
airshed; and prediction – deriving the risk and probability of future outcomes.
In 2009 CCME designed an initiative called Regional Strategic Environmental Assessment (RSEA) that is designed to combine cumulative effects assessment with strategic effects assessment
to arrive at desirable, rather than inevitable, future outcomes. The full CCME document on RSEA is available from: http://www.ccme.ca/ourwork/environment.html?category_id=135. In
addition, the Cumulative Environmental Management Association (CEMA) based in Fort
McMurray Alberta, includes a NO x SO 2 Management Working Group that, while focused on the
acidification and ground-level O 3 elements of RN emissions, does subscribe to many of the
elements of CEA in their approach to RN (see http://www.cemaonline.ca/index.php/workinggroups/nsmwg).
34
Since there is currently no widely accepted community of practice for CEA, and since the
concepts of CEA are discussed thoroughly elsewhere (including CCME publications), a full
discussion of the theory and practice of CEA is not given here. However, because CEA is
evolving as a useful way to approach complex environmental problems; the use of VEC,
indicators, scoping and prediction are discussed in the following with reference to their utility in
RN and eutrophication assessment and management.
4.0 Tools for Monitoring and Management in Canada
This section gives a critical overview of the current tools and programs for assessing RN
deposition effects in Canada as of early in 2010. There have been, and continue to be, a number
of efforts across Canada that attempt to quantify RN deposition, assess ecosystem sensitivity to
RN deposition, and to quantify the deposition loads at which undesirable effects may occur.
These efforts, their strengths, and their weaknesses, are discussed in the following. It includes
both tools that are currently utilised in Canada, and those used elsewhere that may have particular
relevance to Canadian ecosystems. The overview includes a critical analysis of approaches such
that recommendations may be made in section 5.0. While there are a myriad of ways to approach
eutrophication science and assessment, particular methods stand out in the literature. Most of
these techniques have been developed in Europe and were brought to Canada, when ecosystems
here began to experience symptoms of deposition similar to those experienced in Europe. Some
of the eutrophication efforts overlap with the critical loads efforts. However these two
ecosystems states are still approached separately in terms of tools and mapping and much
progress is to be made towards a truly integrated assessment of the ecosystem effects from
atmospheric RN deposition.
In order to assess the vulnerability of ecosystems to either eutrophication or acidification an
approach called “critical loads” (CL) is often used, alone or in combination with other tools.
Developed in Europe, a critical load is defined as “a quantitative assessment of one or more
pollutants below which significant harmful effects on specified sensitive elements do not occur
according to present knowledge” (Nilsson and Grennfelt 1988). This measure is similar to that of
“critical levels” used as thresholds for concentrations of various pollutants. For example,
Canada’s National Ambient Air Quality Objectives (CCME 1999) are a type of critical level, but
do not apply to deposition (fluxes) of pollutants. CL for nutrient N consider the flux of RN (in
kg/ha/yr) as a balance between N input and output; whereas CL for acidic deposition also account
for the atmospheric deposition of chloride (Cl-) and base cations, and the release of base cations
from soil weathering. Both CL for nutrient nitrogen (i.e. N eutrophication) and acidification are
compared with deposition estimates or measurements to assess ecosystem vulnerability to
undesirable change via one of these effect pathways. These measures are termed Critical Load
Exceedance (CL(EX)) expressed as the difference between the actual deposition and CL or:
CL(EX) = Deposition – CL
where deposition (in this case) is RN in eq/ha/yr. A similar expression is the deposition index
(Di) where:
Di = CL – Deposition
35
which is used to identify sensitive forest ecosystems as a measure of the extent of nutrient
depletion in forest soils (Forest Mapping Working Group 2003). Average Accumulated
Exceedance (AAE) may also be calculated as:
AAE = (CL(EX) x area exceeded)/total area of assessment
thus giving a measure of the magnitude of exceedance and giving relevance to effects on
biodiversity (Jones et al. 2009). The exceedance for eutrophication CL is determined using the
atmospheric deposition of N only, whereas the acidification CL(EX) considers the deposition of
both S and N. Calculations of CL for both acidification and eutrophication are discussed in
sections 4.2 and 4.3, respectively. Ways of quantifying deposition for use in CL(EX) calculations
and other strategies, are given below.
4.1 Tools for Quantifying Deposition
While atmospheric emissions of acidifying sulphur have been significantly reduced over the past
decades in Europe, the U.S. and Canada, nitrogen emissions continue to rise causing increased
atmospheric RN deposition (Erisman and de Vries 2000, Vitousek et al. 2000, Seip and Menz
2002, Emmett 2006, Gundersen 2006). In order to understand the effect that this increased RN
may have on ecosystem, we must first quantify the amount of RN that is reaching the ecosystem
through atmospheric pathways. Direct measurements are always the best way to quantify
anything. However, when the required measurements are numerous – as is true for a country the
size of Canada, and the measurement equipment expensive – as is true for devices that measure
RN deposition, comprehensive measurement programs are not always feasible. In this case,
measured data may need to be supplemented with data provided by mathematical models.
4.1.1 Measured Deposition
Deposition can be measured by a variety of means. Because the deposition of RN can occur in a
variety of physical (wet, dry, gaseous and particulate), and chemical (NO x , HNO 3 , NH 3 , etc.)
forms (Table 2), it is often necessary to employ multiple methodologies for measuring total, or
bulk RN deposition.
Wet deposition is relatively simple to measure. While occult precipitation may be neglected by
most sampling systems, the collection of rainwater and snowmelt, which is subsequently
analysed in a laboratory, provides valuable quantification of the wet deposition of dissolved
gaseous- and particulate-RN species in precipitation. The Canadian Air and Precipitation
Monitoring Network (CAPMoN) measures atmospheric (wet) deposition of pollutants in
rainwater and snowmelt (Figure 6). Measurements of wet deposition can severely underestimate
total or bulk deposition in drier climates (Fenn et al. 2009) – such as Canada’s north. Therefore
some CAPMoN sites also measure particulate and gaseous compounds including RN. Dry
deposition however, can not be measured directly and must be inferred, or calculated, from
concentration measurements and surface or atmospheric characteristics (i.e. deposition velocity).
36
Figure 6. CAPMoN stations and what they measured in 2009. “Air” refers to filter-pack measurements of
daily gaseous HNO 3 ; and daily particulate NH 4 + and NO 3 -. Source: Robert Vet Personal Communication
February 2010.
Data collected by CAPMoN are stored in the National Atmospheric Chemistry (NAtChem)
Database (Vet et al. 2005; NAtChem 2007). As of 2009 CAPMoN sites included 15 sites making
daily measurements of nitric acid (gas), ammonium and nitrate (particles), using filter pack
collection; precipitation collection with analysis by ion chromatography at 27 sites, and
continuous (hourly) gaseous measurements of NO x , NO and NO 2 at 3 sites. These 3 sites were
located in Kejimkujk, Nova Scotia; Egbert, Ontario; and Saturna Island, British Columbia
(Figure 6). Measured deposition data of both particles and gases from the CAPMoN atmospheric
monitoring network are often used in conjunction with air quality modelling for model
calibration, performance evaluation and sensitivity analysis.
CAPMoN precipitation monitoring sites are supplemented by provincially operated programs
such as the Newfoundland Environment Precipitation Monitoring Network (NLPMN), or the
British Columbia Precipitation Chemistry Sampling Network (BCPCSN). While Prince Edward
Island, and the Northwest, Yukon and Nunavut Territories do not have independent precipitation
monitoring programs (Figure 7); Saskatchewan initiated their own atmospheric chemistry
monitoring program in the summer of 2009.
37
Figure 7. Map of precipitation monitoring stations (circles) in Canada past (white) and present (black) –
as of June 2006. CAPMoN sites, of particular relevance to RN deposition, are represented by red outlines.
Source: Canadian National Atmospheric Chemistry Precipitation Database (2006).
The NAtChem database is a compilation of Particulate Matter, Precipitation Chemistry, and
Toxics monitoring data from a total of 1140 monitoring stations comprising 23 past and present
monitoring networks in both Canada and the United States (including CAPMoN, NLPMN and
BCPCSN). Particulate matter and gaseous pollutant concentrations are collected as part of the
National Air Pollution Surveillance Network (NAPS). A total of 432 Canadian precipitation
monitoring sites have existed in Canada at some time (Figure 7), and as of 2006, there were 119
sites remaining (Canadian National Atmospheric Chemistry Precipitation Database 2006). More
information on NAtChem including data downloads and mapping utilities can be found on the
NAtChem website (http://www.msc.ec.gc.ca/natchem). Although most of the stations in the
database are no longer recording measurements, these data provide useful baseline conditions,
especially in regions where high deposition is a relatively new phenomenon (such as western
Canada).
Ion Exchange Resin (IER) columns can be used for the measurements of wet deposition. The
columns are made of piping lined with a resin topped with a funnel that collects either rain or
throughfall. IER columns are particularly useful in the measurement of throughfall to the forest
38
floor, representing the deposition flux directly to the soil (Fenn et al. 2009). The resin, like
precipitation water, is analysed in a laboratory for compounds of interest using the appropriate
technique. These columns have been used to successfully measure RN deposition in the United
States by the US Forest Service (Fenn et al. 2008; Fenn et al. 2009), European ICP forest
monitoring plots (Bleeker et al. 2003) and more recently in Canada, as part of an extensive
monitoring program in Alberta’s oilsands region (Mark Fenn Personal Communication February
2010). A similar technology but with a slightly different design, the resin bag, has also been used
in Alberta to measure NO 3 -, NO 2 - and NH 4 + deposition at the soil surface (Köchy and Wilson
2005). In theory, the resin bag, can absorbs gas, particles and wet aerosols in a manner similar to
that of a leaf surface and therefore registers total (bulk deposition) (Köchy and Wilson 2005), but
there is some uncertainty as to the dry depositon fluxes that these types of monitors are able
absorb.
Dry deposition fluxes can be estimated using numerous techniques. Passive samplers use a filter
coated with a chemical that reacts with nitrogen species (NO 2 , HNO 3 ) when they deposit on its
surface (Krupa and Legge 2000). The filters are contained within a small body made of an inert
material such as Teflon and usually employ a rain shield. These samplers, like rainwater and IER,
measure a cumulative dose over the period of measurement and are analysed in a laboratory using
the appropriate technique (such as chromatography or colorimetry) for compounds of interest
(particularly NO 3 - and NH 4 +). Other techniques for dry deposition measurements include the
analysis of surface soil and branches (branch washing) directly; however these techniques require
dry conditions to avoid wash-off and leaching and are not suitable for locations without a welldefined wet season (Fenn et al. 2009). Surrogate surface approaches fall somewhere between a
passive sampler and resin tubes. They measure deposition to a “surrogate” open water or soil
surface using filters in combination with greased disks as collection surfaces (see example in
Raymond et al. 2004) and have been used successfully in Canada in the measurement of
deposition velocities and fluxes of polycyclic aromatic hydrocarbons (St-Amand et al. 2009)
Both passive sampling and resin tubes are inexpensive when compared to active monitors and
flux towers and neither require an external power source. They do however require sophisticated
laboratory analysis of samples once the exposure period ends. The location of sampling is very
important, including placement within a vegetation canopy. This is due to different levels of flux
and uptake by vegetation and soil, as well as the turbulent or roughness characteristics of various
surfaces. Flux towers that employ a numerical analysis technique called “eddy correlation” are
traditionally used to measure vertical variations in gas fluxes of vegetation canopies (Edwards
and Ogram 1986). However, the use of vertical passive samplers has shown (from ozone research
in southern BC) to be a promising and inexpensive alternative (Krzyzanowski 2004). The vertical
placement of passive samplers has shown to bear greatly on measurements of RN deposition in
Alberta (Bytnerowicz et al. 2010). The effects of edges and exposure are also important
considerations in any study of RN deposition or concentration. Deposition fluxes using passive
samplers are calculated based on the characteristics and leaf surface area of surrounding
vegetation (Fenn et al. 2009).
All of the above methods, ion exchange resins, passive samplers and lichen analysis, measure
cumulative RN exposure and are therefore not capable of discerning between individual
deposition events. The continuous monitoring of air pollutants in Canada is coordinated by
NAPS. The network began in 1974 and now includes Provincial and Federal stations set up in all
39
13 Provinces and Territories (Figure 8). Stations are usually located in urban environments and
measure ambient concentrations of NO 2 , SO 2 , O 3 , VOC, PM, providing data for larger-scale
modelling and research projects (see: http://www.etc-cte.ec.gc.ca/NAPS/index_e_html). NAPS
sites measure RN as NO 2 only (Figure 8) and do not take into account contributions of other RN
species such as HNO 3 or NH 3 to total bulk deposition. Nonetheless, deposition fluxes can be
calculated inferentially using continuous monitoring of RN concentrations to determine the
timing and magnitude of deposition events. However, these calculations are often inaccurate due
to difficulties in estimation the actual deposition velocity and related surface characteristics such
as leaf area index (Fenn et al. 2009). The use of flux towers, mentioned above, can be used to
refine some of these inferred deposition estimates but are labour and cost intensive. Atmospheric
models also provide estimates of the occurrence and magnitude of deposition events.
Relative to other parts of western Canada, the Georgia Basin in southwestern British Columbia
has been studied extensively for N deposition and its effects. Wet deposition data collected in the
late 1970s by Krumlik (1978) and Feller and Kimmins (1979) compared with NAtChem data
from the 1990s shows that RN deposition has increased considerably since the early studies.
Environment Canada’s National Agri-Environmental Standard Initiative (NAESI), a
four-year (2004-2008) project that among other things, included monitoring of NH 3 emissions in
the Lower Fraser Valley (part of British Columbia’s Georgia Basin), reported that annual mean
NH 3 concentration was 20.4 µg/m3 at a site near a poultry barn (equivalent to 4.2 kg/ha/yr) due to
chicken husbandry, the application of poultry manure to fields, a dairy cattle operation, and to a
lesser extent, vehicular exhaust emissions of NH 3 in the Lower Fraser Valley (i.e. Greater
Vancouver) (Environment Canada 2006a). It was also found that NH 3 concentrations within the
area may saturate the Fraser Valley airshed during meteorologically stagnant periods when the
wind speed is less than 8 m/sec (Environment Canada 2006b). Although there is little evidence of
coastal eutrophication in Canada, the Georgia Basin with high RN inputs, is the most at potential
risk of eutrophication of any Canadian coastal area (Jeffries et al. 1998). This is exacerbated by
RN inputs from increased marine aquaculture in recent years
In Alberta, N deposition has increased dramatically in the oilsands region and the EdmontonCalgary Corridor since 1990s due to N emissions from intensified industrial activities of
upstream oil and gas extraction and refining (CEMA 2008). Köchy and Wilson (2001) reported
that bulk N deposition (wet + dry) was 8 kg/ha/yr in Jasper area, but increased up to 22 kg/ha/yr
on the Elk Island, relevant to nutrient nitrogen CL discussed in the next section and elsewhere
(e.g. Bobbink et al. 2002). This high bulk deposition implies a potential risk of eutrophication.
However, throughout less industrialised parts of Alberta, N deposition remains low, and in
southern Alberta total nitrate (NO 3 ) deposition was about 4 kg/ha/yr (Schreier et al. 1999).
40
Figure 8. Location of NAPS Network sites that measure concentrations of NO 2 . Source: National Air
Pollution Surveillance Monitoring Program, Environment Canada, Cèline Audette, Personal
Communication March 2010.
Total N deposition measured from 1994-1998 was 211 mol c /ha/yr in Manitoba, and 163
mol c /ha/yr in Saskatchewan (Aherne and Watmough 2006), equivalent to 2.95 and 2.28 kgN/ha/yr, respectively). Aherne et al. (2004) also reported that N deposition was less than 2
kg/ha/yr for most of areas in Manitoba, and below 1.5 kg/ha/yr in the western prairies. Similarly,
Miller (2006) reported that annual inorganic N deposition in Atlantic Provinces was between 2.8
and 13.8 kg/ha/year, of which 41-67% was nitrate (NO 3 -). Deposition levels of total N in
southwestern Quebec and southern Ontario may reach 8 kg /ha/yr at CAPMoN sites, vary with
year and are lower in the rest of Canada (Vet et al. 2005). In the great Lakes region of southern
Ontario Schindler et al. (2006) found that RN deposition reached levels of 12-15 kg/ha/yr near
the border, > 9 kg/ha/yr in southern Ontario and Quebec, and < 6 kg/ha/yr just east of the
Manitoba-Ontario boarder. NH 3 deposition was highest in western Ontario at 5-6 kg/ha/yr,
compared to 2.5 kg/ha/yr in other areas. It was found that dry deposition accounted for 17-41% of
total N.
Based on the CAPMoN wet nitrate deposition data collected from 1995-2005 (Figure 9), it was
found that the highest, wet NO 3 deposition was in southern Quebec, south-central Ontario, and
southern Lake Erie and Lake Ontario reaching annual means of 15-20 kg/ha/yr (over 2000-2004),
41
and totals of 12-15 kg/ha/yr in 2005, but decreasing in general through the 1990-2005 period. All
regions of highest deposition are clustered around large urban centres (Canadian National
Atmospheric Chemistry Precipitation Database 2007), however this is also where RN deposition
is most actively measured. Nitrate (NO 3 ) has been found to account for 41-67% of total N
deposition in 21 forested catchments in Canada, the US and Europe from 1990-1999 (Watmough
et al. 2005). Although less monitoring of N deposition has been carried out in Canada’s northern
regions, recent research shows atmospheric N increases in the globe’s arctic regions including
Canada’s Alert bay that are not necessarily of local emission influence, but relate to variable
long-range winter transport (Hole et al. 2009).
a)
b)
c)
d)
Figure 9. Changes in the spatial distribution of measured wet nitrate deposition in a) 1990, b) 1995, c)
2000, d) 2005. Deep green is ≤ 5kg/ha/yr, whereas red is >30 kg/ha/yr. Source: International Joint
Commission (2008), with data from NAtChem (Canadian National Atmospheric Chemistry Precipitation
Database (2007)) using measurements from CAPMoN and NADP (National Atmospheric Deposition
Program of the US).
4.1.2 Modelled Deposition
The first step to using any atmospheric model to estimate deposition is the development or
retrieval of an atmospheric emissions inventory for the area of interest. Emissions inventories act
as a tool for both science and policy. As the main input into regional-scale air pollution models,
emissions inventories allow us to estimate what the potential impacts of air pollution may be.
However, many have noted inaccuracies with Canada’s emission inventory data (NARSTO 2005;
42
Zhang et al. 2005; Krzyzanowski 2009) and there are no federal strategies to improve that
accuracy. In addition, Canadian reporting policy offers exemptions to many industrial sectors
based on activity or emission thresholds. Without the inclusion of all emission sources, it is
unlikely that modelled deposition will reflect actual deposition. This is important because models
are used to fill in the spatial (and sometimes temporal) gaps in deposition measurements that
exist out of logistic necessity. You can’t directly measure everywhere, or measure the future.
Despite discrepancies between emission inventories and the real world, in areas with large
emission sources and more comprehensive inventories, the modelled pollution forecasts they
produce can be an invaluable source of information.
A Unified Regional Air-quality Modelling System (AURAMS) was developed by Environment
Canada for air quality issues of Canadian interest (Zhang et al. 2002). The model AURAMS is
used by Canadian scientists and managers to estimate the concentration and deposition of
pollutants at regional and national scales. This large scale regional modelling is especially useful
in estimating the fate of pollutants in regions that lack monitoring stations. Regional models like
AURAMS are also capable of producing predictions of future conditions based on different
emission scenarios (e.g. emissions growth or caps) using known atmospheric chemistry and
physical parameters. These types of models are crucial in allowing us to make predictions about
future emission changes and their outcome in terms of human and ecosystem health.
Regional models generally include 3 parts: a meteorology portion, an emissions processor, and
the chemical transport model itself. AURAMS simulates the chemical transformation, transport
and deposition of various pollutants. Currently, AURAMS is used in tropospheric O 3 and smog
forecasts as well as in the transport and deposition of 22 different RN compounds (Moran et al.
2009). Moran et al. (2008) used AURAMS to estimate or predict total nitrogen deposition for the
year of 2002 in Canada at a 42 km resolution (Figure 10). Results of their modelling showed
atmospheric deposition rates of 2-28 and 0.1-5.0 kg/ha/yr for NO 3 and NH 4 , respectively.
AURAMS has also been used in Alberta at a finer (3 km) resolution over a much smaller domain
size as part of the PrAIRie2005 Field Study (Makar 2008). Both studies report the model
performing well in comparison to measured data.
The Community Multiscale Air Quality Model (CMAQ) was developed by the US-EPA and is
also used in Canada to estimate regional pollutant concentrations and deposition. CMAQ uses
similar physics and chemistry as AURAMS in its predictions of both concentration and
deposition. CMAQ has been used to predict pollutant concentrations in the Georgia Basin of
southern British Columbia (Delle Monache et al. 2007) and across Canada (Aherne 2008a).
Annual deposition estimates are not typically an output from user-defined CMAQ fields.
However, both CMAQ and AURAMS have the capability to predict total annual deposition when
utilising a full year of meteorological data. Both CMAQ and AURAMS are data, skill and
resource intensive requiring powerful computers and skilled programmers. These two models
have been shown to predict Canadian concentrations of O 3 and PM with comparable accuracy
(Smyth et al. 2009).
43
Figure 10. AURAMS simulated total N deposition across (western) Canada using year 2002 emissions.
Source: Mike Moran, received from Julian Aherne Personal Communication (January 2010).
Less complex ‘refined’ models such as AERMOD and CALPUFF also estimate deposition but do
so on a smaller scale and do not incorporate pollutant chemistry in deposition estimates. These
models are less resource and skill intensive, and some have user interface software (see
http://www.weblakes.com). Primarily used by industry for emissions permitting requirements and
the prediction of local air concentrations, these models recently became endowed with algorithms
for predicting the dry and wet deposition of chemical species such as RN. In western Canada’s oil
and gas rich province’s of Alberta and British Columbia, RN deposition has been estimated using
CALPUFF (NO x SO 2 Management Working Group 2008) and AERMOD (Krzyzanowski 2010),
respectively. Although the accuracy of deposition estimates predicted by these models is under
question, it was found in California that CALPUFF provided the best estimates of modelled N (as
NO x and NH 3 ) deposition when compared to AERMOD, CMAQ or ISCT3 (not discussed)
(Tonnesen et al. 2007). CALPUFF is generally an optimum choice of refined models in areas of
complex terrain, however it does require more detailed inputs and user knowledge than simpler
straight-line Gaussian models such as AERMOD. The results from passive sampling data can
also be used as input into dispersion models such as AERMOD (Fenn et al. 2009).
All of these models require detailed land-use and surface information to be able to approximate
deposition fluxes. The more accurate the data that goes into the model, the more accurate the
output of the model will be. This is true for surface data, meteorological data and emissions data.
For instance using ensembles of meteorological forecasts from various weather forecast models
has been shown to increase CMAQ’s predictive ability in the Georgia Basin of British Columbia
(Delle Monache et al. 2007). Similarly, ensemble modelling of global pollution using multiple
dispersion and circulation models has shown to predict transport with increased accuracy
(Dentener et al. 2006).
There is generally a high uncertainty associated with all models and the data they use. For this
reason modelling alone should not be the basis of assessment. For instance here is high
uncertainty in modelling the dry deposition of NH 3 . Volatilization rates for NH 3 vary widely
44
from zero to 15% in Europe, up to 25% in tropical areas compared to 4% in N-P-K mixed
fertilisers and 2% in Ca+NH 4 +NO 3 blended fertilisers (Schjorring 2008). These types of
unknown variation add large amounts of error to modelling predictions. In addition, the
calculation of deposition velocity in the conversion of concentration estimates to deposition
estimates by the model, are riddled with simplifications, assumptions and unknowns.
Despite the uncertainty associated with atmospheric dispersion models, unlike the passive and
cumulative sampling techniques referred to in the previous section, models are able to discern
particular deposition events – assuming that the data includes some temporal allocation of
emission. Although the exact location or magnitude may be misrepresented, the ability to make
deposition estimates seamlessly across a continuous landscape provides and invaluable tool. The
inaccuracies associated with atmospheric dispersion models are discussed in detail in the models’
manuals and elsewhere by users.
4.1.3 Other tools and indicators
Lichens are proving to be useful tools for the measurements of bulk deposition on a relative basis
(i.e. high, medium and low) in the oilsands region of northern Alberta. Because they have no
stomata to regulate uptake, the total N in lichen tissue relates well to the RN deposited
atmospherically in its environment. Lichens are being used in combination with other monitoring
techniques in northern Alberta to monitor change and deposition related to atmospheric RN; and
are faring well in their comparisons with RN deposition measured using IER and passive
sampling techniques (Shanti Berryman Personal Communication, February, 2010). In this way,
lichens can be collected in remote areas and their tissue concentrations of RN analysed. Nitrogen
isotope studies show promise for the speciation of various RN compounds in lichen tissue.
Lichen species that have proven to be especially useful in the foliar N measurements include the
epiphytic Hypogymnia physodes and Envernia mesomorpha that are fairly pollution tolerant and
therefore apparent in both high and low deposition environments (Shanti Berryman Personal
Communication February, 2010). Attempts to relate lichen tissue from the species Platismatia
glauca, Parmelia sulcata and Hypogymnia physodes to N deposition in the Georgia Basin has
thus far proven unsuccessful; however moss and lichen N content related well with CMAQ
estimated deposition fields (Raymond et al. 2010).
In 2004 the National Ecological Monitoring and Assessment Network EMAN founded a
terrestrial biodiversity program that uses lichen abundance and biodiversity as indicators of air
pollution concentrations in and around the city of Hamilton. In 2005 this program was extended
to parts of Nova Scotia. More information on the lichen monitoring project and lichen monitoring
protocols for Canada’s mixed, boreal and west coast forests can be found on the EMAN website
(http://www.eman-rese.ca/eman/ectools/protocols/terrestrial). Lichens are particularly useful
indicators of air pollution and can therefore provide an early warning of ecosystem change. More
information on the use of lichens as measures and indicators of air pollution can be found in
Geiser and Neitlich (2007) and Thormann (2006).
Measurements of foliar nitrogen in various native plant and tree species can be used to examine
relative doses and areas of concern (Innes 1995; Piticairn et al. 2001) in a similar manner to total
N measured in lichens. This technique can be particularly valuable in areas lacking larger
epiphytic lichens, or in more urbanised environments with low lichen biomass. Nitrogen content
45
in the bark of hardwood trees, particularly silver maple (Acer saccharinum) is also showing
promise as an indicator of atmospheric nitrogen exposure within the ecological range of this
species. A new study in southern Ontario being conducted by researchers from Brock University
is examining this technique as a tool for air quality assessment and management (Daniel
McCarthy Personal Communication April 2010).
4.2 Nitrogen Loading
Ecosystems respond to RN releases to land and water in addition to atmospheric releases, and
therefore it is necessary to consider RN inputs from as many sources as possible. Nitrogen
loading is the sum of atmospheric RN deposition + RN released from non-atmospheric
anthropogenic sources. Total nitrogen loading includes sources from sewage and fertilisers that
substantially increase the RN available to a system and play an important role in nitrogen
enrichment. Total nitrogen loading to a system needs to be known before appropriate estimates of
eutrophication and acidification risk can be made. The most recent total N-loading data for all of
Canada is from 1996 (Chambers et al. 2001), however a combination of data from the NPRI
(Table 4) and other sources such as provincial surface water quality monitoring network data are
used in the following discussion.
Table 4. National 2008 NPRI facility reported RN (NO x and NH 3 ) releases to land, water and air (tonnes)
in Canada from Environment Canada (2010).
Compound/Pathway
a
NO x
NH 3
b
Total RN
Land
867.3
296.7
508.1
Water
61,955.9
48,269.4
58,561.6
Air
765,399.0
21,158.0
250,427.1
Total
828,222.2
69,724.1
309,496.8
NO x refers to NO 3 - in solution at pH>6.0 when referring to land and water releases, and NO+NO 2 expressed in
NO 2 unit equivalents when representing emissions to air.
b
Calculated using molar mass ratios from section 2.2 of 0.822 for NH 3 and 0.304 for NO 2 . This gives RN as
elemental N, which can then be compared or summed across reactive forms.
a
Over 300,000 tonnes of N entered Canadian soils, surface water and air, as either reactive NO x or
NH 3 , from industrial and municipal point sources in 2008 (Table 4). However, these values are
limited to facility waste and therefore do not include agricultural fertiliser or manure applications,
which contribute substantially to RN loads. In Canada, agricultural inputs of RN total
approximately 2.47 million tonnes N per year, about 50% of which is lost to the environment
(Niemi et al. 2009). This additional 1.2 million tonnes N per year brings the 2008 annual RN
input to the environment up to approximately 1.5 million tonnes N (in reactive forms), or five
times the NPRI reported facility releases (Table 4). Additionally, in 2005 marine aquaculture in
Canada produced 98,441 tonnes of farmed salmon (DFO 2005), which according to relationships
in Islam (2005) equates to the addition of over 45,500 tonnes of N in reactive form to marine
ecosystems, mostly on the western coast of British Columbia.
Major RN loss pathways include the volatisation of NH 3 , and NO 3 leaching (Niemi et al. 2009).
Ammonia losses from agriculture are estimated monthly and included in the federal inventory - at
46
the time of this report the 2008 Criteria Air Contaminants inventory was incomplete and facilityreported NO x emissions to air (Table 4) represent approximately one third of all atmospheric
NO x emissions once agricultural, mobile and open sources are considered (Environment Canada
2010) (see Figure 2 for 2007 totals). However, NPRI values also neglect smaller facility emission
sources due to reporting requirements and exemptions (Canada 2007). Although incomplete,
NPRI facility reported data on emissions to air, land and water, provides essential information on
RN loading to the Canadian environment.
There are no clear sources of recent and comprehensive information on nutrient loading in
Canada. Data on fertiliser sales from 1967-1997 show a 6-fold increase in nitrogen inputs nationwide over the 30-year period. Western Canada was responsible for the use of 83.7% of the
country’s 1.4 million tonnes of N in fertiliser in 1997. Phosphorus, which limits growth in
freshwater aquatic ecosystems, showed a similar 3-fold increase between 1967-1997, reaching
0.55 million tonnes in 1997 (Korol and Larivière 1998). These numbers do not include manure
applications, only chemical fertiliser by sales, which by deduction, add another 1 million tonnes
of RN to Canadian soil each year.
4.3 Tools for Eutrophication
Various stages of terrestrial ecosystem response to RN enrichment have been suggested. By
defining the progression of eutrophication in this way, a common language can be developed
across jurisdictions. This common language should be the first step in developing standard tools
of assessment. Three simple characteristic stages include: 1) N limitation, 2) N saturation and 3)
N excess ((Aber et al. 1989; Rosen et al. 1992). A more detailed approach was developed by
Stoddard (1994) that includes a more detailed method using five distinctive classifications to
categorise various stages of nitrogen enrichment including:





Stage 0: highly RN limited, low foliar N content, low net primary productivity, high
competition for NH 4 +, leaching NO 3 - is minimal, and N 2 O production is undetectable.
Stage 1: increased RN deposition, reduced N limitation on biological function, increased
N mineralisation and cycling rates, increased primary productivity.
Stage 2: net nitrification, increased saturation of RN, NO 3 - leaching is detected, and N 2 O
efflux is low.
Stage 3: RN is not limiting to plant growth, RN leaching is increased as is NO and N 2 O
production, and biological RN retention declines.
Stage 4 or final stage: leaching of NO 3 -, release of Al 3 +, and depletion of nutrient cations
(Ca2+, Mg2+, K+, Na+). This stage is characterised by limited denitrification, enhanced
microbial productivity, and imbalanced plant nutrition.
Many European countries have ecosystems in Stage 1 or Stage 2 (Aber et al. 1989, Stoddard
1994). Most Canadian ecosystems are still at Stage 0. However, some studies (e.g. Köchy and
Wilson 2005) have reported symptoms of Stage 1 and Stage 2 conditions being observed in areas
receiving comparably high RN deposition. Very few Stage 3 cases have been observed in North
America and studies show that most of North America is still in the early stages of nitrogen
saturation.
47
The complexity lies in the overlap of states (see Figure 4), and that different species and
ecosystems may respond to increased RN inputs at different times or in different ways. However
chemical, physical and biological indicators of these states can be used to help assess ecosystem
sensitivity to RN deposition effects and identify areas that are particularly sensitive to change. In
contrast, in aquatic systems - increases in dissolved N, increased algal biomass, reduced light
attenuation, increased organic matter content and reduced levels of oxygen - are some of the
ecological symptoms that would describe stages of eutrophication.
4.3.1 Critical Loads of Eutrophication (Nutrient Nitrogen)
In calculating the CL for eutrophication or nutrient nitrogen, all forms of reactive nitrogen in the
soil complex must be considered including NO 3 - and NH 4 +. The terrestrial CL for nitrogen as a
nutrient is expressed as kg/ha/yr of RN in N equivalents (i.e. using molar ratio conversions). In
general, CL for eutrophication will be less than those for acidification – in other words
eutrophication happens first. When plants take up an anion (e.g. NO 3 -) they release an anion (OH) to balance it, which neutralises any associated H+. Therefore as long as there is plant uptake of
NO 3 -, there is no net acidification (Julian Aherne Personal Communication January 2010).
Instead, there is enhanced growth and eutrophication.
The critical loads concept and tools for assessing the CL of both eutrophication and acidification
have been largely been developed and applied using European ecosystems. Critical loads of
acidity have been applied extensively in Canada and across the globe. Critical loads for nutrient
nitrogen (eutrophication) have not been applied as extensively in Canada and their relevance to
Canadian ecosystems is largely unknown. Studies in Europe and the United States have
developed N eutrophication critical loads and/or other tools to manage eutrophication (e.g.
Erisman and de Vries 2000, Bouwman et al. 2002, Cunha et al. 2002, Burns 2004, Bergstrom and
Jansson 2006); however whether these critical loads are relevant to Canadian ecosystems remains
to be tested. For the time being, we have CL for similar ecosystems that may assist in forming
thresholds in the interim.
Exceedances of critical loads and deposition indices are useful tools for eutrophication
management. Maps of critical loads and exceedances/deposition indices, which are defined and
created by incorporating all the effects from acid deposition and eutrophication, can be used to
quantify the reductions of N emissions and loadings in a given ecosystem. Maps of the critical
loads in Europe are compared with the “total potential acid deposition” for each of the grid cells,
and the exceedance (= total potential acid deposition minus critical load) of this critical load is
calculated. These CL and CL(EX) maps are coupled with descriptions of soil characteristics and
information on the sensitivity of ecosystems to aluminium, free acid, and nitrates forming the
basis of environmental treaties under auspice of the European Monitoring and Evaluation Programme
(or EMEP) (UNEP 2007).
Empirical CL are simplistic and based on a combination of previous research and expert
judgement (Table 5). They are the simplest form of the CL approach and generally apply a CL
value based on one or more ecosystem characteristics. While most of the empirical CL for
nutrient nitrogen have been derived for European ecosystems, much of the work has come from
northern regions that contain similar species and soils to parts of Canada, and therefore have
some relevance to Canadian ecosystems. However, care should be taken in their application and
48
Canadian empirical loads need to be developed or refined based on research done on Canadian
soils, waters and ecosystems.
Table 5. Empirical critical loads of nutrient nitrogen for protection of overall ecosystem health and
productivity in selected ecosystem. Adapted from Bobbink et al. (2002). The high end of the range should
be used if phosphorus is limited and the lower range if phosphorus is available.
Ecosystem
CL, kg-N/ha/yr
Temperate Forests
10-20#
Boreal Forests
10-20#
Indication of Exceedance
Northern Wet Heathland
10-20{#}
Dry Heathland
10-20##
Changes in nutrient balances, soil processes, mycorrhiza
and ground vegetation; increased parasite susceptibility
Changes in nutrient balances, soil processes, mycorrhiza
and ground vegetation; increased parasite susceptibility,
increases in free algae
Decline in mosses and lichens, decline in heathers,
transition to grasses
Lichen decline and transition from heathers to grasses
Arctic and Alpine Scrubland
5-15(#)
Decline in mosses, lichens and evergreen shrubs
High Mountain Areas
5-10#
Bryophyte and lichen effects
Tundra
5-10#
Softwater Lakes
5-10##
Changes in biomass, moss communities and decreases in
lichens
Changes in macrophyte communities
Coastal Dune Grasslands
Marine Marches
10-20#
Increases in tall grasses and N leaching
30-40(#)
Increases in productivity and succession
##: reliable; #: quite reliable; (#): expert judgment
Bobbink et al. (2002) described various empirical nitrogen deposition thresholds above which
forest function and plant communities would be invariably affected (Tables 5 and 6). For
example, forest soil processes and ground vegetation change at N depositions of 10-15
kgN/ha/yr, while heathland and scrub vegetation experience declines in lichens, mosses and
evergreen shrubs at depositions of 5-15 kgN/ha/yr. Similarly, changes in tundra biomass and
species composition are predicted at a deposition of 5-10 kg N/ha/yr (Bobbink et al. 2002).
Combinations of these individual loads are used to define a broader range that protects the
ecosystem overall (Table 5). A lack of proper experiments on the response of boreal communities
to RN deposition limits the application of this CL to boreal ecosystems. Furthermore, the
empirical CL for nutrient nitrogen is the same as that for temperate forests (Table 5). As
discussed previously, biological response is largely species specific and CL may depend on the
ecosystem elements that are most desirable to protect. For instance at a RN deposition of greater
than 18-20 kg/ha/yr Sphagnum spp. function may fail (Nordbakken et al. 2003), whereas
Racomitirum spp. of moss have shown negative responses to RN depositions as low as 10
kg/ha/yr (Jones et al. 2002). In addition, different CL levels are defined in order to protect
particular ecosystem functions (such as soil processes) or elements (mycorrhiza, lichens, ground
vegetation) (Bobbink et al 2002).
Table 6. Suggestions to use lower, middle or upper part of critical load range (Table 5) based on
ecosystem characteristics. Adapted from Achermann and Bobbink (2003).
49
Action
Temperature
Soil wetness
Frost period
Move to lower part
Cold
Dry
Long
Base cation
availability
Low
Use middle part
Intermediate
Normal
Short
Intermediate
Move to higher part
Hot
Wet
None
High
Nordin et al. (2005) suggested that a value of 6 kg/ha/yr be set as the amount where no damage to
ecosystems would occur. In the western United States, RN deposition at levels as low as 3-8
kg/ha/yr have significantly altered some aquatic and terrestrial, plant and microbial, communities
(Fenn et al. 2003a). Wet N deposition of 1.5 kg/ha/yr has been shown to cause health declines in
high-elevation ecosystems (Burns 2004). Since deposition in high-elevation Mountain Hemlock
(MH) zone near Vancouver was approximately 1 kg/ha in the 1970s (Krumlik, 1978), and RN
emissions have increased over the past few decades, there may be risks of eutrophication effects
in the Coast Mountains of British Columbia. In the Pacific Northwest of the United States, total
N deposition was 0.5-2.4 kg/ha/yr in 1992 and critical loads for N was suggested at 5 kg N/ha/yr
for the protection of aquatic resources against chronic acidification, (Eilers et al. 1994). This CL
is considerably lower than estimates often cited for forests of Europe and the northeast of the US
(e.g., 7-10 kg/ha/yr). Wright et al. (2001) pointed out that when N deposition values exceed 10
kg/ha/yr for a number of years, NO 3 - is leached into streams, a symptom of chronic acidification.
Aber et al. (2003) reported that in parts of the United States where over 10 kg/ha/yr NO 3 - is
deposited, NO 3 - is leaching from soils.
Nutrient nitrogen CL have been estimated for Canada (Aherne 2007) using empirical methods
(Bobbink et al. 2002) and their exceedances calculated by subtracting AURAMS estimated RN
deposition (Figure 10) from the empirical critical loads (Figure 11). These are the same CL nut
listed above (Table 6). The critical load was exceeded in the Lower Fraser River Valley of
British Columbia, the Calgary-Edmonton corridor of Alberta, and south-eastern Ontario (Figure
11). CL for nutrient nitrogen are also exceeded in the small areas around Fort McMurray, and
southern Manitoba – corresponding to upstream oil and gas and agriculturally dominated areas,
respectively.
These empirical critical loads (Table 5, Figure 11) may be representative of similar ecosystems in
Canada; however, research needs to either prove or disprove their relevance in the Canadian
context. Because responses to RN are known to be species specific it is likely that the empirical
CL described here, will need to be modified and updated for accurate representation of Canadian
ecosystems. Nonetheless, by comparing the empirical CL with N deposition, the empirical
critical nitrogen loads developed in Europe (or elsewhere) may be applied, tested and modified
for Canada.
50
Figure 11. Critical loads for nutrient N and their exceedances using approaches in Bobbink et al (2002).
Source: Aherne (2007).
Georgia Basin specific empirical CL for nutrient nitrogen have been derived as part of the
Georgia Basin Action Plan founded in 2004 under direction of Environment Canada. The
approach was developed to assess the sensitivity of terrestrial ecosystems to atmospheric N
deposition and N cycling in the Mountain Hemlock zone (MH) of the Georgia Basin, British
Columbia (Zhong 2004). Some key indicators were employed, including geographic variables
(e.g., elevation), plant species, soil texture and chemical properties (esp. C:N ratio, and total N),
foliar N concentration, forest N pool and N retention capacity; all of which is closely related soil
moisture and nutrient regimes (Table 7).
51
Table 7. Indicators for the ecosystem sensitivity to nitrogen deposition in the Mountain Hemlock
Zone of the Georgia Basin (adapted from Zhong, 2004b).
Sensitivity Class
Very low
Low
Intermediate
High
Very High
Hardwood, dist.
(e.g., harvest)
Young stand
average growth
Low/short
Mixedwood
dist.
Medium age
stand
Medium/
medium
Mid-slope
unknown
Conifer
dist
Mature
stand
High/
long
Upper
slope
Windward
medium mt
unknown
Conifer
no dist.
Over-mature
stand
Very high/
very long
Indicative plant**
Hardwood,
no disturbance
Young stand
vigorous
Very low/
Very short
Very low
near sea-level
Leeward slope
high mt.
unknown
Mt ridge or top
Windward
high mt
unknown
Quantitative
Size of soil N pool
Very large
Large
Medium
Small
Very small
Capacity of N retention
Very high
High
Medium
Low
Very low
Soil organic horizon
C:N ratio#
Soil organic horizon
mineralizable N
Foliar N concentration
>> 29
> 29
29
< 29
<< 29
Very low
Low
Medium
High
Very high
Very low
Low
Medium
High
Very high
Indicator
Qualitative
Forest type/disturbance
Forest age/health
Growth rate/growing
season
Elevation
Slope aspect*
Low
valley bottom
Leeward slope
medium mt.
unknown
Neutral slope
Note:
*As atmospheric N movement is dependent on wind, the deposited amount of atmospheric N is closely related to slope aspect of a
mountain, a windward slope generally receives more atmospheric N than a leeward slope. The greater the elevation of a mountain,
the greater the slope effect.
**Some N-sensitive bryophyte plant species (moss, lichen, fern etc.) cover or importance can be used to indicate the degree of N
deposition. However, no data is available for the sensitivity of bryophyte species.
Nutrient nitrogen CL may also be calculated using mass balance methods. The critical loads for
nutrient N (CL nut (N)) is calculated as follows:
CL nut (N) = N u + N i + N de + N le(acc)
where N u = net N uptake rate (i.e. biomass accumulation); N i = acceptable net N accumulation
rate in the soil; N de = soil denitrification rate and N le(acc) = acceptable level of N leaching from
the rooting zone. All values are in eq/ha/yr (UNECE-CLRTAP 2004). A modified version
replaces N de + N le(acc) with N l.crit /(1-f de ) (Hettelingh et al. 1995) to give the form:
CL nut (N) = N u + N i + N l.crit /(1-f de )
where N l,crit = N le(acc) from above, and f de is the fraction of RN that is denitrified (0 ≤ f de ≤ 1) and
is a function of soil type (Downing et al. 1993; Fenn et al., 2008).
Exceedance of nutrient CL for nutrient nitrogen is:
52
CL nut N(EX) = (NO 3 -+NH 4 +) dep – CL nut (N)
Where the subscript dep refers to atmospheric deposition in kgN/ha/yr. Although the mass balnce
method for nutrient nitrogen CL has been used in the United States (e.g. Fenn et al. 2008) and
Europe (e.g. UNECE-CLRTAP 2004; Thimonier et al. 2009), the approach has not yet been
applied in Canada.
4.3.2 Alternative tools, models and indices
Environmental indicators are useful tools in all areas of natural resource management. They
make large problems more manageable and translate science in a way that is more easily
understood. Indicators are utilised in cumulative effects assessment, sustainable forestry
management and in assessing agricultural sustainability. Some of the indicators used in Canada
for other purposes, may prove useful to RN deposition and eutrophication management. For
instance, it was noted previously that lichens are very sensitive to relatively low levels of
pollution. Therefore, measures such as lichen diversity, percent cover, or foliar nitrogen can be
used as indicators of the effects, or levels, of pollutant deposition. Similarly thresholds can be set
for various measures at which change is considered “undesirable”. This is much like critical
loads, however other indicators can also be used.
The Sustainable forest management indicators of area and percentage of water bodies with a
change in physical, biological and chemical properties; and water bodies with significant
discharges of effluent (Sustainable Forest Management Indicators. Knowledge Base 2008), can
be modified and used for the purposes of eutrophication management. Similarly, residual soil
nitrogen is a Canadian agricultural sustainability indicator. It is equal to the amount of nitrogen
added to a crop minus the amount that the crop takes up (Figure 12). The N added is an estimate
of N from all land, water and atmospheric sources, including atmospheric deposition and fixation
by legumes (Drury et al. 2005). This indicator is also useful in the assessment of atmospheric RN
deposition to ecosystems and the equation could easily consider uptake by native, rather than
cultivated, vegetation. The amount of excess N contributed by each source could be managed
individually but with a unified target. The target would also need to take other social and
economic factors into account – for instance a city may favour limits on fertiliser use on lawns
and parklands, and improvements to sewage treatment, over limits to vehicular traffic.
Another related indicator used by Agriculture and Agri-Food Canada is the Indicator of the Risk
of Water Contamination by Nitrogen (IROWC-M). This indicator uses estimates of the N lost
and the concentration of nitrate in runoff to derive a level of risk. The greatest limitation with this
method is the estimate of runoff (precipitation – evapotranspiration) (De Jong et al. 2009) as is
also true for many critical load calculation methods. Recent research in western Canada that
defines runoff using isotopes, shows much promise in the refinement of runoff estimates across
Canada (Gibson et al. 2010). These physical or chemical indicators may be more reliable than
biological indicators whose response may be to a number of stresses. However, in this way
chemical and physical processes can also be affected by changes in land-use or climate regime.
53
Figure 12. Residual Soil Nitrogen under 2001 agricultural management practices. From: Drury et al.
(2005). The classes are from green to red: 0-9.9, 10-19.9, 20-29.9, 30-39.9, >40 kgN/ha.
Because of the complexity of chemical and biological responses associated with RN deposition
and enhancement, there exists a suite of potential tools and indicators for assessing and
monitoring the effects of RN enrichment and eutrophication. Chemical analyses of soil, water or
vegetation are of particular utility (CCME 2008a). Measurements that indicate current nitrogen
status or limitation in forest stands for instance, can be used to assess at what point eutrophication
may occur in that forest, by indicating how much nitrogen vegetation is using and at what point
the system may become N-saturated (i.e. have more RN than it can use in its current condition).
Indictors of nutrient status and limitation include: net N mineralization, net nitrification and foliar
N content – which if low, indicate N limitation (Aber et al 1993) and therefore increased
54
sensitivity to eutrophication. A low or decreasing C/N ratio may indicate increased RN
deposition and an increased risk of nitrate leaching (Gundersen et al. 2009; Reinds et al. 2009;
Thiomer et al. 2009). Forest crown health has also proven to be a good indicator of sensitivity to
RN deposition across eastern Canada and Vermont (NEG-ECP 2003). Additional models and
mathematical tools used for eutrophication assessment require the inputs of measured chemical
values that by themselves may also prove as useful indicators. Conversely, many of the models
may be used to assess or quantify indicators.
In aquatic ecosystems, indicators of RN sensitivity include algae biomass, high N, low oxygen
and high phosphorus concentrations (Walk 2007). Similarly, surface water nitrate is a good
indicator of RN leaching and deposition in watersheds (Aber 1997) and emissions of N 2 O from
freshwater systems can indicate the removal of excess nitrate through denitrification (Hill 1996;
(NO x SO 2 Management Working Group of the Cumulative Environmental Management
Association, 2008) also Fenn et al. 2008; Roobroeck et al. 2009). Measures of lake trophic status
(i.e. whether a system is oligotrophic – nutrient poor; eutrophic – nutrient rich, or mesotrophic –
somewhere in between) can provide information on the changing nutrient status of lakes.
Indicators that are commonly used for trophic status include measures of chlorophyll a – an
indicator of algal productivity; and total phosphorus – an indicator of whether or not the system’s
growth is limited by the nutrient P rather then the nutrient N.
A tool called the Planktonic Index of Biotic Integrity was developed for measuring ecosystem
changes in Lake Erie using measurements of phytoplankton and zooplankton diversity, biomass
etc. to indicate trophic status and changes (Kane et al. 2009). This technique holds promise for
other freshwater systems and can be linked with paleolimnological studies of plankton fossils
providing a sort of baseline measure from which change can be determined. Paleoecological
studies have been used to recreate ecological histories in Canadian marine (Mudie et al. 2002)
and freshwater (Moos et al. 2009) environments by analysing the contents of sediments with
depth – deeper sediments represent the ecosystem composition further back in time. Similar to
plankton, other invertebrates are also good indicators for nitrate levels in lakes and streams.
Dynamic models are built on more complex equations relating to ecosystem processes that
change over time. The Simulation Model for Acidification’s Regional Trends (SMART) soil
chemistry model was devolved for acidification purposes, and although it does not simulate the
processes of nitrogen cycling, the processes of nitrification and denitrification are modelled using
simplified processes. SMART may be used for predicting soil nitrate and aluminium
concentrations based on the deposition and soil concentrations of the elements in question.
Measures of nitrate leaching and aluminium can serve as indicators of nitrogen saturation and
acidification (discussed in the next section). The SMART model has been applied to nitrate
assessments in Europe (e.g. Posch et al. 2003; Reinds et al. 2009), but it does not appear to have
been applied in Canada.
The DayCent model has been used with success in California to estimate the values for
eutrophication indicators such as C/N ratios, tree growth, decomposition and RN inputs and
outputs for current and historical environmental states (Fenn et all. 2008) and has also been
applied successfully in the Great Plains of the United States (Del Grosso et al. 2001) for which it
was developed. DayCent is a daily time-step version of the CENTURY biochemical model. With
a full description available from http://www.nreal.colostate/edu/projects/centruy5/ CENTURY
55
simulates growth and C and N cycling in grasslands, forests and crops using a variety of chemical
and physical inputs parameters including: temperature, precipitation, deposition, soil texture, soil
nutrients (C, N, P) plant nutrients, and plant lignan. Appropriate use of CENTURY and its
derived counterparts is limited to N-limited soils and is not suitable for use in areas characterised
by P-limited soils (Gijsman 1996). CENTURY has been used in Alberta for predicting
Greenhouse Gas cycling and inputs from the Soil Landscapes of Canada (SLC); however most of
Alberta was not characterised by appropriate data in the SLC database (Sauvé et al. 2000).
PnET is a suite of nested models that simulate water, carbon and nitrogen dynamics in
(temperate) forested ecosystems. PnET-CN is an extension with algorithms for closure and full
portrayal of the N-cycle (see http://www.pnet.sr.unh.edu/ for a full description and model
download). Model input includes foliar nitrogen, leaf area index, foliar mass, radiation flux,
vegetation cover, climatic variables and soil properties such as water holding capacity. A similar
model, but with improved soil organic matter and nutrient cycling algorithms to include
processes such as the influence of N-saturation on decomposition, is a version of IBIS the
Integrated Biosphere simulator (http://www.sage.wisc.edu/download/IBIS.ibis.html). This
updated model includes additional N-feedback controls to better represent northern Canadian
forests (Liu et al. 2005). IBIS simulates growth and competition, and nutrient or energy cycling
in forested ecosystems.
The SWAT model has been used to assess nitrogen loading at a watershed scale and has been
used in Canada to assess variable buffer lengths for streams in Ontario and Quebec (Chambers et
al. 2008). The input requires mapped layers of precipitation, digital elevation, land use and soil
texture, and the output gives changes in nutrient loading dependent on the various input factors.
The Model of Acidification of Groundwaters In Catchments (MAGIC) has been applied to
watersheds in Nova Scotia (Dennis et al. 2005) and both soils and lakes in the Athabasca oilsands
region of Alberta (Whitfield et al. 2010). MAGIC and simulates short-term equilibrium reactions
and longer-term geochemical fluxes within soils. Although N-fluxes simulated by the model are
largely conceptual, updates to the model now include nitrogen dynamics controlled by soil N
pools (Cosby et al. 2001). Input includes soil parameters such as weathering rates, N-retention
and a critical C/N ratio.
4.4 Tools for Acidification
In Canada and abroad tools for assessing the sensitivity of ecosystems to acidification are
generally more developed than their eutrophication counterparts. Due to the on-going research
and knowledge in this area through various programs, these tools are described only briefly here
and the reader is referred to other sources of documentation for full method derivations.
4.4.1 Critical Loads of Acidification
Unlike the CL for nutrient nitrogen, when calculating the CL of acidity, sulphur (in the form
SO 4 2-) must be considered along with nitrogen (in the form NO 3 -). Low-lying lakes and those
associated with first order streams (those at the end of a system) maintain water chemistry
approximately equal to the average chemistry of their basin’s soils. Because this allows one
measurement to be representative over a larger area, freshwater lake CL are often utilised in
exceedance calculations and acidification sensitivity maps (e.g. Aherne et al. 2004; Jeffries et al.
2010). Acidification CL have been estimated in Canada using simple empirical methods
56
(National Research Council of Canada 1981; Saskatchewan Research Council1982; Swain 1987
and Krzyzanowski 2010); simple mass balance methods (Jeffries and Ouimet 2005, Watmough et
al. 2005; Aherne and Watmough 2006; Aherne 2008a; Jefferies 2010); and dynamic models
(Whitfield et al. 2006).
Other indicators of acidification progress and sensitivity are also used. The base cation to
aluminium ratio is often used as a chemical limit in critical loads calculations, but could also be
used as a chemical indicator and measured in the field to track acidification status of ecosystems.
A BC:Al ratio of 10 has been used as a limit across Canada (CCME 2008).In Nova Scotia
(Whitfield et al. 2010), Manitoba and Saskatchewan (Aherne 2008a) and other Canadian studies,
alternative ratios have been used including 6:1 in Fort McMurray, and 1:1 for other parts of
Canada (Tominga et al. 2007). Also, because soil saturation of NO 3 - occurs before acidification,
soil nitrate can be used as an indicator of a future acidification problems (Ojima and Baron 1999)
in addition to being an indicator of eutrophication.
5.0 Directions for RN Effects Assessment and Management in Canada
The following sections address each of the research questions stated in the introduction. The
questions are approached by summarising what we know; what we don’t know; and what we can
do to know more - in terms of RN deposition and effects management in Canada. An underlying
theme within each of the research questions is the need to define what it is we are trying to
protect. What we are protecting determines how we: inventory emissions, direct research, locate
monitoring sites, choose eutrophication limits, define critical loads and manage or identify
“undesirable” effects. A combination of methods is often the most desirable way to approach RN
effects assessment. In Canada and around the world, various techniques have been developed and
used to assess the risks of eutrophication, acidification and essentially ecosystem injury from RN
deposition.
Current Canadian management efforts focus on regions with increasing NO x and NH 3 emission
sources -Alberta Oil Sands Regions (OSR), the Lower Fraser Valley of British Columbia, and
those that have experienced impacts related to eutrophication or acidification in the past such as
the Great Lakes Region or Southern Ontario (Environment Canada 2001, Alberta Environment
2008, Carou et al. 2008). These programs are an invaluable resource, but do not necessarily
provide enough information from which to derive policy and management strategies. This section
addresses what is required to appropriately derive such strategies.
The following discussion combines background information provided in the preceding sections
with information from a survey of researchers. The survey was sent in January 2010, to 53
researchers working in the fields of atmosphere and ecosystem modelling, deposition and effects
measurements, and controlled experiments, related to RN in Canada. Of these researchers 16
responded to the short anonymous survey on current knowledge and gaps pertaining to RN
deposition and effects research in Canada. These survey responses, along with the information
above, are employed in answering each of the research questions.
5.1 Measured and Modelled Deposition
Before limits can be set on atmospheric emissions, or strategies developed to reduce RN inputs to
ecosystems, the current RN inputs must be quantified. Monitoring, modelling, or a combination
57
of the two, are required to measure how much reactive nitrogen is present in the air and being
deposited on the surface at any given time or location. It may be necessary to measure and
calculate both wet and dry RN deposition, respectively, in its various forms (see Table 2).
Different RN compounds cause different responses in ecosystems; and it has been found for
instance that in western Canada dry deposition makes up a larger proportion of total deposition
than once thought (Chambers et al. 2001; Zhang et al. 2005).
The Federal CAPMoN and NAPS monitoring networks collectively measure wet, dry, gaseous
and particulate forms of RN (Figures 6 and 8). A number of provincial programs also monitor
atmospheric forms of RN (Figure 7 and 8). Measurements have shown that air quality and acid
rain related issues have been concentrated in eastern Canada, specifically southern Ontario and
Quebec. Modelling with AURAMS (Figure 10) has also predicted high levels of deposition in
these areas. However, AURAMS also predicts high levels of total nitrogen deposition in
southwest British Columbia, southern Manitoba and Saskatchewan, and throughout much of
Alberta. There are also higher densities of NAPS sites located in these areas.
Dispersion models, although prone to large errors, are particularly useful in predicting problems
areas where monitoring should be focused. However, to ensure that the monitors are capturing all
problem areas, the model needs to accurately capture the area’s emissions. It has been shown
previously in northeast British Columbia that this is not always the case. Year 2000 emissions of
NO x more than double in the region when upstream oil and gas sources that fell below reporting
thresholds were included in the emissions inventory (Krzyzanowski 2009). The same discrepancy
between reported and actual emissions likely occurs for other hydrocarbon producing parts of
Canada including Alberta, Saskatchewan and Manitoba. RN deposition may be higher than
predicted in these regions and there are currently no monitoring network measurements in such
areas with which to compare with modelled values. In northeast British Columbia, when
numerous and unreported small emission sources were included in a small-scale model run using
AERMOD, the ambient air quality objective for NO 2 was exceeded on numerous occasions over
the modelled year (2000) (Krzyzanowski 2010). These high concentrations indicate that the
deposition of RN compounds may too, be higher than anticipated.
Since the year 2000 Canada’s emissions reporting policies have changed dramatically. Emissions
inventories are now federal through the NPRI and only those facilities with 20 000 employee
hours per year, and emitting greater than 20 tonnes of NO x (as NO 2 ) need to report their
emissions; unless they are from particular activities such as agriculture, vehicle repair and
dentistry (Canada 2007). This policy neglects to account for thousands of ‘smaller’ Canadian
industrial emission sources. The inclusion of these sources and removal of such reporting
exemptions would likely increase both the modelled RN deposition and the accuracy of simulated
deposition fields. Guidelines for federal standardisation of emissions reporting due to
inconsistencies across the country have been developed (Marbek Resource Consultants Ltd.
2008), however they fail to acknowledge the impact of underreporting from numerous small
industrial emission sources.
Other sources of error in modelling include the estimation of actual deposition velocities, and in
particular the dry deposition of gases and particles. Actual in-field measurements across a variety
of landscapes could help elucidate this in Canada. In addition to deposition velocities, models
derive other parameters from land-cover datasets that may not be accurate at the desired scale. In
58
general there is a heavy reliance on modelling without the actual measurement of data to support
predicted fields. These issues could be alleviated by a monitoring network that fills in many of
the spatial gaps that exist across Canada’s diverse landscape. Models also do not account for
spatial variability, especially when using a 42 km resolution grid for modelling. There are other
unknowns in terms of how the deposition of different types of RN varies with altitude. If model
output could be validated using field measurements, then a finer model resolution - requiring
much faster and larger computers - would be warranted. Until then, the spatial distribution of RN
deposition across most of Canada will remain unknown.
An expansion of monitoring sites to the northern reaches of the western provinces would help to
determine if there is potential for high RN deposition. There are generally too few routine
measurements of RN in western Canada. However the location at which to place these sites may
be determined by use of models that utilise a comprehensive emissions inventory. Because
continuous monitoring can be expensive and require power sources, other techniques may be
desirable for more remote locations. Use of cumulative deposition measurement techniques such
as passive sampling, precipitation sampling and IER are well suited for assessments that utilise
CL for instance, because CL consider the cumulative deposition over an annual period. In
addition, vegetation, soils and aquatic systems receive a cumulative dose of RN from the
atmosphere. Methods from Fenn et al. (2009) for various modes of bulk deposition measurement
including passive sampling, resin tubes etc. and methods from Krupa and Legge (2000) and
Krupa (2002), can be used to define the best monitors for a particular location over a specified
period. Ideally, these passive or cumulative monitoring sites can be correlated or collocated with
some existing CAPMoN or NAPS sites for validation.
Monitoring results can be used in combination with modelling results to examine if, where and
when RN deposition may have a negative impact on ecosystems. The monitoring of as many
different RN compounds as possible and using a type of monitoring that can speciate between the
different forms of deposited RN, are optimal. Different chemical forms of RN effect ecosystems
differently, and along with ecosystem characteristics, determine whether deposition will be
acidifying, eutrophying, or acutely toxic. For instance, since regional models are generally poor
at predicting the dry deposition of NH 3 , the monitoring of dry NH 3 in high source areas would
help resolve some of the issues related to its modelled values and improve models such as
AURAMS. Similarly, N 2 O is not reported to the NPRI by any facilities, and may contribute to
some of the atmospheric RN deposition. However, the fate and uptake of N 2 O in the environment
is not well understood and it is not monitored anywhere in Canada. Regional-scale measurements
of N 2 O, NH 3 and organonitrates are lacking in Canada. These measurements would facilitate in
estimation of dry deposition fluxes. Organonitrates are numerous forms of organic molecules
containing nitrate. They also provide sources of RN to ecosystems and may be found in measured
precipitation.
Canada’s CAPMoN precipitation monitoring sites (Figure 6) are capable of measuring dry
particle deposition and gaseous deposition in addition to the currently analysed wet RN
contributions. Because the samplers are always open and exposed to the atmosphere they can be
used to measure bulk deposition minus losses from chemical reaction or evaporative loss (Krupa
2002). Because of the link between depositional forms and ecosystem uptake, the separation of
wet and dry fractions in the RN deposition measured by these stations would contribute to our
understanding of potential effects
59
The use of historical data from monitoring networks is invaluable and can be used to develop
baselines atmospheric conditions from which to measure change. Former CANSAP (1977-1985),
CAPMoN (1985-present) and NAPS (1969-present) sites will be especially useful in determining
these baselines. The quantification of RN deposition through either monitoring or modelling is
best paired with the quantification of RN contributions from other sources. While much of this
nitrogen may be measured in ecosystem monitoring of N compounds, correlations between
atmospheric RN deposition and measures of foliar N concentration or plant response, may be
confounded by other sources of RN in the environment.
5.2 Physical, Chemical and Biological Processes
Little is known about the long-term or short-term effects of RN deposition on Canadian
ecosystems and their components. As of yet, Canada has no nitrogen uptake, saturation or
leaching thresholds for different terrestrial ecosystems; nor do we know how climate, soil texture
and hydrology may affect RN uptake, saturation and leaching in terrestrial systems, or input to
aquatic systems
The most convenient way to draw relationships between RN deposition and response is to
monitor responses in an area of known deposition. One way to do this is to co-locate forest or
ecosystem monitoring plots with atmospheric deposition monitors – similar to the ICP Forests
monitoring plots in Europe (see: http://www.icp-forests.org). An expanded CAPMoN or NAPS
network could potentially fill such a purpose. Useful locations for new monitoring sites would be
those that are considered (from research in Europe) to be particularly sensitive to atmospheric RN
deposition such as tundra (Bobbink et al. 2002; Bobbink and Roelofs 2005; Aherne 2007);
softwater lakes; and alpine areas (Bobbink et al. 2002; Table 5). More remote areas in the
sampling network may utilise some of the passive and cumulative techniques discussed
previously. There should be monitoring sites and research done in every type of Canadian
ecozone in both aquatic and terrestrial ecosystems. Because biological response is often species
or ecosystem specific, it is crucial to represent all dominant Canadian ecosystem types.
Another way to monitor responses in an area of know deposition is through experimental and
controlled additions of RN. Such approaches have been used previously in eutrophication science
in Canada shedding light on the P-limitation of freshwater ecosystems (Schindler 1974).
However similar research in terrestrial ecosystems, bogs etc. with varying levels and different
forms of RN would also yield useful results. The results from these small isolated field and
laboratory experiments can then be scaled up and used to develop models or thresholds for
Canadian ecosystems. While microcosm studies are not generally accepted as the basis for
critical loads, they can provide useful information on the sensitivity of a plant of particular
significance (Bobbink et al. 2002). In this way, thresholds can be set for Valued Ecosystem
Components (VEC) and the concept of change simplified in a manner akin that of cumulative,
strategic and integrated effects assessment (see CCME (2009) for the use of VEC in Regional
Strategic Environmental Assessment in Canada). VEC should be chosen for their relevance to
regional scale analysis and biodiversity.
60
Research using either correlated monitoring or controlled experimental approaches needs to
address a number of unknowns regarding the fate and influence of atmospheric RN in Canadian
environments. Specific studies, preferably carried out in a variety of ecosystems include but are
not limited to: monitoring the direct impacts of reactive nitrogen in its various forms on
individual species of plants and micro-organisms; measurements of the rate at which
atmospherically deposited reactive is returned to the atmosphere in a non-reactive form (as N 2 or
N 2 O through denitrification); the alteration of nutrient balances in effort to identify limiting
nutrients and the relationships between them in various terrestrial and aquatic ecosystems; and
measurements of N in plant and lichen tissue to set known relationships between tissue
concentrations and deposition. Each of these studies and the variables they measure need to be
directly correlated (or collocated) with measured or controlled deposition.
Similar to controlled experiments, field measurements carried out near continuous emission
sources offer standard conditions in which to measure the gradual changes of ecosystems an its
elements. While in the field, it is difficult to isolate other variables that may too be impacting the
ecosystem (such as co-emitted pollutants, climate change, or pests).These coexisting stressors
also impact ecosystems and their responses to enhanced RN. Therefore an examination of the
combined effects and synergies occurring between atmospheric RN and other variables offers
guidance for future management and realistic priorities. For instance the effect that RN may have
on the export of dissolved organic carbon from soils has implications for climate change, and
climate would likewise impact the effect that RN has on carbon exports.
Ecosystems under various levels of deposition should be monitored for species changes,
particularly sensitive communities of lichens or tundra ecosystems. Macro- and micro-fossil
material (partially decomposed and identifiable seeds, leaves, woody tissue, etc.) can be used in
some areas, particularly lake sediments and peatlands, to recreate historic species distributions in
key areas where historic data does not exist. For example in coastal areas (Gooday et al. 2009)
and inland waters (Moos et al. 2009) of Canada, micro-fossils have been used as biomarkers of
eutrophication. This is necessary to create a baseline condition from which conditions change.
Without a baseline it is difficult to establish whether or not something has indeed changed; and
controlled cause and effect-based studies are necessary to determine if that change is in fact
caused by reactive nitrogen.
5.3 A Comprehensive Monitoring Program
A multitude of factors are associated with an ecosystem’s sensitivity to nitrogen deposition
effects. Many of such factors need to be known before an assessment of vulnerability to
eutrophication or acidification can be made. A comprehensive monitoring program would include
the measurement or assessment of these variables, once research has identified how each variable
relates or responds to, atmospheric RN deposition.
There is data already available in Canada that can be used to indicate and assess change. For
instance surface water nitrate concentrations (Aber 1997) can be used as an indicator of nitrate
leaching; and measures such as total nitrogen (Chambers et al 2008), phosphorus (Walk 2007)
and ammonia in surface water indicate how prone an aquatic ecosystem may be to eutrophication.
Measures of these chemical attributes in lakes and rivers are made across Canada, usually as part
of comprehensive provincial monitoring networks. However, the measurements that are made
61
and how they are taken varies considerably across the nation.
At a federal level this information is supplied as a water quality index (WQI)) as part of the
Canadian Environmental Sustainability Indicators multi-level effort to define thresholds of
environmental sustainability in Canada (Environment Canada 2008b). The WQI – based on a
number of chemical attributes including nitrate, bacteria and pesticides – describes water quality
as fair, marginal, good or excellent (CCME 2001). What is measured, the concentration guideline
used, and the sampling frequency, vary across jurisdictions making it difficult to compare the
WQI across provinces and impossible to know how the WQI corresponds to eutrophication or
RN levels. Across Canada there are 392 river and 19 lakes monitoring sites that report WQI. All
provinces measure P at some sites and all provinces but Newfoundland include some form of
nitrogen measurement in the WQI (Table 8). However, this does not mean that Newfoundland
does not measure nitrogen compounds in any of its surface water systems.
Table 8. Summary of N and P monitoring at sites reporting WQI through CESI. A = all sites measure the
parameter, S = some sites measure a parameter.
Site Number
Lakes Rivers
AB
0
31
BC
0
32
MB
6
34
NB
1
45
NFL
5
17
NS
6
2
NT
1
9
NU
0
1
ON
0
80
PE
0
11
QC
0
121
SK
0
7
YK
0
2
CANADA 19
392
NO 2
NO 3
NO 2 +
NO 3
S
S
NH 3
Total N
P
S
S
S
S
S
S
A
A
A
S
A
A
A
A
A
A
A
S
S
A
A
S
A
A
A
A
A
A
A
A
A
A
A
A
A
A
Most provinces collect a substantial quantity of surface water data, but they are held and analysed
separately. A standardisation of measurements and/or a communal database, would greatly
improve our knowledge of eutrophication sensitivity and progress fresh water systems and their
watersheds. Since such a diverse network of surface water monitoring already exists in Canada,
such a network should be utilised more extensively and made standard and accessible across
Canada. Chlorophyll-a and lake trophic status are also included in Alberta’s lake data providing
additional indicators of eutrophication. Phytoplankton biomass can provide an additional
indicator of eutrophication. Detailed water quality datasets can be used to assess sensitivity to
eutrophication and monitor chemical or biological changes over time.
62
The data available through NAtChem and NAPS is most useful in determining the state of
atmospheric deposition. However, some spatial gaps exist in the measurement sites, particularly
in regards to Western Canada, where RN is on the rise. There are a variety of geospatial
techniques that have evolved in order to design comprehensive and representative monitoring
networks. Geospatial models can be used as tools for developing such a network. The technique
involved in such a design is called “sequential sampling” and a version of ESRI’s ArcGIS
software to be released later in 2010, includes network optimisation tools that utilise these
techniques (Witold Fraczek Personal Communication January 2010). The derivation and use of
this tool to create an optimal sampling network is described in Fraczek and Bytnerowicz (2007).
The sampling network can then be used to geostatistically interpolate values of deposition etc.
where direct measurements are not made through a mathematical weighting technique called
“kriging”.
Geostatistically, a reliable monitoring network should contain 100-300 measurement points, and
the minimum number of measurements required to be able to come up with a statistically
representative deposition field through kriging is 40 (Witold Fraczek Personal Communication
January 2010). There are currently only 27 CAPMoN sites measuring precipitation chemistry in
Canada, 15 measuring dry gas and particulate RN and 3 with continuous hourly RN gas monitors.
An ideal monitoring network would include a minimum of 40 stations all collecting the same
data the same way. This would mean at least 13 more CAPMoN sites all equipped for the
measurement of precipitation chemistry, dry gas and particulate RN species. Preferably the RN
would be speciated into each of its individual chemical forms. Although there are additional
provincial networks collecting precipitation chemistry data, measurement and analysis techniques
vary, a standardisation of what and how chemical RN species are measured would increase the
ability to compare such data. In addition the measurement of dry gases and particles are not
included in these networks, and the importance of dry and particulate RN deposition is greater
than once thought.
That being said, 40 is not a magic number, just a statistically derived minimum. Variation in
deposition occurs with changes in elevation, surface type (water, forest, rock), etc. Therefore it
may be necessary to place all monitors at the same elevation at the same time because a mix of
elevations will add additional error and unknown to kriged predictions. Alternatively, associating
all measurements with a factor such as elevation or surface type may allow for further statistical
cokriging of measurements to create continuous mapped fields of deposition. It may be advisable
therefore to have 40 stations in the tundra, 40 stations in boreal, 40 stations over freshwater lakes,
etc.
The 319 station NAPS network can help determine the dose of RN that ecosystems are receiving
through measurements of NO 2 , however much of the RN that reaches and eventually impacts
ecosystems comes in forms other than NO 2 (Table 2). For instance the additional monitoring of
NH 3 by this network would greatly increase its relevance to RN deposition and effects
monitoring. However, the use of hourly concentrations for the determination dry deposition is
limited by our knowledge of surface fluxes and deposition velocities for most surfaces and
vegetation types in Canada. Therefore the measurement of actual wet deposition through IER and
similar techniques is desirable in addition to the measurement of gaseous concentrations from
which we may infer dry deposition with knowledge of surface characteristics. Throughfall
measurements are useful in determining the RN that reaches the forest floor and can be compared
63
with open deposition monitoring to indicate the rates of uptake by a vegetation canopy. In
addition most NAPS sites are located near cities and along linear features such as roads for
accessibility. From a statistical perspective monitoring location should distributed in a dispersed
and representative manner. In order to make monitoring cost-effective, passive and cumulative
measurements should be considered along with more frequent collocated samplers at some sites
for calibration and guidance.
Measurements of deposition should be collocated with ecosystem monitoring so that the state and
change of ecosystems under different levels of RN deposition can be determined. A similar
system to that used for ICP forest monitoring in Europe could be used in Canada. It is desirable
to also measure the occurrence of other pollutants and environmental stressors that could too be
responsible for changes in ecosystem health or biodiversity. It is not ideal to have separate
monitoring networks for different pollutants. Monitoring sites should include monitoring of
lichen diversity because these communities are the first to change. In order to be able to assess
how similar ecosystems will respond to similar levels of deposition a number of ecosystem
variables should be determined as potential elements to which effects may be related. The
variables with the best relationship can then be used to predict change in similar ecosystems.
These elements include:
i) Type of ecosystem and plant species composition: Some ecosystems are more sensitive to
the fertilising effects of N than others. Aquatic diatoms and terrestrial lichens are
particularly sensitive to the effects of RN deposition are increasingly being used as
bioindicators and in determining CL (Tkacz et al. 2008; Fenn et al. 2008). Additionally,
the forests that contain trees with high nutrient demands may be more sensitive to nutrient
leaching associated with acidification. Some wild trees like the legume Locust (Gleditsia
spp. or Robinia spp.); and non-legume trees such as Adler (Alnus spp.) (Bormann et al.
1993; Brockley and Sandborn 2003), and Pine (Pinus spp.) (Bormann et al. 1993) as well
as legume and non-legume shrubs (MacConnell and Bond 1957) - fix atmospheric
nitrogen via symbiotic relationships with soil microorganisms, adding RN to the available
pool.
ii) Community structure and dynamics: These include measures of plant species density and
diversity or detailed maps of species distributions around the site such that any changes
over time or in relation to changes in RN inputs can be determined. Changes in lichen or
understory species are documented RN related ecosystem effects. This information can
also be used in deriving deposition velocities in locations measuring gaseous
concentrations using passive or continuous sampling techniques.
iii) Biomass and growth: Long-term monitoring studies at these sites should measure
elements such as biomass accumulation and growth to note any changes related to levels
of atmospheric RN deposition.
iv) Water and soil chemistry: Because freshwater ecosystems are representative of the soils
that surround them, freshwater chemistry should measured and determined in the
watersheds or basins where atmospheric and ecological monitoring take place. Elements
that serve as indicators in surface water (NO 3 , total N, P, chlorophyll a, N:P ratios, etc.),
and in soils (C/N ratio, BC:Al, etc.) should be measured at the monitoring sites. These
measures can be used to determine nutrient status and limitations and the limits of N
saturation and leaching.
v)
Animal species composition/density: The ecosystems composition of faunal species
should also be considered. In aquatic systems some zooplankton and benthic invertebrates
64
vi)
vii)
viii)
Land-use: The proportion of agriculture, aquaculture, natural and semi-natural
ecosystems; industry, settlements and municipalities with waste waster releases, and local
emission sources, around each monitoring location should be determined and mapped.
Within each of the land-use types non-atmospheric anthropogenic releases should be
determined. Estimates can be made through measures of forest harvesting and fires, and
the use of nitrate-based fertilisers, fire-retardants and pesticides through sales and taxation
data.
Runoff estimates: In order to estimate rates of N leaching you need estimates of runoff.
Conventionally quantified using the precipitation surplus method (i.e. precipitation minus
evapotranspiration) it has been found that this approach is inaccurate and may cause
errors in critical load estimates. A more accurate, expensive and complex approach is the
use of stable isotopes (Gibson et al. 2010). These values are necessary in understanding
when RN may become acidifying rather than eutrophying.
Climate and temporal factors: Due to the interactions between climatic variables such as
temperature, and water availability on both atmospheric chemistry and ecosystem
responses to RN deposition, these variables should too be measured. This is especially
important in understanding the potentially complex interactions between climate change
and RN effects. Long-term and continuous measurements of climatic variables can also
provide additional insight into seasonal variations in both atmospheric chemistry and
ecosystem response.
5.4 Eutrophication or Acidification
Since the 1990s, critical loads for acidity have been identified for various ecosystems in Canada
(CCME 2008a,b), while less attention has been paid to critical loads of nutrient nitrogen.
Nevertheless, many of the data collected over the past 20 years for the determination of acidity
critical loads, can also be used for developing eutrophication critical loads and their exceedances.
These data include measurements and estimates of: N deposition, N immobilisation, N uptake,
denitrification and leaching rates. Canada’s National Acid Rain Science Plan
(http://www.ccme.ca/assets/pdf/ar_ntnl_science_plan_e.pdf) was developed by CCME’s ARTG
and recommends catchment-based studies examining the linkages between ambient RN
concentrations, deposition, nitrification, impacts, and relationships with eutrophication. Tools
like critical loads for nutrient N need to be developed for Canadian ecosystems in order to
manage and minimise eutrophication. While measurements of N and P help define nutrient status
in soils, it is also necessary to know the status of other nutrients such as P, Ca2+, Mg+, Na2+ and
K+, their uptake from growth, their loss from harvesting (Watmough and Dillon 2001 and 2003),
their release from fire, and their input from deposition. If a system is limited by another nutrient,
RN deposition will not increase growth. A negative mass-balance for base cations indicates a
long-term nutrient limitation. If all cation mass balances are positive, the system may be sensitive
to eutrophication, but not as sensitive to acidification. Similarly, if a system is P limited additions
of N may not make much of a difference in terms of growth enhancement.
65
In freshwater, NO 3 - measurements, especially increases in water NO 3 - over time, indicate that Nleaching is occurring in the watershed. Nitrate leaching is a symptom of N-saturation in the
ecosystem, meaning that no more uptake can occur through vegetation and soil processes. This
happens after the RN enrichment effects of growth increases and changes in species composition
to promote the growth of nitrophillic species. Once there is no further uptake, NO 3 - becomes
mobile and is accompanied by other exchangeable cations (H+, Al3+, Ca2+, Mg+, etc.) in order to
maintain the soil’s charge balance. Indicators of N-leaching include low C:N ratios; high fluxes
of RN deposition; high throughfall N flux to the soil; enhanced N in foliage or litter; and low soil
pH, while low mean annual temperature is a good indicator of ecosystem susceptibility to Nleaching (Dise et al. 2009). When nitrate is mobile and leaching, acidification is beginning to
occur. Changes in soil pH and decreases in BC:Al ratios are signs of the base cation depletion
and aluminium mobilisation associated with acidification.
Sensitivity to acidification is determined by an ecosystem’s ability to neutralise incoming acidity
(H+). This is often termed the soil’s “buffering capacity” and is linked to the soil’s acid
neutralising capacity (ANC, which is = concentration of strong bases – concentration of strong
acids). Carbonate-rich soils derived from limestone parent material, have a high buffering
capacity (low sensitivity), whereas soils derived from plutonic or igneous rock such as feldspar
and granite, have very low bases and a low ANC or buffering capacity. Various types of
sedimentary rocks fall in between (UNECE-LRTAP 2004; Dupont el al. 2005). Base cations
(BC) may also be supplied atmospherically, and measurement of base cation deposition carried
out by some provincial monitoring networks (for example Alberta’s and British Columbia’s
networks), is of extreme value when considering potential deposition effects. Not only does the
atmospheric deposition of BC increase buffering capacity and therefore ecosystem resistance to
acidification, but abundant BC reduces the risk of nutrient growth limitations and increases the
ecosystem’s susceptibility to eutrophication.
Certain soil properties are of interest when considering the limits (or critical loads) of
eutrophication and acidification for ecosystems. Soils are formed out of parent material that has
eroded and been covered / mixed with organic matter from decaying organisms. Soils that contain
high base cations from their parent materials, have a low sensitivity. Organic soils (e.g. northern
peatlands) are low in base cations and naturally acidic, but have a large buffering capacity due to
a system of acid buffering. In addition the particle size of the soil infers the availability of cation
exchange sites - for instance clay, made of fine particles, has a large negatively charged surface
area in comparison to its volume and therefore holds onto cations better than soils with larger
particles. However, fine particles such as clay being dominant may also mean that the soil has
undergone most of its physical and chemical weathering and that the influx of new base cations
will be low.
The availability of metals to the system is important in determining the type of ecosystem effects
that may occur as the result of NO 3 - leaching acidification. The plant and fish toxicity associated
with acidification is largely caused by toxic aluminium (Palmer and Driscoll 2002; DeVries et al.
2003; Henriksen and Posch 2001), which is released from aluminium silicate minerals by the
addition of (NO 3 -) and H+ to the soil complex. Additionally, indicators that relate Al3+
concentrations to other soil properties such as carbon (C) can be used to assess soil sensitivity to
acidification. Al3+ is not the only metal ion that behaves this way, but it is the most prolific. Other
metals such as zinc (Zn), lead (Pb), cadmium (Cd) and copper (Cu) can also be released from the
66
soil complex by acidification and are toxic to nitrogen fixing symbiotic root fungi (particularly
mycorrhizae and ectomycorrhizae) (Fomina et al. 2005), soil bacteria and plant roots (Mayer
1998) thus affecting the way that nitrogen is cycled in the soil.
5.5 Eutrophication Critical Loads
Aherne (2007) used European empirical critical loads (Bobbink et al. 2002, Tables 6 and 7) for
nutrient N to produce a critical load exceedance map for Canada. Although estimated using
evidence of eutrophication and nutrient enrichment in Europe, this map provides a good starting
point from which to identify potentially sensitive areas and begin developing critical loads of
nutrient nitrogen specific for Canadian ecosystems.
The European nutrient nitrogen / eutrophication critical loads from Bobbink et al. (2002) can be
modified and refined using information from the monitoring described above. However, we also
need to understand the physical, chemical and biological processes that regulate RN deposition
and effects. As our methodologies get more complicated we need to justify the complexity with
known variables, not assumptions. For instance the amount of RN that intercepts the soil surface
depends on snow cover dynamics (Fenn et al. 2009); how much is taken up by the vegetative
canopy; and what the vegetation does with it (i.e. does it emit it through its roots in usable form,
emit gaseous N 2 O, grow, etc.). Some of these factors can be evaluated using a variety of
measurement techniques, but these variables and their constants need to be established as rules to
extrapolate to areas of Canada where measurements cannot be made. Field and laboratory
experiments with nitrogen enrichment can be useful in determining CL for Canadian organisms
of interest. Studies can be done on species defined as indicators of the threshold between
acceptable and unacceptable effects. VECs are useful as such species and can be defined using
cumulative effects concepts. However, studies that consider the effects of enhanced RN
deposition on multiple ecosystem receptors are the most desirable.
The mass balance method for calculating nitrogen critical loads discussed in section 4.2.1 is an
optimal approach for determining the CL of nutrient nitrogen. This mass balance should consider
all inputs and outputs of RN from the system. The calculation requires knowledge of N uptake or
biomass accumulation by vegetation, rates of denitrification; and the determination of acceptable
soil RN accumulation rates and acceptable N leaching from the rooting zone. Typically the
method requires that the fraction of RN that is denitrified be determined for Canadian soils and
uptake rates determined for native species of vegetation. Measurements of soil C/N ratio can be
used to estimate leaching, but the relative retention of RN by soils and vegetation is largely
unknown (Gundersen et al. 2009). This can be partially alleviated through other measurements
such as total N in soil and litter, and drainage characteristics, that too indicate RN leaching in
soils (Dise et al 2009). The variables need to be determined for all main soil and vegetation types
in the country through observation and experiment. This mass balance method is then a desirable
method by which to map CL of nutrient (reactive) nitrogen in Canada.
Much of the forest data required to estimate N uptake or leaching rates may be available in
industry or government forest databases. A forest disturbance index (FDI) for instance can be
derived from readily available Landsat imagery for estimating variations in N in stream-flow as
related to various land-use and forest harvesting practices (Eshleman wt al. 2009). Data on forest
management practices (e.g., forest harvest, fertilisation) and actual RN deposition measurements
67
are necessary to determine with certainty the effects of N deposition on forest structure and
function (Zhong 2004). The nitrate retention and leaching estimates required to calculate mass
balance CL for RN can also be estimated using models such as FAB (First Order Acidity Balance
Model) (Henriksen and Posch 2001; Curtis et al 2005); and dynamic models such as SMART,
CENTURY and MAGIC can be used to assess other elements of nitrogen cycling. However,
these models also require input such as soil nutrient concentrations and plant uptake rates may be
difficult to find for most Canadian environments. Additional model inputs such as runoff
estimates have been shown to be erroneous, but can be greatly improved using isotope analyses
such as those carried out in Alberta (Gibson et al. 2010). It is suggested that a monitoring
program, such as that described above, be developed to measure actual deposition, runoff and
responses of ecosystems to fertilisation. Through geostatistical tools such as kriging, these CL
can be interpolated into a spatially continuous map of nutrient nitrogen critical loads for Canada.
5.6 Management Strategies
The effects of eutrophication, acidification, atmospheric deposition, land-use, nutrient inputs
from land-based sources; sulphur, nitrogen, and phosphorus, are so inextricably tied that it is
necessary to manage them using integrated approaches. Strategies to manage, limit and mitigate
the effects of RN deposition also need to account for and manage non-atmospheric RN sources.
Many of these concepts and tools of cumulative and integrated effects management strategies,
specifically the use of scoping, indicators, zones of influence and VECs – are recommended for
this type of complex management scenario. The concepts of R-SEA may be of special utility
because of the goal to reach a desirable rather than inevitable management end-point.
It is important to decide what outcomes are desirable and at what point change becomes
“unacceptable”. Also, baseline conditions must be defined before “change” can be determined.
Baselines can be formed from a combination of paleoecology and historical data records. Terms
like “desirable” and “acceptable” are subjective with little scientific meaning. However, these are
terms that need to be defined in the context of RN for policy and management purposes. The
definition of these terms needs to be the result of well-balanced decisions that take into account
environmental, economic and social consequences in light of the well-being of all Canadians.
First Nations values, wildlife values, urban values, parks and protected areas may have different
notions of change and what change is acceptable. In this way the concepts of cumulative and
integrated effects – that in theory explore the environmental, social and economic consequences
of an action – are especially relevant.
Zoning has been used as a tool of cumulative effects management in a number of instances
including New Jersey, US (Conway and Lanthrop 2005) and Xiamen, China (Xue et al. 2005).
In terms of integrated atmospheric emissions management in Canada, zones can be created at a
federal level with emission intensity caps or management strategies particular to that zone. Each
zone’s emission allowance is based on a balance of social, economic and environmental factors.
Management is based on particular ecosystem- or human health-based objectives. Models to limit
acidification and eutrophication effects can be applied across zones using acidification or
eutrophication factors as in Bellekom et al. (2009) for separate European Nations. This is an
elegant and simple way of arranging a federal emission strategy and existing land-use zoning can
be utilised. For instance, a protected area should not be located in an air zone influenced by large
individual (or numerous small) emission sources. Similarly residential areas should meet health
68
objectives, but an industrially zoned area my have higher emission and deposition allowances.
The number of zones can be set at a feasible and relevant number, with boundaries, thresholds,
loads and caps based on an interdisciplinary and diverse set of objectives.
Critical Loads represent the most well accepted method of managing nitrogen deposition. CL
methods can be viewed like an indicator of cumulative (or integrated) effects management, and
could be developed for each emissions zone. Methods such as the relatively simple IROWC-N
model can also be used much as it was in De Jong et al. (2009), but with more accurate
deposition and runoff off estimates through methods discussed above. Stable isotopes can be used
to potentially identify sources of atmospheric RN from plants tissue, soil or water samples.
Knowing the source pollutants and the range of pollutant transport can aid in the development of
management strategies. Similarly dispersion model backcasts can be used in sourceapportionment.
The scope or resolution of any assessment needs to be defined. Due to Canada’s size and
diversity, it is not appropriate to manage RN in a uniform manner across its expanse. Similarly,
provincial boundaries have no ecological relevance. Zones are also useful in partitioning the
country into manageable units. Ideally zones would represent airsheds. In defining the boundaries
of an airshed two approaches can be used – one is source-oriented and the other receptor-oriented
frames or reference (Ainsle and Jackson 2009). Atmospheric models are very useful tools in
airshed boundary definition. However, it may be practical to define zones or emission
management units at a larger scale.
A number of tools can be used as part of integrated or cumulative management. Models can aid
in integrated RN management. For example INCA-N (Integrated Catchment Model of Nitrogen)
has been used across Europe in assessing nitrogen transport into watersheds, and in the UK was
utilised alongside an economic optimisation framework to develop a policy for minimising
agricultural N-runoff (O’Shea and Wade 2009). Similarly, linear models linking modelled
deposition with emission inventories have been used at the European-scale and can be applied to
Canada with zoning divisions under which emissions are managed. Methods from Bellekom et al.
(2009) provide such a policy-based tool for emissions management. The WATERSN model may
be used to assess management strategies for decreased N input to ecosystems for marine of
freshwater, forests and estuaries using either measured or modelled RN inputs and natural
measured background RN levels. WATERSN has been used in a variety of application including
Canada’s Oceans Act (Borja et al. 2008) to approach issues of ecosystem integrity with an
holistic ecosystem- rather than species-based approach to management.
Strategies for successful eutrophication management include the reduction of atmospheric RN
emissions and nutrient loadings of N and P. Nutrient loadings to ecosystems can be reduced by
upgrading sewage systems, and diverting storm-water to prevent untreated sewage from entering
surface waters. Other advances in the management of RN include reductions in N and P input
from the discharge of municipal and rural wastewaters and advanced phosphorus removal by
municipal wastewater treatment plants before discharging their wastes to water systems.
Although phosphate limits in detergents and cleaners have helped to limit the amount of
phosphorus entering Canada’s ecosystems, natural freshwater ecosystems are P-limited
throughout most of Canada and therefore in much of the country continued phosphorus
69
management is necessary. It is not sufficient to just reduce atmospheric RN; rather strategies need
to integrate the management of all RN and P sources reaching an ecosystem.
The efficient application of chemical fertilisers and manure can reduce N loss to volatilisation
and leaching in agriculture and forestry. The nutrient requirements of crops must be balanced by
the supply of nutrients to the crops from the soil and from fertilisers. It has been estimated that
only 50% of synthetic RN from fertilisers is retained by the crops or soil (Niemi et al. 2009).
Most provinces in Canada have guidelines for manure application to soils, typically based on
nitrogen application rates. Nutrient management strategies (e.g., transporting surplus manure
from animal producers to crop farms) will improve farmers’ abilities to manage nutrients more
effectively, with the ultimate aim of reducing over-fertilisation.
In areas of intensive livestock production, treatment of animal waste could reduce the risk of
contamination of surface water and groundwater by manure. Crop rotation between efficient
users of nitrogen such as grasses, with less efficient users of nitrogen such as potatoes, can help
to limit NO 3 - leaching from crops (Farm Centre 1998). In addition rotation with nitrogen fixing
leguminous species like soy or clover, helps reduce the need for chemical nitrate fertilisers with
an affinity for leaching. Similarly, in aquaculture operations, 70-80% nutrients added are lost to
the environment as metabolic waste, faeces, and uneaten food fragments (Environment Canada
2001). The development of more nutritionally balanced and digestible feed can reduce waste
discharges from feeding. Environmental impacts associated with nutrient loss in aquaculture
operations could be reduced by placing cages away from sensitive waters and shorelines,
collecting and treating wastewater, and the replacement of open-net techniques with contained
units.
Forest harvesting can have a large effect on lake and soil chemistry by the removal of both base
cations and (organic) nitrogen pools from a basin. A significant reduction of critical loads can
occur from forest harvesting due to the removal of base cations that neutralise incoming acidity.
The amount of tree removed (i.e. whole stem versus debris retention) impacts the amounts of
base cations removed (Watmough et al. 2003; Watmough and Dillon 2003). On the other hand,
the removal of nitrogen and planting of young fast growing trees over older slow growing ones,
can improve RN uptake and reduce the onset of RN saturation and subsequent leaching. In this
way, forest management practices can be used to minimise base cation losses (leaving branches
and debris) and maximise RN uptake to limit the unwanted effects of both eutrophication and
acidification from atmospheric nitrogen deposition. Fenn et al. (1998) recommended maximising
N retention in managed forests by avoiding excess stand maturation, intensified harvesting,
fertilisation with P or other nutrients to overcome N limitation, and designating riparian areas as
traps of N and P runoff.
Numerous studies have riparian forests to be very effective in the removal of RN. Riparian
vegetation can remove 80-98% of NO 3 - from groundwater by processes of uptake and
denitrification, the latter of which removes the nitrate permanently (Hill 1996).Similarly,
wetlands provide excellent sinks for NO 3 - in groundwater and surface runoff, but should be used
with care to avoid wetland eutrophication (Arheimer and Wittgren 1994; Jansson et al. 1994).
The use of riparian, wetland and estuarine systems as part of eutrophication management should
be explored further.
70
Critical loads are useful tools in determining reduction targets of RN emissions and RN loadings
for the recovery of affected ecosystems; however these tools must simultaneously account for
atmospheric and non-atmospheric emissions of RN. Additionally, the loading estimates need to
be derived using comprehensive and accurate inventories of atmospheric emissions and the
surface (waster or land) application of sewage, biosolids, fertilisers and animal wastes. For
instance, critical loads have proven to be a truly effective management tool in Europe where very
large exceedances (CL(EX)) no longer occur, since CL were used to determine N-emission
ceilings and develop integrated policies for regulating N flows. Some of these policies include:
the creation of lower exhaust fuels, changes in automobile engines, and regulations in fertiliser
use, imports, and N levels (van Egmond et al. 2002).
While CL have proven an effective deposition management tool globally, there is generally a
high level of uncertainty associated with exceedance estimates due to numerous factors including
the choice of constants and limits used in calculations, to a lack of adequate deposition
monitoring (Aherne 2008a; Aherne 2008b). Tools from risk management may also provide
reasonable management solutions. The idea of risk accommodates numerous outcomes, whereas
critical loads utlise a single concrete threshold or limit creating more opportunity for uncertainty.
The potential for numerous outcomes is helpful in optimising an RN management program
through the consideration of environmental, social and economic goals. R-SEA techniques can be
used to ensure the desired outcomes. Following the optimisation of atmospheric and nonatmospheric RN reduction programs, various methods can be used to see the reduction through to
fruition. Tax incentives employed for reducing N loading in agricultural regions for instance has
been successful in the UK (O’Shea and Wade 2009).
6.0 Conclusion and Recommendations
It is recommended that a representative Canada-wide RN monitoring and management program
include these main elements:
1) A realistic quantification of RN deposition
2) A realistic quantification of non-atmospheric RN and P inputs
3) A description of what is being protected
4) A description of the vulnerability to adverse effects of what is being protected
5) Knowledge of adverse effects already occurring
6) Strategies for impact mitigation
The point at which N saturation is occurs is important in determining when RN changes from
being a growth promoter, to an acidifier – a process still unknown in Canadian ecosystems. In
Norway it was found that N saturation did not occur at wet nitrate (NO 3 -) depositions of <10
kg/ha/year (Dise and Wright 1995). Similarly, Forsius et al. (1996) reported from Europe that
reductions of deposition to 8-10 kg NO 3 /ha/year are needed to avoid saturation, and even lower
thresholds if high levels of RN contents are already present in the soil.
From the maps of residual soil N (Figure 12) it appears that much of southern Canada is at risk of
effects from RN enrichment through various source inputs. Areas such as southern Ontario and
Quebec, also receive the highest atmospheric depositional load (Figure 10). Although they are not
71
thought to be the most sensitive ecosystems to acidification, the high RN loads in these areas
parts of northern Alberta, and the southern Prairies and Maritimes) puts them at risk of changes
due to nitrogen enrichment. In southwest British Columbia, marine vessel emissions and
aquaculture contribute to the high RN loads from the transportation and agricultural sectors.
According to present knowledge, these parts of Canada receive the highest RN loads, however
their sensitivity to eutrophication and the quantification of totals RN loadings, remains to be seen.
Any management strategy used needs to account for the diversity of RN sources and chemical
forms; and also the diversity of Canadian ecosystems and economies across the country.
Atmospheric RN deposition comes from a variety of transportation and industrial based sources
across Canada. The form of the RN deposition, whether it be modelled or measured is
particularly important in determining how ecosystems will respond. It is recommended that
emission zoning be explored as a potential tool to manage RN emissions and account for this
variability.
Since eutrophication is inherently tied to non-atmospheric inputs of RN and the availability of
other nutrients such as phosphorus, integrated or cumulative approaches are recommended for of
the assessment and management of RN deposition and effects. Additionally, because RN can lead
to acidification it is strongly tied to atmospheric emissions of sulphur, integrated management
should include a representation of multiple emission types and sources reaching ecosystems via
air, land or water. Despite the unknowns and uncertainty surrounding RN deposition and effects
in Canada, use of existing information and monitoring networks, combined with ecosystem
monitoring, will shed light on the sensitivity of Canadian ecosystems to atmospheric RN
facilitated eutrophication and its related effects.
72
7.0 References
Aber, J.D., Nadelhoffer, K.J., Steudler P., and Melillo, J.M.., 1989. Nitrogen saturation in
northern forest ecosystems. BioScience 39: 378-386.
Aber, J.D., Magill, A., Boone, R., Melillo, J.M., Steudler, P., and Bowden, R., 1993. Plant and
soil responses to chronic nitrogen additions at the Harvard Forest Massachusetts.
Ecological Applications 3(1): 156-166.
Aber J.D., Olinger S.V., and Driscoll, C.T., 1997. Modeling nitrogen saturation in forest
ecosystems in response to land use and atmospheric deposition. Ecological Modeling 101:
61-78.
Aber, J., McDowell, W., Nadelhoffer, K., Magill, A., Bernston, G., Kamakea, M., McNulty, S.,
Currie, W., Rustad, L., and Fernandez, I., 1998. Nitrogen Saturation in Temperate Forest
Ecosystems: Hypothesis revisited. Bioscience 48(11): 921-934.
Aber, J.D., Goodale, C.L., Ollinger, S.V., Marie-Louise, S., Magill, A.H., Martin, M.E., Hallett,
R.A., and Stoddard, J.L., 2003. Is nitrogen deposition altering the nitrogen status of
northeastern forests? BioScience. 53(4): 375-389.
Achermann B., and Bobbink R., 2003. Empirical critical loads for nitrogen. Expert Workshop.
Beme, Switzerland. 11-13 November 2002. Proceedings. SAFEL – Swiss Agency , for
the Environment, Forests and Landscape. Environmental Documentation 164.
Aherne J., 2008a. Critical Load and Exceedance Estimates for Upland forest Soils in Manitoba
and Saskatchewan: Comparing Exceedance Outputs of Different Models. Final Report,
Environmental and Resource Studies, Trent University, Peterborough, Ontario.
Aherne J., 2008b. Calculating Critical Loads of Acid Deposition for Forest Soils in Alberta:
Critical load, exceedance and limitations. Final Report, Environmental and Resource
Studies, Trent University, Peterborough, Ontario.
Aherne J., Posch M., Dillon P.J., and Henriksen A., 2004. Critical loads of acidity for surface
waters in south-central Ontario, Canada: regional application of the First-order Acidity
Balance (FAB) model. Water, Air, and Soil Pollution: Focus 4: 25–36.
Aherne J., and Watmough S., 2006. Calculating Critical Loads of Acid Deposition for Forest
Soils in Manitoba and Saskatchewan: Data Sources, Critical Load, Exceedance and
Limitations. Final Report prepared for CCME. Environmental and Resource Studies,
Trent University, Peterborough, ON.
Ainslie, B., and Jackson, P.L., 2009. The use of an atmospheric dispersion model to determine
influence regions in the Prince George, B.C. airshed from the burning of open wood
waste piles. Journal of Environmental Management 90: 2393-2401.
73
Alberta Environment, 2008. Alberta Acid Deposition Management Framework. 22pp.
Anderson, D.M., Burkholder, J.M., Coclan, W.P., Gilbert, P.M., Gobler, C.J., Heil, C.A., Kudela,
R.M., Parsons, M.L., Rensel, J.E.J., Townsend, D.W., Trainer, V.,L., and Vargo, G.A.,
2008. Harmful algal blooms and eutrophication: Examining linkages from selected
coastal regions of the United States. Harmful Algae 8: 39-53.
Arheimer B., and Wittgren H.B., 1994. Modelling the effects of wetlands on regional nitrogen
transport. Ambio 23: 378-386.
Avila-Segura, M., Barak, P., Posner, J.L., and Laird, D.A., 2002. Coupling of H+ and N cycles on
several common agroecosystems of Midwestern United States. Wisconsin Integrated
Cropping Trial Project. URL: http://www.cias.wisc.edu/wicst/pubs/acidify.htm [Accessed
January 2010]
Baron, J.S., Bowman, W.D., Sverdrup, H., Driscoll, C.T., Stoddard, J.T., and Hartman, M.D.,
2009. Acidification and eutrophication of temperate North American and European
ecosystems from atmospheric deposition. Climate Change: Global Risks, Challenges and
Decisions. IOP Conf. Series: Earth and Environmental Science 6 (2009) 462012.
doi:10.1088/1755-1307/6/6/462012
Bellekom, S., Hettelingh, J-P., and Aben, J., 2009. Spatial aspects affecting acidification factors
in European acidification modelling. Environmental Modelling and Software 24: 463472.
Belzer W., Evans C., and Poon A., 1997. Atmospheric Nitrogen Concentrations in the Lower
Fraser Valley. Environment Canada, Aquatic and Atmospheric Science Division,
Vancouver, B.C. DOE FRAP 1997: 1–23.
Berendse E., van Breeman N., Rydlin H., Buttler A., Heijmans M., Hoosbeek M.R., Lee J.A.,
Mitchell E., Saarinen T., Vasander H., and Wallen B., 2001. Raised atmospheric CO 2
levels and increased N deposition cause shifts in plant species composition and
production in Sphagnum bogs. Global Change Biology 7: 591-598.
Bleeker, A.G., Draaijers, G., van der Veen, D., Erisman, J.W., Mols, H., Fonteijn, P., and
Geusebroek, M. 2003. Field intercomparison of throughfall measurements performed
within the framework of the Pan European intensive monitoring program of EU/ICP
forest. Environmental Pollution 125: 123-138.
Bobbink R., Ashmore M., Braun S., Fluckiger W., and van den Wyngaert, I., 2002. Empirical
nitrogen critical loads for natural and semi-natural ecosystems: 2002 update.
Environmental Documentation 164, Swiss Agency for the Environment, Forests and
Landscape CLC 2000 - Corine Land Cover, CEEA, Copenhagen, 2006.
Bobbink, R., and Roelofs, I.G.M., 2005. Nitrogen Critical Loads for Natural and Semi-Natural
Ecosystems: the Empirical Approach. Water, Air and Soil Pollution 85: 2413-2418.
74
Borja, A., Bricker, S.B., Dauer, D.M., Demetriades, N.T., Ferreira, J.G., Forbes, A.T., Hutchings,
P., Jia, X., Kenchington, RMarques, J.C., and Zhu, C., 2008. Overview of integrative
tools and methods in assessing ecological integrity in estuarine and coastal systems
worldwide. Marine Pollution Bulletin, 56: 1519-1537
Bormann, B.T., Bormann, F.H., Bowden, W.B., Piece, R.S., Hamburg, S.P., Wang, D., Snyder,
M.C., Li, C.Y., and Ingersoll, R.C., 1993. Rapid N 2 Fixation in Pines, Alder, and Locust:
Evidence From the Sandbox Ecosystems Study 74(2): 583-598.
Bouwman A.F., van Vuuren, D.P., Derwent R.G., and Posch M., 2002. A global analysis of
acidification and eutrophication of terrestrial ecosystems. Water, Air and Soil Pollution.
141: 349-382.
Brandes, J. A., and Devol, A. H., 2002. A global marine-fixed nitrogen isotopic budget:
Implications for Holocene nitrogen cycling. Global Biogeochemical Cycles 16(4): 1120.
Britto, D.T., and Kronzucker, H.J., 2002. NH 4 + toxicity in higher plants: a critical review. Journal
of Plant Physiology 159(6): 567-584.
Brockley, R.P., and Sanborn, P., 2003. Effects of Sitka alder in the growth and foliar nutrition of
young lodgepole pine in the central interior of British Columbia. Canadian Journal of
Forest Research 33: 1761-1771.
Burke, M., Oleson, K., McCullough, E., Gaskell, J., 2009. A Global Model Tracking Water,
Nitrogen, and Land Inputs and Virtual Transfers from Industrialized Meat Production and
Trade. Environmental Modeling and Assessment 14(2): 1573-2967.
Burns D.A., 2004. The effects of atmospheric nitrogen deposition in the Rocky Mountains of
Colorado and southern Wyoming, USA – a critical review. Environmental Pollution
127(2): 257-269.
Butzler, J.M., and Chase, J.M., 2009. The effects of variable nutrient additions on a pond
mesocosm community. Hydrobiologia 616: 65-73.
Bytnerowicz, A., Dueck, T., and Godzik, S., 1998. Nitrogen Oxides, Nitric Acid Vapour, and
Ammonia. In: Recognition of Air Pollution Injury to Vegetation – A Pictorial Atlas, ed.
R. B. Flager. Air and Water Management Association, Pittsburgh, Pennsylvania.
Bytnerowicz, A., Sanz, M.J., Arbaugh, M.J., Padgett, P.E., Jones, D.P., and Davila, A., 2005.
Passive sampler monitoring ambient nitric acid (HNO 3 ) and nitrous acid (HNO 2 )
concentrations. Atmospheric Environment 39: 2655-2660.
Bytnerowicz, A., Omasa, K., and Paoletti, E., 2007. Integrated effects of air pollution and climate
change on forests: A northern hemisphere perspective. Environmental Pollution 147: 438445.
75
Bytnerowicz, A., Fraczek, W., Schilling, S., and Alexander, D., 2010. Spatial and temporal
distribution of ambient nitric acid and ammonia in the Athabasca Oil Sands Region,
Alberta. Journal of Limnology 69 (Suppl. 1): 11-21. DOI: 10.3274/JL10-69-S1-03.
Calderón, S, M., Poor, N.D., and Campbell, S.W., 2007. Estimation of the particle and gas
scavenging contributions to wet deposition of organic nitrogen. Atmospheric
Environment 41(20): 4281-4290.
Canada, 2007. Notice with respect to substances in the National Pollutant Release Inventory for
2007. Supplement: Department of the Environment, published under section 71(1)(b) of
the Canadian Environmental Protection Act. Canada Gazette 141(9), Part 1. Ottawa, ON,
March 3, 2007.
Canadian National Atmospheric Chemistry Precipitation Database, 2006. NAtChem/Precipitation
Database - Station Information, SAS Output. URL:
http://www.msc.ec.gc.ca/natchem/precip/metadata_e.html [Accessed February 2010].
Canadian National Atmospheric Chemistry Precipitation Database, 2007. NAtChem Database.
Precipitation Chemistry and Particulate Matter. User Login / Data Download. URL
http://www.on.ec.gc.ca/NatChem/Login/:Login.aspx [Accessed January 2010].
Carou, S., Dennis, I., Aherne, J., Ouimet, R., Watmough, S.A., DeMerchant, I., Shaw, M., Vet,
R., Bouchet, V., and Moran, M., 2008. A National Picture of Acid Deposition Critical
Loads for Forest Coils in Canada. Report prepared by Environment Canada for CCME.
Carroll J.A., Johnson D., Morecroft M., Taylor A., Caporn S.J.M. and Lee J.A., 2000. The effect
of long-term nitrogen additions on the bryophyte cover of upland acidic grasslands.
Journal of Bryology 22: 83–89.
CCE (Coordination Centre for Effects). 2007. Progress report 201pp. URL:
http://www.rivm.nl/bibliotheek/rapporten/500090001.pdf [Accessed Janruary 2010].
CCME, 1999. Canadian National Air quality Objectives: Process and Status. In: Canadian
environmental quality guidelines, 1999, Canadian Council of Ministers of the
Environment, Winnipeg.
CCME, 2003. Canadian water quality guidelines for the protection of aquatic life: Nitrate Ion. In:
Canadian environmental quality guidelines, 1999, Canadian Council of Ministers of the
Environment, Winnipeg.
CCME, 2004. Canadian water quality guidelines for the protection of aquatic life: Phosphorus:
Canadian Guidance Framework for the Management of Freshwater Systems. In: Canadian
environmental quality guidelines, 2004, Canadian Council of Ministers of the
Environment, Winnipeg, MB.
CCME, 2005. Five-year review of the Canada-wide Acid Rain Strategy for Post-2000.
Federal/provincial/Territorial Ministers of Energy and Environment. PN 1353 23pp.
76
CCME, 2008a. 2006-2007 Progress Report on the Canada-wide Acid Rain Strategy for Post2000. Federal/provincial/Territorial Ministers of Energy and Environment. PN 1410 40pp.
CCME, 2009. Regional Strategic Environmental Assessment in Canada: Principles and
Guidance. Canadian Council of Ministers of the Environment, Winnipeg, MB.
CCME, 2010. Who belongs to CCME and how do they set priorities. URL: www.ccme.ca/about
[Accessed May 2010].
CEMA, 2008. Proposed Interim Nitrogen (Eutrophication) Management Recommendations and
Work Plan for the Regional Municipality of Wood Buffalo area. Prepared by the NO x SO 2
Management Working Group of the Cumulative Effects Management Association
(http://www.cemaonline.ca/content/view/78/192/)
Chambers P.A., 1996. Nutrient enrichment in the Peace, Athabasca and Slave rivers: assessment
of present conditions and future trends. Synthesis Report No. 4. Edmonton: Environment
Canada and Alberta Environmental Protection, Northern River Basins Study.
Chambers P.A., Guy M., Roberts E.S., Charlton M.N., Kent R., Gagnon C., Grove G., and Foster
N., 2001. Nutrients and their impact on the Canadian environment. Agriculture and AgriFood Canada, Environment Canada, Fisheries and Oceans Canada, Health Canada and
Natural Resources Canada, Ottawa. 241 p.
Cole D.W., Gessel S.P., and Dice S.F., 1967. Distribution and cycling of nitrogen, phosphorus,
potassium, and calcium in a second-growth Douglas-fir ecosystem. In: Young, H.E. ed.
Symp. On Primary Productivity and Mineral Cycling in Natural Ecosystems. University.
of Maine Press, Orono, ME.
Colvile, R.N., Hutchinson, E.J., and Warren, R.F., 2002. The transport sector as a source of air
pollution. In: Air Pollution Science for the 21st Century, Developments in Environmental
Science, Volume 1. Austin, J., Brimblecome, P., and Sturges, W. (eds.). Elsevier Science
Ltd., Oxford, UK. Pp. 1537-1565.
Conway, T., and Lathrop, R.G., 2005. Alternative land use regulations and environmental
impacts: assessing future land use in an urbanizing watershed. Landscape and Urban
Planning 71:1-15.
Cosby, B. J., Ferrier, R. C., Jenkins, A., and Wright, R. F., 2001. Modelling the effects of acid
deposition: refinements, adjustments and inclusion of nitrogen dynamics in the MAGIC
model, Hydrology and Earth System Sciences 5: 499–517.
Crohn, D., 2004. Nitrogen mineralization and its importance in organic waste recycling. In:
Proceedings, National Alfalfa Symposium, 13-5 December, 2004, San Diego, Ca, UC
Cooperative Extension, University of California, Davis 95616.
77
Cunha A., Power S.A., Ashmore M.R., Green P.R.S., Haworth B.J., and Bobbink, R. 2002.
Whole ecosystem nitrogen manipulation: an updated review. JNCC report series.
Curtis, L., Rea, W., Smith-Willis, P., Fenyves, E., Pan, Y., 2006. Adverse health effects of
outdoor air pollutants. Environment International 32: 815–830.
De Jong, R., Drury, C.F., Yang, J.Y., and Campbell, C.A., 2009. Risk of water contamination by
nitrogen in Canada as estimated by the IROWC-N model. Journal of Environmental
Management 90:3169-3181.
De Solla, S.R., Pettit, K.E., Bishop, C.A., Cheng, K.M., and Elliott, J.E., 2002. Effects of
agricultural runoff on native amphibians in the lower Fraser River Valley, British
Columbia, Canada. Environmental Toxicology and Chemistry 21: 353–360.
Del Grosso, S.J., Parton, W.J., Mosier, A.R., Hartman, M.D., Keough, C.A., Peterson, G.A.,
Ojima, D.S., and Schimel, D.S., 2001. Simulated effects of land use, soil texture and
precipitation on Ngas emissions using DATCENT. In: Follett, R.F., Hatfield, J.L. (Eds.)
Nitrogen in the Environment: Sources, Problems and Management. Amsterdam,
Netherlands. Pp. 413-432.
Delle Monache L., Deng X., Zhou Y., Modzelewski, H., Hicks, G., Cannon, T., Stull R.B, and di
Cenzo, C., 2007. Air Qulaity Ensemble Forecast Over the Lower Fraser Valley, British
Columbia. In: Borrego, C., and Norman A-L. (Eds.) Air Pollution Modeling and Its
Application XVII. Pp. 399-402.
Dennis, I.F., Clair, T.A., and Cosby, B.J., 2005. A new model validation approach to test the
MAGIC model in highly organic, low conductivity waters from Nova Scotia, Canada.
Environmental Modeling and Assessment 10(4): 303-314.
Dentener, F., Drevet, J., Lamarque, J.F., Bey, I., Eickhout, B., Fiore, A.M., Hauglustaine, D.,
Horowitz, L.W., Krol, M., Kulshrestha, U.C., Lawrence, M., Galy-Lacaux, C., Rast S.,
Shindell, D., Stevenson, D., Van Noije, T., Atherton, C., Bell, N., Bergman, D., Butler,
T., Cofala, J., Collins, B., Doherty, R., Ellingsen, K., Galloway, J., Gauss, M., Montanaro,
V., Müller, J.F., Pitari, G., Rodriguez, J., Sanderson, M., Solmon, F., Strahan, S., Schultz,
M., Sudo, K., Szopa, S., and Wild, O., 2006. Nitrogen and sulfur deposition on regional
and global scales: A multimodel evaluation. Global Biogeochemical Cycles 20: GB4003,
doi:10.1029/2005GB002672.
DFO, 2005. Fact Sheet - Why Farm Fish and Seafood? Department of Fisheries and Oceans,
Canada. URL: http://www.dfo-mpo.gc.ca/aquaculture/sheet_feuillet/seafood-fruitdemereng.htm [Accessed March 2010].
Dise, N.B., Rothwell, J.J., Gauci, V., van der Salm, C., and de Vries, W., 2009. Predicting
dissolved inorganic nitrogen leaching in European forests using two independent
databases. Science of the Total Environment 407: 1798-1808.
78
Domene, L.A.F., and Ayres, R.U., 2000. Nitrogen’s Role in Industrial Systems. INSEAD, R&D,
Cedex, France, 73pp.
Dow C.L., and Dewalle D.R., 1997. Sulfur and nitrogen budgets for five forested Appalachian
plateau basins. Hydrological Processes 11:801-816.
Downing, J., Hettelingh, J.P., and de Smet, P., 1993. Calculation and Mapping of Critical Loads
in Europe: Status Report 1993. Coordination Center for Effects. RIVM, The Netherlands.
Driscoll C.T., Lawrence G.B., Bulger A.J., Butler T.J., Cronan C.S., Edgar C., Lambert K.F.,
Likens G.E., Stoddard J.L., and Weathers K.C. 2001. Acidic deposition in the
northeastern United States: sources and inputs, ecosystem effects, and management
strategies. BioScience 51(3): 180-198.
Drury, C.F., Yang, J.Y., De Jong, R., Huffman, T., Kirkwood, V., Yang, X.M., and Reid, K.,
2005. Nitrogen Use Efficiency. In: Environmental Sustainability of Canadian Agriculture.
Agri-Environmental Indicator Report Series. Report #2. Agriculutre and Agri-Food
Canada. Pp. 69-75.
Dupont J., Choate J., Clair T.A., Estabrook R., Fredette C., Hilborn J., Jeffries D.S., Kahl J.S.,
Kellogg J.H., Kemp A., Khan H., Stacey P., Taylor D., Thompson R., and van Arsdale A.,
2000. Is nitrogen deposition a serious issue. Report for the Acid Rain Steering Committee
by Water quality monitoring workgroup, Conference of New England Governors and
Eastern Canadian Premiers. 22pp.
Dupont, J., Clair, T.A., Gagnon, C., Jeffries, D.S., Kahl, J.S., Nelson, S.J., and Peckenham, J.M.,
2005. Estimation of Critical Loads of Acidity for Lakes in Northeastern United States and
Eastern Canada.
Dupre, S., 2009. Annex 8: Canada’s Greenhouse Gas Emission Tables, 1990-2007. In:
Environment Canada, National Inventory Report, 1990-2007. Greenhouse Gas Sources
and Sinks in Canada. The Canadian Government’s Submission to the UN Framework
Convention on Climate Change, 661 pp. ISBN: 978-1-100-12999-0.
Edwards, G.C., and Ogram, G.L., 1986. Eddy correlation measurements of dry deposition fluxes
using a tunable diode laser absorption spectrometer gas monitor. Water, Air and Soil
Pollution 30(1-2): 187-194.
Eilers J.M., Rose C.L., and Sullion T.J., 1994. Status of air quality and effects of atmospheric
pollutants on ecosystems in the Pacific Northwest region of the National Park Service.
Final report prepared for the Air quality Division, National Park Service. 33 pp.
Emmett B., 2006. National Soil Monitoring Network: Review and assessment study. Final
Report. Centre for Ecology and Hydrology, 42pp. (CEH Project Number: C03179, Project
LQ09) (Unpublished).
79
Environment Canada, 2001. Nutrients in the Canadian Environment – Reporting on the state of
Canada’s Environment. 90 pp.
Environment Canada, 2006a. Preliminary literature review on agricultural ammonia abatement
techniques. National Agri-Environmental Standards Initiate (NAESI) Report No. 2-9.
16pp.
Environment Canada, 2006b. Preliminary literature review on atmospheric ammonia. National
Agri-Environmental Standards Initiate (NAESI) Report No. 2-10. 23pp.
Environment Canada, 2008a. Nation Air Pollution Surveillance (NAPS) Network. Annual Data
Summary for 2005 and 2006. Environmental Science and Technology Centre,
Environment Canada 367pp.
Environment Canada, 2008b. Canadian Environmental Sustainability Indicators 2008. Freshwater
Quality Indicator: Data Source and Methods 39pp.
Environment Canada, 2009a. 2007 Total Air Pollutants Emissions for Canada and
Provinces/Territories Version 1. National Pollutant Release Inventory Downloadable
Datasets. URL: https://www.ec.gc.ca/inrp-npri/Default.asp?lang=En&n=0EC58C98-1
[Accessed January, 2010].
Environment Canada, 2009b. 1985-2007 Historical NO x Emissions for Canada. Version 2.
National Pollutant Release Inventory Downloadable Datasets. URL:
https://www.ec.gc.ca/inrp-npri/Default.asp?lang=En&n=F3B474E9-1#3-1-2
[Accessed January, 2010].
Environment Canada, 2010. National Pollutant Release Inventory (NPRI) 2008 Facility Data
Summary. URL: http://www.ec.gc.ca/inrp-npri/default.asp?lang=EN&n=CFA83D44-1
[Accessed February 2010].
Erisman, J.W., Brydges, T., Bull, K., Cowling, E., Grennfelt, P., Nordberg, L., Satake, K.,
Schneider, T., Smeulders, S., Van der Hock, K.W., Wisniewski, J.R., and Wisniewski, J.,
1998. Nitrogen, the confer-N-S, First international nitrogen conference 1998: Summary
statement. Environmental Pollution 102: 3-12.
Erisman J.W., and de Vries, W., 2000. Nitrogen deposition and effects on European forests.
Environmental Reiews 8: 65-93.
Erisman J.W., Grennfelt P., and Sutton M., 2003. The European perspective on nitrogen emission
and deposition. Environment International 29(2-3): 311-325.
Eshleman K.N., Morgan II R.P., Webb J.R., Deviney F.A., and Galloway J.N., 1998. Temporal
patterns of nitrogen leakage from mid-Appalachian forested watersheds: Role of insect
defoliation. Water Resources Research 34:2005-2116. 19
80
Eshleman, K.N., McNeil, B.E., and Townsend, P.A., 2009. Validation of a remote sensing based
index of forest disturbance using streamwater nitrogen data. Ecological Indicators 9: 476484.
Fangmeir A., Hadwiger-Fangmeir A., van der Eerden L.J.M., and Jager H.J., 1994. Effects of
atmospheric ammonia on vegetation - a review. Environmental Pollution 86, 43-82.
Farm Centre, 1998. Nitrate – Agricultural sources and fate in the environment – perspectives and
direction. Proceedings of a Workshop help April 17, 1998. URL: http://www.ccseswcc.nb.ca/publications/english/nitrate.pdf [Accessed January 2010]
Feller M.C., and Kimmins J.P., 1979. Chemical characteristics of small streams near Haney in
Southwestern British Columbia. Water Resources Research. 15(2): 247-258.
Fenn, M.A., Poth, M.A., Aber, J.D., Baron, J.S. , Bormann, B.T. , Johnson,, D.W. , Lemly, A.D..,
McNulty, S.G. , Ryan, D.F. , and Stottlemyer, R., 1998. Nitrogen excess in North
American ecosystems: Predisposing factors, ecosystem responses, and management
strategies. Ecological
Fenn, M.E., Poth M.A., Schilling S.L., and Grainger D.B., 2000. Throughfall and fog deposition
of nitrogen and sulfur at an N-limited and N-saturated site in the San Bernardino
Mountains, southern California. Canadian Journal of Forest Research 30: 1476-1488.
Fenn, M.E., Baron J.S., Allen E..B., Rueth H.M., Nydick K.R., Geiser L., Bowman, W.D.,
Sickman J.O., Meixner T., Johnson D.W., and Neitlich P., 2003. Ecological effects of
nitrogen deposition in the Western United States. Bioscience 53(4): 404-420
Fenn, M.E., Jovan, S., Yuan, F., Meixner, T., and Gimeno, B.S., 2008. Empirical and simulated
critical loads for nitrogen deposition in California mixed conifer forest. Environmental
Pollution: 155: 492-511.
Fenn, M.E., Sickman, J.O., Bytnerowicz, A., Clow, D.W., Molotch, N.P, Pleim, J.E., Tonnesen,
G.S., Weathers, K.C., Padgett, P.E., and Campbell, D.H., 2009. Methods for Measuring
Nitrogen Deposition Inputs in Arid and Montane Ecosystems of Western North America.
Chapter 8, In: Air Quality and Ecological Impacts, Relating Sources to Effects.
Developments in Environmental Science, Volume 9. Legge, A. (ed.). Elsevier Science
Ltd., Oxford, UK. Pp. 179-228.
Fichtner, K., and Schulze, E.D., 1992. The effect of nitrogen nutrition on growth and biomass
partitioning of annual plants originating from habitats of different nitrogen availability.
Oecologia 92: 236-241.
Findlay, D., 2003. Response of Phytoplankton Communities to Acidification and Recovery in
Killarney Park and the Experimental Lakes Area, Ontario. Ambio 32: 190-195.
81
Fomina, M.A., Alexander, I.J., Colpaert, J.V., and Gass, G.M., 2005. Solubilization of toxic
metal minerals and metal tolerance of mycorrhizal fungi. Soil Biology and Biochemistry
37: 851-866.
Forest Mapping Working Group, 2003. Assessment of Forest Sensitivity to Nitrogen and Sulfur
Deposition in New England and Eastern Canada. Conference of New England Governors
and Eastern Canadian Premiers. Forest Mapping Group Pilot Phase Report, 4pp.
Fraczek, W., and Bytnerowicz, A., 2007. Making Effective Use of Geostatistics. In: ArcUser
Online. July-September 2007, URL:
http:www.esri.com/news/arcuser/0807/geostatistics.html [Accessed February 2010].
Gahnstrom, G., Blomqvist, P., and Fleischer, S., 1993. Are key nitrogen fluxes changed in the
acidified aquatic ecosystems? Ambio 22: 318-324.
Galloway, J. N. , Aber , J. D., Erisman , J. W., Seitzinger , S. P., Howarth , R. W., Cowling , E. B.
and Cosby, B. J., 2003. The nitrogen cascade. BioScience 53(4): 341–356.
Gangolli, S.D., van der Brandt, P.A., Feron, V.J., Janzowsky, C., Koeman, J.H., Speijers, G.J.A.,
Spiegelhalder, B., Walker, R., and Wishnok, J.S., 1994. Nitrate, nitrate and N-nitroso
compounds. European Journal of Pharmacology 292: 1-38.
Geiser, L.H., and Neitlich, P.N., 2007. Air pollution and climate gradients in western Oregon and
Washington indicated by epiphytic macrolichens. Environmental Pollution 145: 203-218.
Gibson, J.J., Birks, S.J., Scott, K., Jeffries, D., Aherne, J., and Shaw, P., 2010. Site-specific
estimates of water yield applied in regional acid sensitivity surveys across western
Canada. Journal of Limnology 69 (Suppl. 1): 56-66. DOI: 10.3274/JL10-69-S1-03.
Gijsman, A.J., Oberson, A., Tiessen, H., and Friesen, D.K., 1996. Limited applicability of the
CENTURY model to highly weathered tropical soils. Agrononomy Journal 88: 894–903
Gooday, A.J., Jorissen, L.A., Levin, L.A., Middelburg, J.J., Naqvi, S.W.A., Rabablais, N.N.,
Scranton, M., and Zhang, J., 2009. Biogeosciences 6: 1707-1745.
Government of Alberta, 2004. A Preliminary Assessment of Carbon Dioxide and Nitrous Oxide
Emissions from Agricultural Soils in Alberta. URL:
http://www1.agric.gov.ab.ca/$department/deptdocs.nsf/all/aesa8424 [Accessed Jamuary
2010]
Government of Canada, 1996. The state of Canada’s environment - 1996. Ottawa, ON.
Graveland, J., van deWal, R., van Balen, J.H., van Noordwijk, A.J., 1994. Poor reproduction in
forest passerines from decline of snail abundance on acidified soil. Nature 368: 446-448.
Grennfelt, P., and Hultberg, E., 1986. Effects of nitrogen deposition on the acidification of
terrestrial and aquatic ecosystems. Water, Air and Soil Pollution 30:945-963.
82
Grier, C.C., and Cole, D.W., 1972. Elemental transport changes occurring during development of
a second-growth Douglas-fir ecosystem. In: Research on coniferous forest ecosystems:
first year progress in the Coniferous Forest Biome, US/IBP: Proceedings of a symposium;
1972 March 23-24; Bellingham, WA. Portland, OR: U.S. Department of Agriculture,
Forest Service, Pacific Northwest Forest and Range Experiment Station: 103-113.
Gundersen, P., 1999. Nitrogen status and impact of nitrogen input in forests – indicators and their
possible use in critical load assessment. Presented at Conference on Critical Loads,
Copenhagen, November 1999.
Gundersen, P.; Berg, P.; Currie, B.; Dise, W. S.; Emmett, B. A.; Gauci, V.; Holmberg, M.;
Kjonaas, O.J.; Mol-Dijkstra, J.; van der Salm, C.; Schmidt, I. K.; Tieterna, A.; Wessel, W.
W.; Vestgarden, L. S.; Akselson, C.; De Vries, W.; Forsius, M.; Kros, H.; Matzner, E.;
Moldan, F.; Nadelhoffer, K. J.; Nilsson, L.-O.; Reinds, G. J.; Rosengren, U.; Stuanes, A.
O.; Wright, R. F., 2006. Carbon - Nitrogen interactions in forest ecosystems. Final report.
Danish Centre for Forest, Landscape and Planning. Forest & Landscape Working Papers,
17. 61pp.
Gundersen, P., Sevel, L., Riis Christiansen, J., Vesterdal, L., Hansen, K., and Bastrup-Birk, A.,
2009. Do indicators of nitrogen retention and leaching differ between coniferous and
broadleaved forests in Denmark? Forest Ecology and Management 258: 1137-1146.
Hall, R.I., Leavitt, P.R., Quinlan, R., Dixit, A.S., and Smol, J.P., 1999. Effects of agriculture,
urbanization and climate on water quality in the northern Great Plains. Limnology and
Oceanography 44: 739–756.
Hanson H. (editor). 2008. “Acid deposition fact sheet.” In: “Encyclopedia of Earth. Eds. Cutler J.
Cleveland (Washington, D.C.: Environmental Information Coalition, National Council for
Science and the Environment). [First published in the Encyclopedia of Earth June 25,
2008; Last revised July 24, 2008; Retrieved January 11, 2009].
www.eoearth.org/article/Acid_deposition_fact_sheet
Harrison, W.G., 1978. Experimental measurements of nitrogen remineralization in coastal water.
Limnology and Oceanography 23: 684-694.
He, B., Oki, T., Kanae, S., Mouri, G., Komori, D., and Seto, S., 2009. Integrated biogeochemical
modelling of nitrogen load from anthropogenic and natural sources in Japan. Ecological
Modelling 220: 2325-2334.
Hegmann, G., C. Cocklin, R. Creasey, S. Dupuis, A. Kennedy, L. Kingsley, W. Ross, H. Spaling,
and D. Stalker. 1999. Cumulative Effects Assessment Practitioners Guide. Prepared by
AXYS Environmental Consulting Ltd. and the CEA Working Group for the Canadian
Environmental Assessment Agency, Hull, Québec. URL: http://dsppsd.pwgsc.gc.ca/Collection/En106-44-1999E.pdf [Accessed January 2010].
83
Heijmans, M.M.P.D., Klees, H., de Visser, W. and Berendse, F. 2002. Effects of increased
nitrogen deposition on the distribution of 15N-labeled nitrogen between Sphagnum and
vascular plants. Ecosystems 5: 500–508.
Heisler, J., Glibert, P.M., Burkholder, J.M., Anderson, D.M., Cochlan, W., Dennison, W.C.,
Dortch, Q., Gobler, C.J., Heil, C.A., Humphries, E., Lewitus, A., Magnien, R., Marshall,
H.G., Sellner, K., Stockwell, D.A., Stoecker, D.K., and Suddleson, M., 2008.
Eutrophication and harmful algal blooms: A scientific consensus. Harmful Algae 8: 3–13.
Henriksen, A., and Posch,, M., 2001. Steady-state models for calculating the critical loads of
acidity for surface waters. Waster, Air and Soil Pollution: Focus 1 375-398.
Hessen D.O., Hindar A., and Holtan G., 1997a. The significance of nitrogen runoff for
eutrophication of freshwater and marine recipients. Ambio 26: 312-320.
Hessen, D.O., Henriksen, A., Hindar, A., Mulder, J., Torseth, K., and Vagstad, N., 1997b. Human
impacts on the nitrogen cycle: a global problem judged from a local perspective. Ambio
26:321-325.
Hill, A.R., 1996. Nitrate removal in stream riparian zones. Journal of Environmental Quality 25:
743-755
Hole, L.R., Christensen, J.H., Ruoho-Airola, T., Tørseth, K., Ginzburg, V., and Glowacki, P.,
2009. Past and future trends in concentrations of sulphur and nitrogen compounds in the
arctic. Atmospheric Environment 43: 928-939.
Horvath, S.M., Bedi, J.F., and Dreshcer-Parks, D.M., 1986. Effects of peroxyacetyl nitrate alone
and in combination with ozone in healthy young women. Journal of the Air Pollution
Control Association 26: 265-270.
Houle, D., 2005. Effects on forests and soils. Chapter 5 In: 2004 Canadian Acid Deposition
Science Assessment. ISBN 0-662-68662-4. EC, Downsview, p.163-202 (available from
http://www.msc-smc.ec.gc.ca/saib/acid/acid_e.html)
INDITE, 1994. Impacts of Nitrogen Deposition In Terrestrial Ecosystems (INDITE) in the
United Kingdom. Report of the UK Review Group on Impacts of atmospheric nitrogen.
Prepared for the Department of the Environment. HMSO, London.
Innes, J.L., 1995. Influence of air pollution on the foliar nutrition of conifers in Great Britain.
Environmental Pollution 88: 183-192.
International Joint Commission, 2008. Canada-United States Air Quality Agreement, Progress
Report 2008. URL: http://www.epa.gov/airmarkets/progsregs/usca/docs/2008report.pdf
[Accessed February 2010].
84
Islam, M.S., 2005. Nitrogen and phosphorus budget in coastal marine cage aquaculture and
impacts of effluent loading on ecosystem: review and analysis towards model
development. Marin Pollution Bulletin. 50: 48-61.
Jansson, M., Andersson, R., Berggren, H., and Leonardson, L., 1994. Wetlands and lakes as
nitrogen traps. Ambio 23: 320- 325.
Jassal, R.S., Black, T.A., Cai, T., Ethier, G., Pepin, S., Brümmer, C., Nesic, Z., Spittlehouse,
D.L., and Trofymow, J.A., 2010. Impact of nitrogen fertilization on carbon and water
balances in a chronosequence of three Douglas-fir stands in the Pacific Northwest.
Agricultural and Forest Meteorology 150(2): 208-218.
Jeffries, D.S., 1990. Snowpack storage of pollutants, release during melting, and impact on
receiving waters. In: Norton SA, Lindberg S & Page AL (Eds) Soils, aquatic processes
and lake acidification, Acidic Precipitation Vol. 4 pp 107-132). Springer, NY.
Jefferies, R.L., and Maron, J.L., 1997. The embarrassment of riches” atmospheric deposition of
nitrogen and community ecosystem processes. TREE 12(2): 74-78.
Jeffries, D., and Ouimet, R., 2005. Chapter 8: Critical loads: Are they being exceeded? In: 2004
Canadian Acid Deposition Science Assessment. Environment Canada, Ottawa.
Jefferies, D.S., Semkin, R.G., Gibson, J.J., and Wong, I., 2010. Recently surveyed lakes in
northern Manitoba and Saskatchewan Canada: characteristics and critical loads of acidity.
Journal of Limnology 69 (Suppl. 1): 45-55.
Jenkin, M.E. and Clemitshaaw, K.C., 2002. Ozone and other secondary photochemical
pollutants” chemical processes governing their formation in the planetary boundary layer.
Chapter 11, In: Air Pollution Science for the 21st Century, Developments in
Environmental Science, Volume 1. Austin, J., Brimblecome, P., and Sturges, W. (eds.).
Elsevier Science Ltd., Oxford, UK. Pp. 285-388.
Johnson, D.W., Turner, J., and Kelly, J.M. 1982. The effects of acid rain on forest nutrient status.
Water Resources Research 18: 449-461.
Johnson, P.T.J., Chase, J.M., Dosch, K.L., Hartson, R.B., Gross, J.A., Larson, D.J., Sutherland,
D.R., and Carpenter, S.R., 2007. Aquatic eutrophication promotes pathogenic infection in
amphibians. Proceedings of the National Academy of Science 104: 17430-17434.
Jones, M.L.M., Oxley, E.R.B. and Ashenden, T.W. 2002. The influence of nitrogen deposition,
competition and desiccation on growth and regeneration of Racomitrium lanuginosum
(Hedw.) Brid. Environmental Pollution, 120: 371–378.
Jones, T., Evans, C., Freeman, C., Ostle, N., and Burden, A., 2007. The impact of changing
acidity on carbon and nitrogen cycling in peatlands. In: Proceedings of Peatlands and
Climate. Pp.201-204.
85
Jones, M.L.M, Norris, D., Hall, J, and Petit, S., 2009. Developing an indicator-modelling
approach to forecast changes in nitrogen critical load Exceedance across Europe arising
from agricultural reform. Ecological Indicators, In Press.
Kane, D.D., Gordon, S.I., Munawar, M., Charlton, M.N., and Culver, D.A., 2009. The Planktonic
Index of Biotic Integrity (P-IBI): An approach for assessing lake ecosystem health.
Ecological Indicators 9: 1234-1247.
Kaste O., and Dillon P.J., 2003. Inorganic nitrogen retention in acid-sensitive lakes in southern
Norway and southern Ontario, Canada – a comparison of mass balance data with an
empirical N retention model. Hydrological Processes 17(12): 2393-2407.
Karnosky, D.F., and Thakur, R.C., 2004. Genetic aspects of air pollution and climate change. In:
Burley, J., Evans, J., Youngquist, J. (eds.), Encyclopedia of Forest Sciences. Academic
Press, London, pp. 223-229.
Kelly, C.A., Rudd, J.W., and Schindler, D.W., 1990. Acidification by nitric acid - future
considerations. Water Air and Soil Pollution 50:49-61.
Kennedy, F. 2003. How extensive are the impacts of nitrogen pollution in Great Britain’s forests?
Forestry Commission of Great Britain, pp.66-75. URL:
http://www.forestresearch.gov.uk/pdf/FR0102nitro.pdf/$FILE/FR0102nitro.pdf
[Accessed January 2010].
Kivimaki, S.K., Bruneau, P., Grace, J. Leith, I.D., Rees, B., and Sheppard, L.J. 2007. Carbon and
nitrogen cycling in Sphagnum capillifolium in respect to nitrogen deposition and form. In:
Proceedings of Peatlands and Climate. Pp. 592-594.
Köchy, M., and Wilson, S.D., 2001. Nitrogen deposition and forest expansion in the northern
Great Plains. Journal of Ecology 89(5): 807-817.
Köchy, M., and Wilson, S.D., 2005. Variation in nitrogen deposition and available soil nitrogen
in a forest-grassland ecotone in Canada. Landscape Ecology 20: 191-202.
Korol, M., and Larivière, É., 1998. Chapter II – The Structure of the Canadian Fertilizer Industry.
In: Fertiliser Pricing in Canada. Agriculture and Agri-Food Canada. URL:
http://www4.agr.gc.ca/AAFC-AAC/display-afficher.do?id=117916606252&lang=eng
[Accessed February 2010].
Kowalchuk, G.A., and Stephen, J.R., 2001. Ammonia-Oxidizing Bacteria: A Model for
Molecular Microbiology. Annual Review of Microbiology 55: 485-529.
Krumlik, G.J., 1978. Comparative study of nutrient cycling in the Subalpine Mountain Hemlock
Zone of British Columbia. Ph.D. Thesis, Faculty of Forestry, University of British
Columbia, Vancouver, Canada.
86
Krupa, S.V., and Legge, A.H., 2000. Passive sampling of ambient, gaseous air pollutants: an
assessment from an ecological perspective. Environmental Pollution 107: 31-45.
Krups, S.V., 2002. Sampling and physico-chemical analysis of precipitation: a review.
Environmental Pollution 120: 565-594.
Krupa, S.V., 2003. Effects of atmospheric ammonia (NH 3 ) on terrestrial vegetation: a review.
Environmental Pollution, 124(2): 179-221.
Krzyzanowski, J., 2004. Ozone Variation with Height in a Forest Canopy - Results from a
Passive Sampling Field Campaign. Atmospheric Environment 38 (35): 5957-5962.
Krzyzanowski, J., 2009. The importance of policy in emissions inventory accuracy – a lesson
from British Columbia, Canada. Journal of the Air Waste Management Association 59:
430–439.
Krzyzanowski, J., 2010. Approaching Cumulative Effects through Air Pollution Modelling.
Water , Air and Soil Pollution.. In Press.
Laudon H., Clair T.A., and Hemond, H.F., 2002. Long-term response in episodic acidification to
declining SO 4 -2 deposition in two streams in Nova Scotia. Hydrology and Earth System
Sciences 6: 773-781.
Laungani, R., and Kmops, J.M.H., 2009. Species-driven changes in nitrogen cycling can provide
a mechanism for plant invasions, PNAS 106(30): 12400-12405.
Lawton, L.A., and Codd, G.A., 1991. Cyanobacterial (Blue-Green Algal) Toxins and their
Significance in UK and European Waters. Water and Environment Journal 5(4): 460-465.
Lillyman, C., Buset, K., Beaney, G., and Mullins, D., 2009. The Development of Regionally
Based Performance Standards for Agricultural Ammonia Emissions in the Context of
Particulate Matter Formation in Canada. National Agri-Environmental Standards
Initiative Synthesis Report No. 2, Environment Canada. Gatineau, Quebec. 80pp.
MacConnell, J.T., and Bond, G., 1957. Nitrogen Fixation in Wild Legumes. Annals of Botany
21: 185-192.
MacDonald, R.W., Barrie L.A., Bidleman T.F., Diamond M.L., Gregor D.J., Semkin R.G.,
Strachan W.M., Li Y.F., Wania F., Alaee M., Alexeeva L.B., Backus S.M., Bailey R.,
Bewers J.M., Gobeil C., Halsall C.J., Harner T., Hoff J.T., Jantunen L.M., Lockhart W.L.,
Mackay D., Muir D.C., Pudykiewicz J. Reimer K.J., Smith J.N., and Stern G.A. 2000.
Contaminants in the Canadian Arctic: 5 years of progress in understanding sources,
occurrence and pathways. The Science of Total Environment 254(2-3): 93-234.
Majumdar, D., 2009. Chapter 8: Nitrous Oxide from Crop Fields and Its Role in Atmospheric
Radiative Forcing. In: Singh, S.N. (Ed), Climate Change and Crops. Springer-Verlag
Berlin Heidelberg, 147-190.
87
Mandaville, S.M., 1999. Appendix A – A brief treatise on Eutrophication of Lakes. In:
Bioassessment of Freshwaters using Benthic Macroinvertebrates. URL:
http://chebucto.ns.ca [Accessed January 2010].
Marbek Resource Consultants Ltd., 2008. Assessment of Regulatory Quantification and
Reporting Requirements for Air Emissions from Industrial Facilities. Executive
Summary, submitted to CCME. 14pp.
Martin, J.L., Haya, K., Burridge, L.E., and Wildish, D.J., 1990. Nitzschia pseudodelicatissima – a
source of domoic acid on the Bay of Fundy, eastern Canada. Marine Ecology Progress
Series 67: 177-182.
Mayer, R., 1998. Soil acidification and cycling of metal elements: cause-effect relationships with
regard to forestry practices and climatic changes. Agriculture, Ecosystems and
Environment 67: 145-152.
Mills, K.H., Chalanchuk, S.M., and Allan, D.J., 2000. Recovery of fish population in Lake 223
from experimental acidification. Canadian Journal of Fisheries and Aquatic Sciences 57:
192-204
Moos, M.T., Laird, K.R., and Cumming, B.F., 2009. Climate-related eutrophication of a small
boreal lake in northwestern Ontario: a palaeolimnological perspective. The Holocene 19
(3): 359-367.
Moran, M., Makar, P., and Vet, R., 2009. Atmospheric transportation, transformation and
deposition of Nr affecting Canada. Workshop on Reactive Nitrogen, April 14-15,
Environment Canada, Downsview, ON.
Mosier A., and Kroeze C., 2000. Potential impact on the global atmospheric N 2 O budget of the
increased nitrogen input required to meet future global food demands. Chemosphere Global Change Science 2: 465-473
Moss, B., Beklioglu, M., Carvalho, L., Kilinc, S., McGowan, S., and Stephen, D., 1997.
Vertically-challenged limnology: Contrasts between deep and shallow lakes.
Hydrobiologia. 342: 257-267.
Mudie, P.J., Rochon, A., and Levac, E., 2002. Palynological records of red tide-producing
species in Canada: past trends and implications for the future. Palaeogeography,
Palaeoclimatology, Palaeoecology 180: 159-186.
National Research Council of Canada, 1981. Acidification in the Canadian Aquatic Environment:
Scientific criteria for assessing the effects of acidic deposition on aquatic ecosystems.
National Research Council of Canada, Associate Committee on Scientific Criteria for
Environmental Quality. NRCC No. 18475, 369pp.
88
Neff, J.C., Holland, E.A., Dentener, F.J., McDowell, W.H., and Russell, K.M., 2002. The origin,
composition and rates of organic nitrogen deposition: A missing piece of the nitrogen
puzzle. Biogeochemistry 57/58: 99-136.
NEG/ECP Forest Mapping Group (FMG), 2003. Assessment of forest sensitivity to nitrogen and
sulphur deposition in New England and Eastern Canada. FMP Pilot Phase Report 4pp.
NEG-ECP-FMG, 2008. Mapping forest sensitivity to atmospheric acid deposition. 2006-2007
Annual Report for the Committee on the Environment of the Conference of New England
Governors and Eastern Canadian Premiers. Acid Rain Action Plan, Action Item 4: Forest
Mapping Project 12 pp.
Nellemann, C., and Thomsen, M.G., 2001. Long-term changes in forest growth: Potential effects
of nitrogen deposition and acidification. Water, Air and Soil Pollution 128: 197-205.
Niemi, D., 2004. Emissions of pollutants related to acid deposition in North America, Chapter 2.
In: Morrison H.A. (Ed) Canadian acid deposition science assessment 2004 pp.
http://www.msc-smc.ec.gc.ca/saib/acid/assessment2004/toc_e.html#ch2002 Environment
Canada, Downsview, ON.
Niemi, D., Bittman, S., Drury, C.F., Grant, C., McGinn, S.M., and Rochette, P., 2009. Canadian
Nr Emissions. Workshop on Reactive Nitrogen, April 14-15, Environment Canada,
Downsview, ON.
Nilsson, J., and Grennfelt, P., 1988. Critical Loads for Sulfur and Nitrogen. Nordic Council of
Ministers, Copenhagen, Denmark. Pp. 31.
Nordbakken, J.F., Ohlson, M., and Högberg, P., 2003. Boreal bog plants: nitrogen sources and
uptake of recently deposited nitrogen. Environmental Pollution 126: 191–200.
Nordin, A.P., Hogberg, P., and Nasholm, T., 2001. Soil nitrogen form and plant nitrogen uptake
long a boreal forest productivity gradient. Oecologia 129:125-132.
Nordin, A., Strengbom, J., Witzell, J., Nasholm, T., and Ericson, L. 2005. Nitrogen deposition
and the biodiversity of boreal forests: implications for the nitrogen critical load. Ambio
34(1): 20-24.
NO x SO 2 Management Working Group of the Cumulative Environmental Management
Association, 2008. Proposed Interim Nitrogen (Eutrophication) Management
Recommendation and Work Plan for the Regional Municipality of Wood Buffalo Area,
May 12, 2008. 64pp.
Ogbebo, F.E., Evans, M.S., Waiser, M.J., Tumber, V.P., and Keating, J.J., 2009. Canadian
Journal of Fisheries and Aquatic Sciences 66: 247-260.
89
Ojima, D.S., and Baron, J.S. 1999. Modeling the timeline for acidification from excess nitrogen
deposition in Rocky Mountain National Park. Final completion report. DOI-NPS-National
Park Service 15pp.
O’Shea, L., and Wade, A., 2009. Controlling nitrate pollution: An integrated approach. Land Use
Policy 26: 799-808.
Ouimet, R., Arp, P.A., Watmough, S.A., Aherne, J., and DeMerchant, I., 2006. Determination
and mapping critical loads of acidity and exceedances for upland forest soils in eastern
Canada. Water, Air and Soil Pollution 172: 57-66.
Palmer, S. M., and Driscoll, C. T., 2002. Acid deposition: decline in mobilization of toxic
aluminium. Nature 417: 242–243.
Pardo, L. H. and C. T. Driscoll. 1996. Critical loads for nitrogen deposition: case studies at two
Piticairn, C.E.R, Leith, I.D., Fowler, D., Hargreaves, K.J., Moghaddam, M., Kennedy, V.H., and
Granat, L., 2001. Foliar nitrogen as an indicator of foliar nitrogen deposition and critical
loads on a European scale. Water, Air and Soil Pollution 130: 1037-1042.
Posch, M., Forsius, M., Johansson, M., Vuorenmaa, J., and Kamari, J., 2003. Modelling the
recovery of acid sensitive Finnish headwater lakes under present emission reduction
agreements. Hydrology and Earth Systems Science 7(4): 484-403.
Posch, M., and Reinds, G.J., 2009. A very simple dynamic soil acidification model for scenario
analyses and target load calculations. Environmental Modelling & Software 24:329-340.
Rabalais, N.N., 2002. Nitrogen in aquatic ecosystems. Ambio 31: 102-112.
Raymond, B.A., Bassingthwaighte, T., and Shaw, D.P., 2010. Measuring nitrogen and sulphur
deposition in the Georgia Basin, British Columbia, using lichens and moss. Journal of
Limnology 69(Suppl. 1): 22-32.
Reinds, G.J., Posch, M., and de Vries, W., 2009. Modelling the long-term soil response to
atmospheric deposition at intensively monitored forest plots in Europe. Environmental
Pollution. 157: 1258-1269.
Rixen, C., Freppaz, M., Stoeckli, V., Huovinen, C., Huovinen, K., and Wipf, S., 2008. Altered
Snow Density and Chemistry Change Soil Nitrogen Mineralization and Plant Growth.
Arctic, Antarctic, and Alpine Research. 40(3): 568-575.
Roelofs, J.G.M., Kempers, A.J., Houdijk, L.F.M., and Jansen, J., 1985. The effect of air-borne
ammonium sulphate on Pinus nigra var. maritima in the Netherlandds. Plant and Soil 84:
45-56.
Roobroeck, D., Brüggeman, N., Butterbach-Bahl, K., and Boeckx, P., 2009. Dynamic of nitrate
limitation on gaseous nitogren exchanges from pristine peatlands. Geophysical Research
90
Abstracts 11: EGU2009-12298-3.
Rosen, K., Gundersen, P., Tegnhammar, L., Johansson, M., and Frogner, T., 1992. Nitrogen
enrichment of nordic forest ecosystems - The concept of critical loads. Ambio 21: 364368.
Rouse, J.D., Bishop, C.A., and Struger, J., 1999. Nitrogen Pollution: An assessment of it threat to
amphibian survival. Environmental Health Perspectives 107: 7990803.
Roy, S., Hegde, M.S., and Madras, G., 2009. Catalysis for NO x abatement. Applied Energy 86:
2283-2297.
Russow, R., and Böhme, F., 2005. Determination of the total nitrogen deposition by the 15N
isotope dilution method and problem in extrapolating results to a field scale. Geoderma
127: 62-70.
Saskatchewan Environment and Public Safety, 1988. Toxicity of Ammonia under varying
Temperature and pH Conditions. In: Surface Water Quality Objectives. Water Quality
Branch. From: Environment Canada, Total Ammonia Nitrogen URL:
http://205.189.8.19/water/fa01/fa01d03/fa01s09.en.html [Accessed February, 2010].
Saskatchewan Research Council, 1982. Evaluation of the Status of Surface Water Sensitivity
Mapping for Acidic Deposition in Western Canada. Coordinating Committee on Surface
Waters. SRC No. C-805-11-E-82, 90pp.
Sauvé, J.L., Goddard, T.W., and Cannon, K.R., 2000. A model assessment of greenhouse gas
emissions from agricultural soils in Alberta. 4th International Conference in Integrating
GIS and Environmental Modeling (GIS/EM4): Problems, Prospects and Research Needs.
Banff, Alberta, Sept. 2-8, 2000.
Savard, M.M., Bégin, C., and Smirnoff, A., 2009. Tree-Ring Nitrogen Isotopes Reflect
Anthropogenic NO x Emissions and Climatic Effects. Environmental Science and
Technology 43(3): 604-609.
Schindler, D.W., 1974. Eutrophication and Recovery in Experimental Lakes: Implications for
Lake Management. Science 184(4139):897-900.
Schindler, D.W., Dillon, P.J., and Schreier, H., 2006. A review of anthropogenic sources of
nitrogen and their effects on Canadian aquatic ecosystems. Biogeochemistry 79: 25-44.
Schindler, D.W., Hecky, R.E., Findlay, D.L., Stainton, M.P., Parker, B.R., Paterson, M.J., Beaty,
K.G., Lyng, M., and Kasian, S.E.M., 2009. Eutrophication of lakes cannot be controlled
by reducing nitrogen input: Results of a 37-year whole ecosystem experiment. PNAS
105(32): 11254-11258.
91
Seip, H.M. and Menz, F. 2002. Effects of acid deposition on natural ecosystems – a short
summary. Presentation in Air Pollution as a Climate Forcing: A Workshop. April 29 May 3, Honolulu, Hawaii.
Smith, V. H., Tilman, G. D., and Nekola, J. C., 1999. Eutrophication: impacts of excess nutrient
inputs on freshwater, marine, and terrestrial ecosystems. Environmental Pollution 100:
179-196.
Smyth, S.C., Jiang, W., Roth, H., Moran, M.D., Makar, P.A., Yang, F., Bouchet, V.S., and
Landry, H., 2009. A comparative performance evaluation of the AURAMS and CMAQ
air-quality modelling systems. Atmospheric Environment 43(5): 1059-1070.
St-Amand, A.D., Mayer, P.M., and Blais, J.M., 2009. Modeling PAH uptake by vegetation from
the air using filed measurements. Atmospheric Environment 43: 4283-4288.
Stevens C.J., Dise N.B., Mountford, J.O., and Gowing D.J., 2004. Impact of nitrogen deposition
on the species richness of grasslands. Science 303(5665): 1876-1879.
Stevens, C.J., Maskell, L.C., Smart, S.M., Capron, S.J.M., Dise, N.B., and Gowing, D.J.G., 2009.
Identifying indicators of atmospheric nitrogen deposition impacts in acid grasslands.
Biological Conservation 142: 2069-2075.
Stoddard, J.L. 1994. Long-term changes in watershed retention of nitrogen – its causes and
aquatic consequences. Environmental Chemistry of Lakes and Reservoirs 237: 223-284.
Summers P.W., Bowersox, V.C., and Stensland, G.J., 1986. The geographical distribution and
temporal variations of acidic deposition in eastern North America. Water, Air, and Soil
Pollution 31: 523-535.
Sustainable Forest Management Indicators. Knowledge Base, 2008. Sustainable Forest
Management Research Group, University of British Columbia, Vancouver, British
Columbia. URL: http://www.sfmindicators.org [Accessed February 2010].
Swain, L.G., 1987. Second Report on Chemical Sensitivity of B.C. Lakes to Acidic Inputs.
British Columbia Ministry of Environment and Parks, Victoria, B.C., 31pp.
Tang, Y.S., Simmons, I., van Dijk, N., Di Marco, C., Nemitz, E., Dämmgen, U., Gilke, K.,
Djuricic, V., Vidic, S., Gliha, Z., Borovecki, D., Mitosinkova, M., Hanssen, J.E.,
Uggerud. T.H., Sanz, M.J., Sanz, P., Chorda, J.V., Flechard, C.R., Fauvel, Y., Ferm, M.,
Perrino, C., and Sutton, M.A., 2009. European scale application of atmospheric reactive
nitrogen measurements in a low-cast approach to infer dry deposition fluxes. Agriculture,
Ecosystems and Environment 133: 183-195.
Thimonier, A., Pannatier, E.G., Schmitt, M., Waldner, P., Walthert, L., Schleppi, P., Dobbertin,
M.m and Kräuchi, N., 2009. Does exceeding the critical loads for nitrogen alter nitrate
leaching, the nutrient status of trees and their crown condition at Swiss Long-term Forest
92
Ecosystem research (LWF) sites? European Journal of Forest Research. On-line First
DOI: 10.1007/s10342-009-0328-9.
Thormann, M.N., 2006. Lichens as indicators of forest health in Canada. The Forestry Chronicle
82(3): 335-343.
Tominga, K.S., Watmough, S., and Ahrene, J., 2007. Review of chemical critical limits in critical
load conditions, Unpublished report, Trent University.
Tonnesen, G., Wang, Z., Omary, M., and Chien, C.-J., 2007. Assessment of Nitrogen Deposition:
Modeling and Habitat Assessment. California Energy Commission, PIER Energy-Related
Environmental Research. CEC-500-2005-032, 105pp.
Transport Canada, 2009. Canada / US Emission Control Area (ECA) for ships. Consultation
Presentation, Vancouver, Ottawa, Halifax Environment and Transport Canada, January
2009.
Treseder, K., 2004. A Meta-Analysis of Mycorrhizal Responses to Nitrogen, Phosphorus, and
Atmospheric CO 2 in Field Studies. New Phytologist 164(2): 347-355.
Turetsky, M.R., Halsey L.A., Vitt D.H., and Wieder, R.K., 2003. Enhanced N deposition and
peatland carbon stocks in boreal Alberta. Abstract #B31C-0309 at Fall Meeting of
American Geophysical Union.
Turkington, R., John, E., Krebs, C.J., Dale, M.R., Nams, V.O., Boonstra, R., Boutin, S., Martin,
K., Sinclair, A.R., and Smith, J.N., 1998. The effects of NPK fertilization for nine years
on boreal forest vegetation in northwestern Canada. Journal of Vegetation Science 9: 333346.
Turner, J., 1975. Nutrient cycling in a Douglas-fir ecosystem with respect to age and nutrient
status. Ph.D. Thesis, Univ. of Washington, Seattle, WA.
UNECE-CLRTAP, 2004. United Nations Economic Commission for Europe - Convention on
Long-range Transboundary Air Pollution. Manual on methodologies and criteria for
modelling and mapping critical loads and levels, and air pollution effects and trends,
Berlin, Germany, 202pp. URL:
http://icpmapping.org/cms/zeigeBereich/5/manual_und_downloads.html [Accessed
January 2010].
UNECE-CLRTAP, 2009. United Nations Economic Commission for Europe - Convention on
Long-range Transboundary Air Pollution, Working Group on Effects. Effects of Airborne
Nitrogen. Genvea, Switzerland. ECE/EB.AIR/WG.1/2009/15, 12pp.
UNEP, 2007. Reactive Nitrogen in the Environment – Too Much or Too Little of a Good Thing.
United Nations Environmental Program, Sustainable Consumption and Production
Branch, Paris, Prance, 56pp.
93
US-EPA, 2010. Dissolved Oxygen Depletion in Lake Erie, Great Lakes Monitoring Program.
URL: http://www.epa.gov/glindicators/water/oxygenb.html [Accessed January 2010]
Vaicys, M., and Armolaitis, K., 1986. Gas resistance and regernation of forest damaged by
industrial emissions. In: Plants in Forest Protection. Kiev, Ukraine, 360pp.
van Egmond, K., Bresser, T., and Bouwman, L., 2002. The European Nitrogen Case. Ambio
31(2): 72-78.
Vaughan, H., 2002. Canada’s Ecological Monitoring and Assessment Network. Environment
Canada, Burlington, ON. URL: http://www.eman-rese.ca/eman/program/background.html
[Accessed January 2010].
Vet, R., Brook, J., Ro, C., Shaw, M.., Narayan, J., Zhang, L., Moran, M., and Lusis, M., 2005.
Chapter 3: Atmospheric response to past emission control programs. In: The 2004
Canadian Acid Deposition Science Assessment. Environment Canada, Ottawa.
Vitousek, P.M., and Howarth, R.W., 1991. Nitrogen limitation on land and in the sea: How can it
occur? Biogeochem 13: 87-115.
Vitousek, P.M., Aber, J., Howarth, R.W., Likens, G.E., Matson, P.A., Schindler, D.W.,
Schlesinger, W.H., and Tilman, G.D., 2000. Huamn alteration of the global nitrogen
cycle: causes and consequences. In: “Issues in Ecology”. Ecological Society of America.
Walk, M.F., 2007. Acid deposition research needs in Massachusetts. Expert Summit Report
funded by the MA Department of Environmental protection, Bureau of Waste Prevention.
20 pp. URL: www.umass.edu/tei/wrrc/ [Accessed January 2010].
Wang, X., Zhang, Y., Chen, H., Yang, X., and Chen, J., 2009. Particulate Nitrate Formation in a
Highly Polluted Urban Area: A Case Study by Single-Particle Mass Spectrometry in
Shanghai. Environmental Science and Technology 43: 3061-3066.
Watmough, S.A., and Dillon, P.J., 2001. Base cation losses from a coniferous catchment in
central Ontario. Water, Air and Soil Pollution: Focus 1 507-524.
Watmough, S.A., and Dillon, P.J., 2003. Base cation and nitrogen budgets for seven catchments
in central Ontario, 1983–1999. For. Ecol. Manage. 177: 155–177.
Watmough, S.A., Ahern,e J., Alewell, C., Arp, P., Bailey, S., Clair, T., Dillon, P., Duchesne, L.,
Eimers, C., Fernandez, I., Foster, N., Larssen, T., Miller, E., Mitchell, M., and Page, S.,
2005. Sulphate, nitrogen and base cation budgets at 21 forested catchments in Canada, the
United States and Europe. Environmental Monitoring and Assessment 109: 1-36.
Wedin D.A., and Tilman D., 1996. Influence of nitrogen loading and species composition on the
carbon balance of grasslands. Science 274(5293): 1720-1723.
Wesely, M.L., 2007. Parameterization of surface resistances to gaseous dry deposition in
94
regional-scale numerical models. Atmospheric Environment 41:S52-S63.
Westerveld, S.M., McKeown, A.W., McDonald, M.R., and Scott-Dupre,e C.D., 2004.
Chlorophyll and Nitrate Meters as Nitrogen Monitoring Tools for Selected Vegetables in
Southern Ontario. ISHS Acta Horticulture 627: XXVI International Horticultural
Congress: Toward Ecologically Sound Fertilization for Field Vegetable Production.
Toronto, ON.
Whitfeild, C.J., Watmough, S.A., Aherne, J., and Dillon, P.J., 2006. A comparison of weathering
rates for acid-sensitive catchments in Nova Scotia, Canada and their impact on critical
load calculations. Geoderma 136: 899-911.
Whitfield, C.J., Aherne, J., Cosby, B.J., and Watmough, S.A., 2010. Modelling catchment
response to acid deposition: a regional dual application of the MAGIC model to soils and
lakes in the Athabasca Oil Sands Region, Alberta. Journal of Limnology 69(Suppl. 1):
147-160.
Wigington, Jr P.J., Davies, T.D., Tranter, M., and Eshleman, K.N., 1992. Comparison of episodic
acidification in Canada Europe and the United States. Environmental Pollution 78: 29-35.
Wiklander, G., Nordlander, G., and Andersson, R., 1991. Leaching of nitrogen from a forest
catchment at Soderasen in Southern Sweden. Water Air and Soil Pollution 55: 262-282.
Wrona, F.J., Gummer, W.D., Cash, K.J., and Crutchfield, K., 1996. Cumulative impacts within
the northern river basins. Synthesis Report No. 11. Edmonton: Environment Canada and
Alberta Environmental Protection, Northern River Basins Study.
Xue, X., Hong, H. and Charles, A.T., 2004. Cumulative environmental impacts and integrated
coastal management: the case of Xiamen, China. Journal of Environmental Management.
71: 271-283.
Zhang, L., Brook, J. R., Vet, R., Wiebe, A., Mihele, C., Shaw, M., O’Brien, J. M., and Iqbal, S.,
2005. Estimation of contributions of NO 2 and PAN to total atmospheric deposition of
oxidized nitrogen across Eastern Canada. Atmospheric Environment 39: 7030-7043.
Zhong, A.L., 2004. Assessment of ecosystem sensitivity to N deposition for the Mountain
Hemlock Zone of the Georgia Basin. Contract report prepared for Environment Canada.
34pp.
95