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In the past, the phrase ‘environmental allocations of water’ has most often been taken to mean allocation of water to rivers. However, it is now accepted that groundwater-dependent ecosystems are an important feature of Australian landscapes and require an allocation of water to maintain their persistence in the landscape. However, moving from this theoretical realisation to the provision and implementation of a field-based management regime is extremely difficult. The following four fundamental questions are identified as being central to the effective management of groundwater-dependent ecosystems (GDEs): (1) How do we identify GDEs in the field; put another way, which species or species assemblages or habitats are reliant on a supply of groundwater for their persistence in the landscape; (2) what groundwater regime is required to ensure the persistence of a GDE; (3) how can managers of natural resources (principally water and habitats), with limited time, money and other resources, successfully manage GDEs; and (4) what measures of ecosystem function can be monitored to ensure that management is effective? This paper explicitly addresses these questions and provides a step-by-step theoretical and practical framework for providing answers. In particular, this paper provides an introduction to some of the relevant literature and from this, presents a synthesis, presented in the form of a functional methodology for managing groundwater dependent ecosystems.
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CSIRO PUBLISHING Australian Journal of Botany,2006, 54, 97–114
A functional methodology for determining the groundwater regime needed
to maintain the health of groundwater-dependent vegetation
Derek EamusA,C, Ray FroendB, Robyn LoomesB, Grant HoseAand Brad MurrayA
AInstitute for Water and Environmental Resource Management and Department of Environmental Sciences,
University of Technology, Sydney, PO Box 123, NSW 2007, Australia.
BCentre for Ecosystem Management, Edith Cowan University, Joondalup, 100 Joondalup Drive, WA 6027, Australia.
CCorresponding author. Email:
Abstract. In the past, the phrase ‘environmental allocations of water’ has most often been taken to mean allocation
of water to rivers. However, it is now accepted that groundwater-dependent ecosystems are an important feature of
Australian landscapes and require an allocation of water to maintain their persistence in the landscape. However,
moving from this theoretical realisation to the provision and implementation of a field-based management regime
is extremely difficult. The following four fundamental questions are identified as being central to the effective
management of groundwater-dependent ecosystems (GDEs): (1) How do we identify GDEs in the field; put another
way, which species or species assemblages or habitats are reliant on a supply of groundwater for their persistence in
the landscape; (2) what groundwater regime is required to ensure the persistence of a GDE; (3) how can managers
of natural resources (principally water and habitats), with limited time, money and other resources, successfully
manage GDEs; and (4) what measures of ecosystem function can be monitored to ensure that management is
effective? This paper explicitly addresses these questions and provides a step-by-step theoretical and practical
framework for providing answers. In particular, this paper provides an introduction to some of the relevant literature
and from this, presents a synthesis, presented in the form of a functional methodology for managing groundwater
dependent ecosystems.
If the 19th century was dominated by the acquisition
and defence of land (territory) and the 20th century was
dominated by the acquisition and control of oil and energy
resources, then the 21st century will be dominated by the
politics of water. Globally, secure access to potable water has
been identified as the key political, humanitarian and military
flash point (Anon. 2000; Postel 2003). However, in addition
to the human need for water, it has become increasingly clear
that the maintenance of a supply of water for the environment
is equally important to human welfare. Environmental water
requirements support the obvious economic activities of
irrigated crops and pastures but also maintain the often-
overlooked direct and indirect values of a large number
of ecosystem services (Eamus et al. 2005). Ecosystem
services, or ecosystem goods and services (hereafter the two
are deemed to be contained within the phrase ecosystem
services) are those processes and attributes of an ecosystem
(or part of an ecosystem) that benefit humans (Costanza
et al. 1997). Examples of ecosystem services include
soil formation, prevention of soil erosion, regulation of
water flow (surface, subsurface and groundwater recharge),
water purification, tourism value and carbon sequestration
(Costanza et al. 1997).
When groundwater-dependent ecosystems (GDEs) are
threatened by insufficient supply of groundwater, all the
ecosystem services provided by that ecosystem may be
threatened. In this review, groundwater is defined as the
saturated zone of the regolith and its associated capillary
fringe. It is distinct from soil water which is water held
under tension in soil pore spaces and is generally not
saturated. We do not concern ourselves with differentiating
between regional, local or perched watertables as this paper
is principally concerned about the relationship between
groundwater regime and vegetation function, irrespective of
the nature of the groundwater resource per se.
As demand for water by humans increases, the utilisation
of groundwater reserves increases, especially in arid and
semi-arid zones. However, it has become apparent that
even in humid and wet–dry tropical climates, including
northern Australia, significant ecosystems are dependent
on groundwater supply for their continued existence and
the unavailability of groundwater will have significant
deleterious effects on all GDEs (Murray et al. 2003).
© CSIRO 2006 10.1071/BT05031 0067-1924/06/020097
98 Australian Journal of Botany D. Eamus et al.
Consequently, there is a clear need to be able to manage
groundwater resources so that environmental requirements
remain protected. However, there are four major impediments
to the sustainable management of GDEs. These can be stated
as the following four questions:
(1) how do we identify GDEs in the field; put another way,
which species or species assemblages or habitats are
reliant on a supply of groundwater for their persistence
in the landscape;
(2) what groundwater regime is required to ensure the
persistence of a GDE;
(3) what are the safe limits to change in groundwater regime.
Put another way, what do managers of natural resources
(principally water and habitats), with limited time, money
and other resources, actually manage a GDE; and
(4) what measures of ecosystem function can be monitored
to ensure that management is effective?
The purpose of this paper is to provide a conceptual and
methodological framework (see Table 1) to explicitly address
these questions and to explore underlying assumptions and
the information required to tackle them. The related question,
i.e. how to prioritise allocations of limited resources to
competing GDEs, is dealt with by Murray et al. (2006).
Where possible, published examples are provided to illustrate
the application of field methods to the determination
of groundwater dependency of terrestrial vegetation and
the determination of the groundwater regime required to
maintain the health of these systems.
The remainder of this paper has been divided into five
sections, each dealing with a question listed above. The
first section addresses the question of how to identify
GDEs. In addressing this question we first present a new
simple classification system of GDEs. We then discuss
how groundwater dependency for each class may be
inferred, or measured, in the field. We also ask what
degree of dependency is expressed and what vegetation
processes are groundwater-dependent within GDEs? The
second section addresses the question of what attributes
of a groundwater regime are important to a GDE? Five
attributes of groundwater are deemed important ecologically
and these are discussed. The third section deals with the
most vexed question facing managers, i.e. what is the
likely response of vegetation to a degraded groundwater
regime? Both practical examples of how this question has
been addressed in the past and new, potential approaches
for future application, are discussed. The vexed issue of
defining the response function of a species or ecosystem to
changes in groundwater regime, are discussed. The fourth
section discusses the question at individual, community
and population scales. Criteria for determining which
measures of individuals, communities and populations are
optimal for quantification of the response to changes
in groundwater regime, are discussed. Finally, the last
section provides concluding comments about putting theory
into practice.
How to identify GDEs
Groundwater-dependent ecosystems can be located in
marine, coastal, riparian, in-stream, terrestrial and in cave
and aquifer environments. To assist in future discussion and
management of GDEs, we propose three simple primary
classes of GDEs. These are the following:
(I) Aquifer and cave ecosystems, where stygofauna
(groundwater-inhabiting organisms) reside within the
groundwater resource (for discussion see Humphreys
2006). These ecosystems include karstic, fractured rock
and alluvial aquifers. We also consider the hyporheic
zones of rivers and floodplains in this category because
these ecotones often support stygobites (obligate
groundwater inhabitants). For further discussion of the
hyporheic zone see Boulton and Hancock (2006).
(II) All ecosystems dependent on the surface expression
of groundwater. This therefore includes base-flow
rivers and streams, wetlands, some floodplains and
mound springs and estuarine seagrass beds. While it
is acknowledged that plant roots are generally below
ground, this class of GDEs requires a surface expression
of groundwater, which may, in many cases, then soak
below the soil surface and thereby become available to
plant roots.
(III) All ecosystems dependent on the subsurface presence
of groundwater, often accessed via the capillary fringe
(non-saturated zone above the saturated zone of the
water table) when roots penetrate this zone. This
class includes terrestrial ecosystems such as river
red gum (E. camuldulensis) forests on the Murray–
Darling basin, and Banksia woodland on the Gnangara
mound of Western Australia. No surface expression of
groundwater is required in this class of GDE.
Implicit within Classes II and III are the associated
faunal, microbial and fungal populations. However, within
this review, we focus on consideration of the vegetation
components of the ecosystem, for which there tends to be
more information.
Application of this simple classification system will assist
managers in identifying the appropriate techniques (selecting
the correct tool) for identifying the presence, timing and
nature of groundwater dependency. This classification differs
markedly from earlier attempts (e.g. Sinclair Knight Merz
(SKM) 2001). Our approach to this classification scheme
has been to identify ecosystems associated with a common
groundwater resource. For instance, Class II consists of those
ecosystems that utilise groundwater once it is expressed
above the land surface, whereas Class III consists of
those ecosystems that utilise groundwater before its surface
expression. Class I ecosystems are simply subterranean
Management of groundwater-dependent ecosystems: functional methodology Australian Journal of Botany 99
Table 1. A summary of the key questions to be addressed and methods that may be applied in the process of identifying and managing
groundwater-dependent ecosystems
Not all of these are discussed in this paper. GW, groundwater
Key question Methods that may be applied to answer the question
(1) Which populations or
species of an ecosystem
are GW-dependent?
(a) Observations to answer the nine questions listed above in section Ecosystems reliant on surface
expressions of groundwater. (b) Stable isotope comparison of GW, soil and vegetation water content.
(c) In the case of base flow systems, the concentration of a conserved tracer, such as CFC, magnesium,
oxygen-18 and radon in a river and GW can be measured.
(2) If some populations or
species of an ecosystem
are GW-dependent, what
degree of dependency is
(a) Establish the degree of dependency by assessing the proportion of water required that is GW-derived
(based on annual water budgets and isotope/radon analyses) and assume that the degree of dependence
is proportional to the fraction of the annual water budget that is GW-derived. (b) Use the approach
of Scott et al. (1999) and Shafroth et al. (2002) who quantified the relationship between patterns of
vegetation response and changes in GW depletion and matched this with quantitative data on rate, depth
and duration of declines in the water table and apply this approach to all three dependence classes.
(c) Examine seasonal patterns in climate, leaf area index, stable isotope data and vegetation function to
determine the timing and nature of likely GW dependency; deduce rules about nature of dependency
from these relationships.
(3) What patterns in
GW-dependency are
(a) Establish threshold values for GW attributes that support the processes and populations/species
identified as being GW-dependent, using best-available knowledge (expert reviews) of relationships
between processes that are dependent and GW availability. (b) Establish consumptive rates of GW use,
where appropriate; (c) establish temporal and spatial distribution of GW use and availability.
(4) What processes are
(a) For wetlands, for example, examine relationships among timing of seasonal flooding, climate and
key ecosystem processes, especially growth rate, flowering, seed germination and dispersal. Long-term
records for frequency, duration and depth of flooding and population demography, can be examined to
establish a link between the important components of the water regime and recruitment and population
structure. (b) For wetlands, for example, examine the relationships among tolerance to flooding of key
species, the historical record of floods and zonation of species distribution; (c) similar relationships can
be examined also for terrestrial and baseflow systems.
(5) What attributes of GW
(level, flux, hydraulic head,
quality) are important
to the dependent
(a) Establish benchmarking (or reference condition) for comparable systems in comparable climate
envelopes where GW abstraction or availability has not changed and compare vegetation/ecosystem
responses. (b) Interrogate data from long-term monitoring sites (where available), and where not
available, establish expert assessment of the likely impacts of changes in flux, level, quality or pressure
on GW dependent systems; (c) apply the approach of Loomes (2000), who analysed the distribution of
60 wetland species in relation to surface-water data to determine mean minimum and maximum water
depths to establish EWRs, to the other classes of GW-dependent systems (phreatophytic and baseflow),
by comparing sites with known differences in GW attributes (e.g. level, flux).
(6) What are the safe limits to
changes in the attributes of
GW that are important?
(a) Compare and contrast the community composition, structure and functioning of GW-dependent
communities having different GW availabilities, within each class of GW-dependent ecosystem
(i.e. within the following three classes: wetlands, baseflow systems and phreatophytic systems).
(b) Apply the methodology and results of Loomes (2000), who analysed the distribution of
60 wetland species in relation to surface-water data to determine mean minimum and maximum
water depths to establish EWRs, to the other classes of GW-dependent systems (phreatophytic and
baseflow), by comparing sites with known differences in GW attributes (e.g. level, flux). (c) Examine
which ecosystem process (e.g. consumptive water use by trees; or germination and recruitment of
seeds/seedlings) is most dependent on GW and thereby infer safe limits of GW change. For example,
if different vegetation habits (e.g. tree water use v. shrub water use) have different dependencies (as
found by Froend and co-workers in Western Australia) then safe limits for changes in GW depth
can be determined for a site by using this information. (d) Establish rooting depths of GW-dependent
systems and seasonal patterns of depth to watertable. (e) Compare the response of hyporheic invertebrate
communities in baseflow systems to changes in flow permanence and timing. Consider the significance
of the loss of habitat distinction (e.g. surface riffles and pools) if baseflows are limited to the hyporheic
zone. (f) Assess the likelihood of baseflow dependence by analysis of low-flow stream signatures,
hydrographic baseflow separation techniques, isotopic analysis and partitioning of water sources, and
calibration of local stream–aquifer interaction models. (g) Use Hotspots software to calculate how
much GW can be extracted before GW pressure/level is adversely affected and to model
stream–aquifer interactions.
100 Australian Journal of Botany D. Eamus et al.
Tab l e 1 . (continued)
Key question Methods that may be applied to answer the question
(7) What is the response
function of key species or
(a) Compare and contrast the community composition, structure and functioning of GW-dependent
communities having different groundwater availabilities, within each class of GW-dependent ecosystem
the community to changes
in GW regime (supply/flux/
pressure/quality or level)?
(i.e. within the following three classes: wetlands, baseflow systems and phreatophytic systems).
(b) Establish the ranking of change in cover, recruitment, growth rate or other attribute of individual
species’ health to changes in water regime, within each of the three classes of dependent systems.
(c) Examine relationships between net primary productivity and water availability of a large number of
diverse sites to establish the response function of NPP to changes in water availability.
(8) What values are assigned
by all stakeholders, to the
GD vegetation of the
(a) Community consultation on all values assigned to the system showing GW dependency. (b) Expert
assessment of biodiversity/rare and endangered species, uniqueness and other values of the ecosystem.
(c) Economic assessment of ecosystem services, including amenity, tourism, conservation, economic
productivity (e.g. pastoral use).
(9) What are the acceptable limits of
change of GW (flux/pressure/
level/quality) that does not
cause unacceptable change in
ecosystem composition/structure/
function or services?
(a) Community consultation on defining the acceptable limits of change in system values, identified
from (8) above. (b) Expert assessment of acceptable limits to system attributes of value. (c) Economic
assessment of responses of economy to changes in provisionof all ecosystem ser vices, including amenity,
tourism, conservation, economic productivity (e.g. pastoral use).
(10) What is the EWR to maintain
the values of the GD vegetation
of the ecosystem?
(a) Apply the methodology and results of Loomes (2000), who analysed the distribution of 60 wetland
species in relation to surface-water data to determine mean minimum and maximum water depths to
establish EWRs, to the other classes of GW-dependent systems (phreatophytic and baseflow) and to
a broader range of wetlands. (b) Use sapflow sensors and open-top chambers to quantify seasonal
patterns of vegetation water use and micro-climate approaches to estimate vegetation water use. (c) Use
expert opinion to define minimum water-level requirements for all three classes of GW-dependent
systems, using existing and new information developed in this project, integrated with data in reports
from (e.g. the Daly River study, which identified EWR for the Daly River using a broad range of
indicators). (d) Compare, contrast, apply and refine the methodologies of Roberts et al. (2000) and
Arthington and Zalucki (1998) protocols for defining water requirements of floodplain wetlands and
riparian systems to several test sites. (e) Develop a matrix of vulnerability (ranked from knowledge of
groundwater dependence) v. ‘value’ (as defined by all stakeholders) to establish a rank for importance
for GD vegetation at a location. In addition, develop a ranking of system responses to several categories
of change in GW attributes (small, medium or large). Combine this with the matrix to establish upper
and lower limits of GW supply to maintain the health of the GD system. (f) Where GW abstraction by
bores impinges on dependent systems, apply Hotspots software tool to determine permissible abstraction
quantities and schedules so that water-level thresholds are not breached.
aquatic ecosystems. This classification also addresses
inherent difficulties in earlier classifications of GDEs.
For instance, it is difficult to compare meaningfully a
‘terrestrial fauna’ GDE with a near-shore marine GDE,
given that they are at vastly different scales and levels of
biological organisation.
A toolbox for identifying GDEs
A toolbox contains the methods, approaches and techniques
required to tackle a problem. In this section we describe
these tools in relation to the first problem confronting
managers, i.e. how to identify the presence of a GDE
in a landscape.
Aquifer and cave ecosystems
All subterranean waters constitute GDEs, even if the
biotic component of the ecosystem is limited to a microbial
flora. The presence of groundwater can be determined by
routine geophysical methods such as electrical resistivity
and seismic refraction (Reynolds 2000). After this inferential
approach, more detailed analysis to identify the presence
of biological organisms will include microbial plating (see
Gounot 1994) and surveys of macro-invertebrates (Hahn
2002). Because the focus of this paper is terrestrial vegetation,
no further discussion of this class of GDEs is made.
Ecosystems reliant on surface expressions
of groundwater
Groundwater presence (1–5 below) and hence potential
for use, or actual use, may be inferred from positive answers
to one or more of the following questions:
(1) does a stream/river continue to flow all year, or
a floodplain waterhole remain wet all year, despite
prolonged periods of zero surface flows (that is, zero or
very low rainfall);
(2) for estuarine systems, does the salinity drop below that
of seawater in the absence of surface-water inputs (e.g.
tributaries or stormwater);
Management of groundwater-dependent ecosystems: functional methodology Australian Journal of Botany 101
(3) does the volume of flow in a stream/river increase
downstream in the absence of inflow from a tributary;
(4) is the level of water in a wetland/swamp maintained
during extended dry periods;
(5) is groundwater discharged to the surface for significant
periods of time each year or at critical times during
the lifetime of the dominant vegetation type (if such a
resource is present, some species present are likely to be
adapted to be using it);
(6) is the vegetation associated with the surface discharge of
groundwater different (in terms of species composition,
phenological pattern, leaf area index or vegetation
structure) from vegetation close-by but which is not
associated (i.e. accessing) this groundwater;
(7) is the annual rate of water use by the vegetation
significantly larger than annual rainfall at the site and
the site is not a run-on site; e.g. low-lying paperbark
(Melaleuca spp.) swamps in the Northern Territory
receive surface and subsurface (lateral) flows of water);
(8) are plant water relations (especially pre-dawn and mid-
day water potentials and transpiration rates) indicative of
lower water stress (potentials closer to zero; transpiration
rate larger) than for vegetation located nearby but not
accessing the groundwater discharged at the surface; (the
best time to measure this is during rainless periods); and
(9) is occasional (or habitual) groundwater release at the
surface associated with key developmental stages of
the vegetation (such as flowering, germination, seedling
Affirmative answers to one or more of these questions
support the inference that the system is likely to be a GDE,
if we accept the view that when a resource is present, at
least some species will be accessing that resource. However,
this inference does not provide any information about the
nature of the dependency (obligate or facultative). Neither
does it provide information about the groundwater regime
(timing of groundwater availability, volume of availability,
location of surface expression, the hydraulic head of the
groundwater aquifer required to support the surface discharge
of groundwater) needed to support the ecosystem. These
issues are discussed in the second section.
More direct evidence that an ecosystem is using
groundwater can be obtained by comparing the stable isotope
(in particular deuterium and oxygen-18) composition of
groundwater, soil water, surface water (where relevant) and
vegetation xylem water (Thorburn et al. 1993; Zencich
et al. 2002). Where there is sufficient variation in isotopic
composition among these sources then it is possible to
identify the single or the most dominant source of water
being used by different species at different times of year
(Zencich et al. 2002). Thus, the use of stable isotopes can
provide information about spatial and temporal variation in
groundwater dependency within and between species and
ecosystems (O’Grady et al. 2006).
For base-flow systems (that is, rivers and streams
showing significant flows during periods of zero surface
or lateral flows and standing-water habitats such as lakes),
measurements of the chlorofluorocarbon, magnesium or
radon concentrations of river and groundwater supply can
identify and quantify the amount and timing of groundwater
inflows into the river (Cook et al. 2003).
Ecosystems reliant on subsurface presence
of groundwater
The presence of ecosystems reliant on subsurface presence
of groundwater may be inferred from positive answers to one
or more of the following questions:
(1) Is groundwater or the capillary fringe above the water
table present within the rooting depth of any of the
(2) Does a proportion of the vegetation remain green and
physiologically active (principally, transpiring and fixing
carbon, although stem-diameter growth or leaf growth
are also good indicators) during extended dry periods of
the year?
(3) Within a small region (and thus an area having the same
annual rainfall and same temporal pattern of rainfall
across its entirety), and in an area not having access to
run-on or stream or river water, do some ecosystems show
large seasonal changes in leaf area index while others
do not?
(4) Questions 5–7 in the section Ecosystems reliant on
surface expression of groundwater.
(5) Are seasonal changes in groundwater depth larger than
can be accounted for by the sum of lateral flows and
percolation to depth (that is, is vegetation a significant
discharge path for groundwater; Cook et al. 1998)?
Clearly, if the error terms in the estimation of lateral
flow and percolation to depth are of similar or greater
magnitude than the rate of vegetation water use, this
method may not be appropriate.
Stable isotopes can be used for these systems too, as
can artificial labelling with tracers, such as lithium or
deuterium. When tracers are added to the groundwater,
the subsequent uptake into vegetation is usually conclusive
proof that access by that vegetation is occurring. However,
the presence of a tracer in a shallow-rooted species can
occur if neighbouring deep-rooted species exhibit hydraulic
lift and the shallow-rooted plants then ‘harvest’ this water
(Caldwell et al. 1998).
What degree of dependency is expressed?
If a habitat or ecosystem is identified as being a GDE,
it is valuable to know the degree of dependency exhibited.
Different GDEs can exhibit a range of dependencies, from
obligate to facultative (Hatton and Evans 1997). Stygofauna
can be assumed to be obligately dependent, as are the
102 Australian Journal of Botany D. Eamus et al.
flora and fauna of mound springs. Obligate dependency
should not be taken to mean that groundwater is required
continuously. Dependency can be deemed to be obligate
if the groundwater is relied on only very infrequently
(6 months in every 10–20 years), or frequently but for
short periods of time (1 month in every 12 months). The
former can be difficult to show when long-lived trees
are being considered. It is conceivable that groundwater
availability may be required only infrequently, for example,
in the establishment of trees on floodplains (George et al.
2005), to sustain the population through to the end of
a particularly long cycle of below-average rainfall, or it
may have played a crucial role in the establishment of
a population or cohort (McBride and Strahan 1984).
Consequently, the impact of the loss of groundwater may
not be apparent for many years, even decades. Recharge of
floodplains by groundwater may be required infrequently,
to recharge soil stores and to support germination and
early establishment of vegetation. However, the ecological
importance of those floodplains is significant (Kingsford
2000). Thus, the timing and degree of groundwater
dependency is important knowledge for landscape
managers. At the species level, obligate groundwater
use is evident if all instances of a species presence are
dependent on either continuous, seasonal or episodic access
to groundwater.
In contrast, facultative dependency occurs when
groundwater is used when it is available but its absence does
not cause the loss of this vegetation element from that site
(O’Grady et al. 2006). Populations of facultative species may
include individuals that access groundwater when at shallow
depth and individuals that have not accessed groundwater
throughout their life, e.g. at higher positions in the landscape
(Zencich et al. 2002).
To establish the degree of groundwater dependency is
difficult and can take many years. Four approaches serve to
highlight this.
(1) The first approach determines the degree of dependency
by quantifying the proportion of annual water use that
is derived from groundwater and then assumes that this
is a measure of the degree of dependency. When there is
a consistent utilisation of groundwater each year, this
approach may be applied, but it is costly and takes a
minimum of 12 months of field work. It also does not
differentiate between an obligate use of groundwater
(where its absence will have a severe negative impact) and
a facultative use (where its absence will not have a severe
effect). Both uses could give the same numeric value
but the management implications of loss of groundwater
availability differ. Furthermore, when groundwater use
becomes significant only occasionally (for example for
the last 2 years of a 10-year cycle of low rainfall), this
approach is unlikely to accurately reflect the degree of
dependency unless the period of study is longer than the
cycle length.
(2) Proportional contributions of groundwater and soil water
(and streamwater) can also be assessed by examining
stable isotope composition of xylem water and water
sources and applying mass-balance mixing models
(Rains and Mount 2002; Zencich et al. 2002).
(3) An alternative approach used by Scott et al. (1999) and
Shafroth et al. (2000, 2002) quantifies the relationship
between patterns of change in groundwater availability
(for example, depth, rate of decline in depth, and duration
of excessive depths of the water table) and vegetation
responses, including changes in crown volume, radial
stem growth and mortality. Froend and co-workers in
Western Australia have applied this approach to Banksia
woodland and wetland vegetation of the Gnangara and
Jandakot mounds (Froend et al. 2004), to ascertain
the response of phreatophytic vegetation to separation
from the groundwater source. This can be achieved
through either long-term monitoring of vegetation
vigour and composition relative to groundwater regime,
or via shorter-term drawdown experiments where
the watertable is manipulated to induce seasonal or
inter-annual change in plant water-source partitioning.
The former technique provides a more accurate
determination of the degree of dependency at the
plant-community or population level; however, it
requires long periods of study. The latter approach is
particularly useful for identifying individual dependency
on groundwater.
(4) A third, inferential approach to estimating the degree
of groundwater dependency is to examine temporal
patterns in soil moisture availability, rainfall and
vegetation variables known to be influenced by soil
moisture content (for example, leaf area index and
vegetation water use). From these patterns, rules about
the likely temporal dependency (rather than the degree
of dependency assessed through quantitative analysis of
groundwater use) are deduced. This approach underpins
the work of Cook et al. (1998).
What vegetation processes are groundwater-dependent?
The maintenance of ecosystem structure and function
(a better aim than that of maintaining ecosystem health
simply because they are easier to define and hence measure)
requires the maintenance of some key ecosystem processes.
These include the following:
(1) flowering, seed set and germination;
(2) growth and persistence;
(3) seedling establishment and recruitment to reproductive
(4) mortality; and
(5) nutrient cycling.
Management of groundwater-dependent ecosystems: functional methodology Australian Journal of Botany 103
These are the foundations of the provision of ecosystem
services as discussed by de Groot et al. (2002) and
Murray et al. (2006).
In order to establish which of these, and other processes,
are most sensitive to changes in groundwater availability,
one of several approaches can be applied. The first
approach requires an analysis of historical records of
vegetation behaviour in relation to observed changes in
groundwater availability (for example, flooding or depth to
groundwater). There are few sites in Australia where these
data are available, but where they are available, important
insights into groundwater-dependent processes have emerged
(see below).
In the USA, the response of riparian and floodplain
Populus spp (cottonwoods). to changes in hydrologic regime
has received extensive study (Scott et al. 1999; Amlin and
Rood 2002; Rood et al. 2003). Riparian cottonwoods are
dependent on shallow alluvial groundwater that is linked to
stream water, especially in semi-arid regions. When this water
becomes unavailable riparian cottonwoods show symptoms
of stress, including stomatal closure, reduced transpiration
and photosynthesis, reduced predawn and midday water
potentials and increased xylem cavitation (Rood et al. 2003).
Morphological responses observed include reduced shoot
and root growth, crown die-back and eventually increased
mortality (Scott et al. 1999).
Shafroth et al. (1998) used historic discharge data and
information on the successful establishment of three native
and one exotic species to model locations along the Bill
Williams River in Arizona where these species would become
established. Total basal area of mature woody vegetation, the
maximum annual depth to ground water, and the maximum
rate of decline of the water table were the variables that best
differentiated between quadrats with or without seedlings,
thereby establishing the importance of these in determining
seedling recruitment.
By using an experimental approach, Amlin and Rood
(2002) showed that the low rates of decrease in water
table (<2 cm per day) promoted root and shoot growth
in saplings and seedlings of Populus spp. and Salix spp.
compared with the zero-decline treatment. Increases in water-
table depth larger than 2 cm per day tended to reduce
growth. They recommended a gradual stream recession along
regulated rivers of 2.5cm per day in late spring in order
to encourage recruitment of seedlings of cottonwood, but a
more gradual recession of 1 cm per day through the lower
zones would support recruitment of other willows. Clearly,
in the short term, changes in the rates of water use and
carbon gain occur as plants become drought stressed and
stomatal closure occurs. A stimulation of root growth is
commonly observed in response to drought. When drought is
prolonged, however, reduced shoot growth occurs, along with
a reduction in leaf area index (crown die-back) and mortality
(Amlin and Rood 2002).
Pettit et al. (2001) compared the relationships between
flow regime and riparian vegetation characteristics for
two rivers in Western Australia. In the Blackwood River
(south-western Western Australia), mean monthly discharge
is highly seasonal but predictable, with low variability
each month. The Ord River is also very seasonal but
variability is very high, reflecting the impact of monsoonal
weather patterns and groundwater input to the river. These
differences in stream hydrology, water depth, duration
of flooding and flood frequency were strongly correlated
with plant population dynamics, floristic composition and
vegetation structure.
On the Blackwood River, species richness and cover
of shrubs reduced with increased frequency and duration
of flooding, whereas cover of exotic species and annual
herbs increased with increased flooding. Importantly, the
germination of tree seedlings was not influenced by flood
regime but tree size increased with flood duration. On the Ord
River, species richness was not influenced by flooding regime
but the cover of perennial grasses increased with flooding
frequency and shrub cover decreased, whereas tree size
decreased as the number of flood events per year increased
(Pettit et al. 2001).
Froend and McComb (1994) applied an alternative
approach to address the following question: which vegetation
processes are sensitive to alterations in hydrologic regime?
They examined, along a water regime gradient, the
distribution, productivity and reproduction of two emergent
macrophytes in eight wetland lakes on the Swan Coastal
Plain in Western Australia. They showed that standing
biomass and ramet and inflorescence densities varied
along the gradient, with maximum values occurring most
often at intermediate water depths. There was a shift in
phenology (ramet emergence, new leaf growth, flowering
and seed production) with increasing mean water depth
and nutrient status. In addition, seasonal values of above-
ground productivity changed along the water-regime gradient
for both species. Clearly, such a comparative approach can
be applied in the absence of historical data to predict the
response of plant communities and populations to altered
groundwater availability.
For floodplains and wetlands, long-term records of
frequency of flooding and depth of flooding have been linked
to population demography and the response of patterns of
zonation of vegetation (Hughes and Rood 2003). Floodplain
forests are dependent on flooding and rely on well timed,
periodic floods for the provision of regeneration sites.
They require a tapered flood recession for the successful
establishment of seedlings. Such overbank flood events
are central to forest regeneration and need occur only
infrequently. In contrast, following the establishment of
immature forest trees, growth of these trees also requires
adequate and variable ‘maintenance flows’ throughout
the year. Regeneration flows are often synonymous with
104 Australian Journal of Botany D. Eamus et al.
flood flows and occur only infrequently. Therefore, the
hydrologic regime required for recruitment differs from
that required to maintain growth from establishment to
maturity. Kingsford (2000) similarly examined long-term
consequences to vegetation of changes in river flow
following regulation (damming) of rivers. He noted that
more than half of all Australian floodplain wetlands on
regulated rivers may no longer flood. Consequently, some
floodplain wetlands are now permanent storages and this
has changed their biota from one that tolerates a variable
flooding regime, to one that tolerates permanent flooding
(Kingsford 2000).
What attributes of a groundwater regime
are important to a GDE?
Having established the presence of a groundwater
dependency at a site, it is important then to establish
which attributes of the groundwater regime are important.
Generally, the following seven attributes are important
(1) the depth of the saturated zone of water (usually the
capillary fringe above the water table), with particular
reference to distance between the capillary fringe and
plant roots;
(2) groundwater flow rate to the site;
(3) the hydraulic head within an aquifer, which determines
the flow of groundwater, to a point of discharge (such as
an artesian spring);
(4) the quality (especially the salt or nutrient or pollutant
concentration) of the groundwater;
(5) location of discharge; changes in the location of discharge
will obviously influence the distribution of the vegetation
associated with that discharge;
(6) the normal pattern of change in groundwater depth,
including the timing, frequency and duration of maxima
and minima (where these exist) of groundwater depth;
(7) the rate of change of groundwater depth.
To determine the fourth point is relatively simple; analyses
of groundwater quality can be routinely undertaken by
analytical laboratories. Generally, poor-quality groundwater
(high in salt or heavy metals, for example) is less supportive of
vegetation. Reduced growth rates, crown die-back, increased
mortality and changes in species composition (from less
saline-tolerant to more saline-tolerant species, for example)
are impacts of poor-quality groundwater.
The location of the discharge is also relatively easy to
determine and predicting vegetation response to a change
of location of discharge can be relatively simple (although
the time frame over which change will occur may not
be). In contrast, ascertaining which of the other attributes
listed above is most important may require expert review
by hydrologists, ecologists and ecotoxicologists who can
determine potentiometric surfaces, flow patterns and aquifer
permeability and yield characteristics as well as water-quality
needs of vegetation before any attempt can be made to predict
how vegetation may respond to changes in level, flow rate or
hydraulic head. It is important to know which attribute of the
groundwater regime is most important at each site as this can
be used as the basis for defining a management target and a
monitoring system for managers.
To provide managers with empirical targets for
management, Loomes (2000) analysed the distribution
of 60 wetland species in relation to data on surface-
water depth, duration and timing of flooding of Swan
Coastal Plain wetlands. From this analysis she was able to
determine the ecohydrological range or average minimum
and maximum water depths required to maintain viable
populations of various wetland species, thereby providing
management goals. Similarly, Froend and Zencich (2001)
examined phreatophytic vegetation on the Gnangara mound
in Western Australia and identified a range of depths of
groundwater that resulted in a range of impacts for different
species. Froend and co-workers identified categories of
groundwater usage by vegetation, depending on groundwater
depth. The greater the depth to groundwater, the lower
the requirement for groundwater and the more tolerant the
vegetation to water-table decline, owing to the corresponding
increase in alternative water sources. These alternative
sources are primarily the larger volume of unsaturated
zone (with increasing depth) exploitable by the plant’s root
system. Currently, quantitative information suggests reduced
importance of groundwater to vegetation where depths to
groundwater exceed 10 m. However, it is assumed that
at depths of 10–20 m there is a possibility of vegetation
groundwater use, although it is thought to be negligible
in terms of total plant water use, and that at depths
of 20+m the probability of groundwater use is low
(Froend and Zencich 2001).
From the above discussion, it appears that the depth to
groundwater is often the most important attribute at sites
relying on subsurface provision of groundwater, whereas it
is the depth of inundation and frequency of inundation that
appear most important to ecosystems relying on both surface
expressions of groundwater and overland flow of surface
waters (floodplains, wetlands, base-flow rivers). However,
hydraulic head is an important attribute too. In the past
50 years, the number of free-flowing artesian springs has
declined at some locations because of the decline in hydraulic
head in the confined aquifer resulting from high rates of
extraction (Australian Water Resources Assessment 2000).
Hydraulic head and flux are also critical for aquifer and cave
ecosystems to maintain a supply of organic matter and oxygen
throughout the aquifer. Clearly, however, water quality is
a key issue across all GDEs and is of critical concern,
given the increasing levels of groundwater contamination
globally (Danielopol et al. 2003). Management strategies for
Management of groundwater-dependent ecosystems: functional methodology Australian Journal of Botany 105
managing these different attributes in order to maintain the
desired characteristics of ecosystem structure and function
(for example, vegetation composition or productivity) are
likely to differ.
What are the safe limits to changes
in groundwater regime?
Information gained during the determination of ecosystem
processes underlying groundwater dependency (see What
degree of dependency is expressed? above), plus the
information gained during determination of those attributes
of groundwater of most importance to the GDE (see What
attributes of a groundwater regime are important to a
GDE? above) will also yield information about the range
of change in groundwater attributes that can be tolerated by
the ecosystem. Implicit in this, of course, is that acceptable
change has been defined by stakeholders (e.g. managers, the
public, scientists and land owners). This point is discussed
in Murray et al. (2006). Also implicit in the question ‘what
are safe limits to change?’ is the requirement to be able to
measure change in ecosystem structure and function in a way
that allows monitoring of change through time. This issue is
discussed in Which features of vegetation can be measured
to monitor ecosystem function?
In a minority of cases, safe limits to groundwater
extraction can be relatively easy to ascertain. When the
relationship between hydraulic head and flow in an aquifer
is known, it might be relatively easy to ensure that the
hydraulic head is maintained so as to ensure a minimum
flow rate for conservation of mound springs, for example.
Presumably, flow rates into base-flow streams can be similarly
predicted as a function of hydraulic gradients. This is the
underlying philosophy behind the prevention of groundwater
pumping too close (within 3 km) to such rivers in the
Northern Territory (NT) of Australia (O’Grady et al. 2006).
Similarly, groundwater levels adjacent to rivers must be
sufficiently maintained to ensure continued flow of water
into the river to maintain the health of riparian vegetation
that uses groundwater in the dry season (Erskine et al. 2003).
Erskine et al. (2003) summarised a wealth of data derived
from several 3-year studies to determine the environmental
water requirements of flora and fauna of the Daly River
in the NT. Most importantly, this report provides explicit
management targets for minimum dry-season river-flow
rates, groundwater pumping regimes and irrigation water
extraction from the river. These targets are directly related to
the needs of specific floral and faunal populations, including
the maintenance of riparian vegetation, the protection of
peak floods to ensure lateral connection to floodplains, the
maintenance of disturbance events necessary for vegetation
regeneration and adequate water of suitable temperature and
provision of suitable breeding sites for pig-nosed turtles
and fish.
In the majority of cases, determining the safe range
of change in the key groundwater attribute is difficult
and time consuming. A priori considerations suggest
that ecosystems may show a proportional (simple linear
or a highly non-linear) response of ecosystem function
to declining groundwater availability (hereafter called a
degrading groundwater regime) or they may show a threshold
response whereby minimal change occurs until a threshold
of availability occurs. Although there are few empirical
studies of these a priori response models, there are several
potential approaches to defining safe limits to changes in the
groundwater regime. These include the following:
(1) Examine long-term hydrological data from monitoring
wells, along with long-term rainfall and vegetation
survey data. Where vegetation survey data are
absent, dendrochronological studies of tree rings and
photographic records (Fensham and Fairfax 2003) may
be useful. Long-term changes in average groundwater
depth and rates of change and the duration of maxima and
minima of groundwater depths are particularly important
in these analyses. Froend and co-workers (Froend
and Zencich 2001; Froend et al. 2004) established
such safe limits to changes in depth to groundwater
through analyses of long-term vegetation surveys and
groundwater data. These comparative data on mortality,
species composition and groundwater depth before and
after groundwater pumping, plus control sites where
groundwater pumping has not occurred, were used to
identify three classes of phreatophytic vegetation and
four classes of risk of impact of groundwater extraction
as a function of rate of drawdown and the magnitude of
drawdown (Froend and Loomes 2004).
(2) Groundwater modelling can be used to calculate the
response of the water table of an unconfined aquifer
and potentiometric surface of a confined aquifer to a
given rate of extraction, from which inferences can
be drawn about likely vegetation responses, by using
information from studies such as those described above.
Froend et al. (2004) used modelled groundwater level
change mapping to determine areas of phreatophytic
vegetation susceptible to drawdown as a result of future
groundwater resource development on the Swan Coastal
Plain, Western Australia. Similarly, models of aquifer–
surface water interactions can be used to predict changes
in stream-flow, from which inferences can be drawn about
likely floral and faunal responses (Erskine et al. 2003).
The most difficult tasks are formulating the inferences
about vegetation and faunal responses to these changes
in groundwater and stream hydrology. There is a paucity
of data describing the response of ecosystem function to
changes in groundwater regime.
(3) Multiple-site comparisons within a single climate
regime, with sites chosen specifically to represent
106 Australian Journal of Botany D. Eamus et al.
known differences in groundwater regime, can be used
to derive relationships between ecosystem structure,
function and groundwater regime. Loomes (2000),
Froend and Zencich (2001) and O’Grady et al. (2006)
applied this approach to establish ecological water
requirements of wetland species, Banksia woodlands and
riparian vegetation.
(4) A combination of detailed biological, hydrological
and topographical knowledge is required to establish
minimum groundwater regimes to maintain viable
populations of key, iconic or threatened species. Three
examples will suffice to show how such information
can be synthesised to provide management goals. In
the first, Georges et al. (2002, cited in Erskine et al.
2003) determined the minimum dry-season flow rates
in the Daly River that are required to maintain viable
populations of the pig-nosed turtle. By clearly identifying
the breeding, feeding and range requirements of the
pig-nosed turtle, plus the use of a one-dimensional
steady-state backwater model, breakpoints in the stream
that would impede the movement and breeding of pig-
nosed turtles were identified. From a knowledge of the
relationship between stream-volume flow rate and stream
topography, it was then possible to set a minimum dry-
season flow rate that would achieve a known level of turtle
nesting along the river.
In the second study, a similar approach was used to
determine flow rate needed to maintain in-stream macrophyte
populations. Preference curves for water depth, distance
from riverbank edge and mean and maximum flow velocities
at 27 river channel cross-sections during the dry season
were established for Vallisneria nana, an important species
providing habitat for turtles and other vertebrates. V. nana
occurred in the dry season within the depth range 0–1.3 m,
with a mean flow velocity range of 0–0.6m s1and a
maximum velocity within the range 0–0.75 m s1. The
probability of occurrence of dense beds of V. nana was
maximal at depths of 0.6 m and at 5 m from the bank.
Consequently, Georges et al. (2002, cited in Erskine et al.
2003) were able to identify the range of flow regimes
required to sustain these plant and associated dependent
faunal populations. By sustaining this keystone species, it
was assumed that sufficient habitat would be maintained to
support the full complement of other species.
Finally, Begg et al. (2001, cited in Erskine et al. 2003)
used a GIS to identify the extent and distribution of wetlands
in the Daly River Basin. They then identified the threats to
wetlands from current and future rates of water use from
the Daly River and provided an assessment of the risks
and groundwater requirements for the maintenance of these
wetlands. Ten different types of wetland were identified (river,
creek, channel billabong, backflow billabong, floodplain
billabong, floodplain, dampland, sumpland, waterhole and
doline). The spatial extent of these was mapped onto
1 :50 000 topographic maps on the basis of the waterlogging
characteristics of each land unit within the NT government
land unit maps, plus the landform and water regime
(permanently flooded, seasonally flooded or seasonally
saturated). Threats were identified as agricultural and
other land-use practices, road and other construction and
urban and mining developments and water extraction
from the river and groundwater sources. From these
threats and land-unit mapping, it was shown that different
proportions of the different wetlands were likely to be
affected by specific changes in land use. From this study,
specific rules about water extraction from the river were
developed. Such rules included the following: no water
extraction to occur on the rising stage and peak of flood
hydrographs during the wet season and water extraction
to a maximum of 20% of streamflow allowed when flood
stage has dropped at least 1 m below peak values in the
wet season.
What is the response function of species or ecosystems
to a degraded groundwater regime?
As previously discussed, a priori considerations suggest that
vegetation may show a proportional or threshold response
to a degrading groundwater regime. Field data to support
this are difficult to obtain. However, several approaches to
this question can be taken. In the first, modelling of links
between vegetation and groundwater at large scales, with
minimal attempt at resolution to species, can be attempted.
Structural features are used (e.g. forest, woodland, grassland)
to describe ecosystems. Bauer et al. (2004) and Rains et al.
(2004) adopted this approach (see below). A second approach
uses measurements of individual plants and data are collected
at the individual-plant scale, either for a dominant or group of
dominant species. Scott et al. (1999) and Groom et al. (2000a,
2001) (see below) adopted this approach. The third approach
is to study ecosystem processes, such as nutrient cycling
(Hefting et al. 2004; see below). Finally, a comparative
approach has been used to examine species diversity and
community structure of sites differing in groundwater regime
(Brunke et al. 2003).
Bauer et al. (2004) used diurnal fluctuation in groundwater
level in Botswana to estimate the contribution of groundwater
to evapotranspiration by phreatic vegetation. These authors
were able to model the influence of local vegetation cover
and soil characteristics. Similarly, Rains et al. (2004)
developed a model linking groundwater and vegetation
to simulate groundwater and vegetation distributions in a
riverine and reservoir-fringe community. Mean depth to
groundwater was modelled for a 20-year period for each
of five vegetation-community types and multiple vegetation
models were developed which generated a probability
of occurrence for each of the vegetation types as a
function of depth to groundwater and flooding. From this,
Management of groundwater-dependent ecosystems: functional methodology Australian Journal of Botany 107
changes in the distribution of each vegetation type were
predicted under various hydrologic scenarios. Importantly,
this approach was able to simulate multiple groundwater
and surface-water management regimes but also to predict
vegetation distributions in the riverine (riparian) and
reservoir-fringe communities.
Scott et al. (1999) established transects across a river
before mining activities which caused the depth of the
water table to increase 1–2 m above a 2-year period. Tree
survival, crown volume and stem increment of Populus
forests in the USA all declined following the decline in
water table. Similarly, increased mortality and changes in
species composition have been observed in a 30-year study
during which depth to groundwater increased, partly as a
result of groundwater pumping and partly because of a
significant decline in rainfall over the 30-year period (Groom
et al. 2000b). Both studies highlighted the value of long-
term vegetation surveys that start before pumping occurs
and continue during the period of pumping. When a survey
before pumping has not been undertaken, a comparison with
reference sites, where pumping is not occurring, must be used
to establish a baseline for comparison.
A threshold response does appear to be evident in the
studies of both Scott et al. (1999) and Groom et al.
(2000aand 2000b), although this could be a function
of the vegetation parameters examined. It is likely that
a more proportional response would be observed if
variables such as stomatal conductance and tree water
use (which show continuous rather than discrete variation)
were measured rather than crown volume or mortality.
Scott et al. (1999) identified branch diameter increment
as being more sensitive to changes in groundwater regime
than crown volume or tree survival and this showed a
proportional response rather than a threshold response.
Interestingly, branch growth is significantly correlated with
stream volume flow in cottonwoods (Willms et al. 1998),
presumably because stream flow increases when groundwater
depth decreases.
An alternative to examining biotic (species or
communities) responses to changes in groundwater
availability is to look at ecosystem processes. Hefting
et al. (2004) examined the influence of water-table depth
on soil nitrogen cycling in riparian wetlands. Nitrification
and denitrification are important processes influencing
plant productivity and are aerobic or anaerobic processes,
respectively. Consequently, flooding, which reduces the
oxygen content of soil, will have differential effects on these
processes. They showed that depth to water table was a
major determinant of soil nitrogen dynamics and observed
three threshold responses. When the water table was close to
the surface (<10 cm) ammonification was the main process
and ammonium accumulated in the soil. When the water
table was at >10-cm but <30-cm depth, denitrification was
the dominant process and nitrogen was lost from the soil.
When the water table was at >30-cm depth, nitrification
dominated and soil nitrate content increased. Unfortunately,
it is not immediately clear what the impact of these changes
will be on vegetation structure, but the approach deserves
further consideration.
Brunke et al. (2003) examined the effect of the exchange
between groundwater and surface water on the distribution
of aquatic invertebrates within a riverine landscape in two
floodplains of a southern Alpine river. The two floodplains
differed in hydrological regime. In the middle floodplain,
lateral inputs and exfiltration of hillslope groundwater were
important processes and bank infiltration of river water
supported subsurface water for only a very short distance
from the riverbank. Aquatic habitats in this floodplain
were relatively homogenous, with high taxon richness and
intra-habitat diversity. In the lower floodplain, groundwater
and river water exchange was more extensive and aquatic
floodplain habitats of this floodplain were mostly supplied
by alluvial groundwater, hyporheic exfiltration and surface
water. These floodplains showed low intra-habitat diversity
and a high inter-habitat diversity was present. Interestingly,
ordinations grouped the aquatic habitats according to the
origin of the water and species turnover was related to
differential lateral and vertical connectivity (Brunke et al.
2003). Consequently, changes in groundwater supply and
interactions with surface waters may result in changes
in species composition, diversity and ecosystem structure.
Such information is an important move towards being
able to predict ecosystem-response functions to changes in
groundwater regime.
Which features of vegetation can be measured
to monitor ecosystem function?
Having established the presence of a GDE and established a
management target (for example maintaining groundwater
depth within a certain range during the growing season),
it is important to have a vegetation response that can be
measured routinely and which will indicate that ecosystem
function is being maintained. Developing a set of triggers,
which prompts a management response to impacts, is a
critical stage in the adaptive management process (Downes
et al. 2002). A reference point, or particular value
of an indicator, is determined and deviance from this
reference point determines management action (Jamieson
and O’Boyle 2001). The importance of developing links
between monitoring programs and management decision-
making becomes evident as it is only through the reporting
of monitoring outcomes (i.e. the deviation of an indicator
from reference values) that management action can be
triggered (Morgan and Davis 1997; Finlayson and Eliot
2001). The response of management to a trigger being
breached will vary depending on the nature of the breach and
the management objectives. For example, Ecological Water
Provisions may have to be adjusted if monitoring indicates
108 Australian Journal of Botany D. Eamus et al.
that the environmental condition of a GDE has declined to a
level greater than is acceptable, or that a GDE appears to be
more resilient than predicted (Sinclair Knight Merz 2001).
Resilience, however, is particularly hard to detect with short-
term monitoring because of the time-lag between change in
groundwater regime and vegetation response.
If we ignore the impacts of polluted groundwater, the
impact of a degraded groundwater regime is essentially one
of a (more or less) gradually increasing imposition of water
(drought) stress. The response of native plants to drought
stress follows a reasonably consistent pathway and therefore
we can predict the cascade of responses (Fig. 1), from
molecular-scale to landscape-scale responses, which occur
over the time frame of days to decades (Smith and Griffiths
1993). It is not possible to review all the techniques that
could be applied in assessing the performance of vegetation
in response to a changing hydrologic balance. Pearcy et al.
(1994) and Le Maitre et al. (1999) provided discussions of
various methods that can be applied in the study of vegetation
responses to drought.
There are several criteria required to decide which
characteristics to measure, including traits that
(1) have a defined relationship with groundwater levels;
there needs to be confidence that a measured response
within a parameter reflects altered groundwater levels
rather than other abiotic/biotic factors;
stages of
stages of
Late stages
of drought
Shoot growth rate declines
Gene activity
Abscisic acid
Stomatal closure
observed Transpiration and
Leaves wilt,
xylem embolism
Leaf area index
over time
New species
present in
population New
to adult
Growth rate
Fig. 1. A conceptual model of short-, medium- and longer-term
changes in ecosystem function that can be monitored at a range of spatial
scales and levels of biological organisation.
(2) characterise risk to the environment; parameters
should identify, where possible, whether impacts to
environmental values are short-term or long-term,
reversible or irreversible and/or minor or major;
(3) are cost-effective and practical; parameters should
be inexpensive enough to measure, although
current monitoring practices may need to change
to accommodate more appropriate additional or
replacement parameters; parameters that reflect
landscape responses by GDEs of wide distribution, such
as remote sensing of phreatophytic vegetation health,
should be considered in light of the cost-effectiveness of
such approaches;
(4) have early warning capabilities; the lead-time (a measure
of how long it takes between the time at which a
parameter indicates there is likely to be a change within
an attribute, to the time that actual change occurs)
should be sufficiently long to provide the opportunity
to implement an appropriate management response;
generally, the better the warning (the longer the period
between potential change and actual change) the lower
the accuracy of the parameter in portraying a response
specific to a given stressor (i.e. depressed groundwater
levels); a balance between these characteristics (lead-
time and accuracy) should be considered to provide
the most appropriate and cost-effective parameters;
further characteristics of early warning indicators and
considerations which need to be taken into account when
deciding on environmental, physical and/or chemical
indicators are detailed in van Dam et al. (1998);
(5) consider the lageffects between changed groundwater
levels and environmental condition and/or health;
response of vegetation parameters influenced by changed
groundwater levels can take a long time to become
manifest and further reductions may occur before
impacts of previous changes are realized; consequently,
parameters with rapid responses are favoured, as they
provide advanced warning of significant stress or
degradation on the system, as well as providing the
opportunity to determine whether intervention or further
investigation is required (van Dam et al. 1998); however,
some GDE values may have to be measured through
parameters with a greater ‘lag’ time (e.g. phreatophytic
vegetation community composition).
The ability to make predictions of the impact of
a modified water regime on ecosystem components
depends on an understanding of the relationship between
the ecosystem component (e.g. wetland vegetation,
phreatophytic vegetation, macro-invertebrates) and the
water regime (Froend and Zencich 2001). Consideration
needs to be given to understanding and quantifying this
relationship at all ecological levels (community, population
and individual), as all are linked; an individual species
Management of groundwater-dependent ecosystems: functional methodology Australian Journal of Botany 109
response has implications for population response which
in turn influences community composition or structure
(Froend and Zencich 2001).
Community-level parameters
GDE water regimes are reflected in vegetation as it provides
numerous functions integral to the continued health of plants
and the organisms they support (Loomes and Froend 2001b).
Understanding the groundwater requirements of a plant
community requires knowledge of community responses
to short- and long-term effects of groundwater fluctuation,
groundwater abstraction and poor recharge events. Generally
though, these responses are poorly understood (Froend
and Zencich 2001). Regular monitoring of different plant
communities along with the underlying hydrology is
required before relationships between floristic structure and
composition and groundwater regimes can be confidently
described. Although more research is required to define
groundwater–vegetation relationships and to improve the
basis for decision-making with regard to predicting the
potential impacts of water-regime modification on vegetation,
Froend and Zencich (2001) provided details on specific
parameters to measure, with which vegetation response
to a given water regime can be correlated/associated
(Froend and Zencich 2001).
Loomes and Froend (2001a) explained that the
distribution, growth and reproduction of wetland vegetation
have strong relationships to depth, duration and amplitude of
seasonal flooding, demonstrated by measuring the response
of wetland vegetation to altered water regimes. As each
species is adapted to specific water-level ranges, changes
may cause a shift in community composition and structure.
For example, lowering water tables can result in a loss of
species intolerant of drying, and their gradual replacement
by terrestrial species. Detailed knowledge of the relationships
between plant-community composition and water regimes is
still lacking (Loomes and Froend 2001a, 2001b), although a
handful of studies have sought to address this issue (Froend
et al. 1993; Groom et al. 2000; Loomes 2000). Loomes (2000)
described the hydrology of 19 Swan Coastal Plain wetlands
in relation to the influence on composition and structure of
wetland vegetation and grouped wetland vegetation species
into hydrotypes, on the basis of the water regimes they
experienced, to predict the impact of altered hydrology on
wetland vegetation composition and structure.
Overstorey species, where present, are generally used to
define vegetation communities within wetlands and have
many important ecological functions, making them suitable
indicator species for monitoring wetland health. Overstorey
species also tend to persist in highly disturbed plant
communities when most other natives disappear (Pettit and
Froend 2001), implying a greater lag response to changes in
groundwater levels. As such, overstorey species are useful as
long-term indicators of environmental condition and health.
Similarly, emergent macrophytes form dominant
communities in wetlands and perform important ecological
functions, also making them suitable indicators for
monitoring wetland health (van Dam et al. 1998). Emergent
macrophytes are highly responsive to inter-annual variability
in wetland surface levels and as such are well suited as
short-term indicators, given their sensitivity to the early
stages of a stressor (i.e. depressed groundwater levels).
As weed species are relatively quick to colonise disturbed
areas, they may be useful as a short-term indicator of
disturbance within GDE. For example, invasive exotic species
such as Typha orientalis and annual exotic grasses can be
measured as an indication of the level of disturbance within a
wetland. This is important as high levels of weed invasion can
have negative effects on species diversity and recruitment of
native species (D¢Antonio et al. 1998) High levels of weed
invasion within wetland vegetation are reflective of changes
to community structure which, in turn, may be a result of
changes within the groundwater regime.
Abundance (area), character (composition, floristic
richness and structural diversity) and condition (collective
vigour) of phreatophytic vegetation can be measured at a
community level (Froend and Zencich 2001).
Distribution (reduction/expansion) of a community along
a water availability gradient may change in response to
altered groundwater regimes. Alteration in the dominant
species composition can be used as an indicator of change
in community distribution.
Species diversity and composition in a community may
change as species more vulnerable to prolonged dry periods
become locally extinct and in severe cases diversity may be
significantly reduced and comprise only xerophytic species;
alternatively, weed invasion may increase on loss of drought-
intolerant species (Froend and Zencich 2001); or structural
changes, such as height structure, may occur in a community
as mature trees senesce as a result of altered groundwater
regime (Froend and Zencich 2001).
Regeneration index over time (divide the number of
seedlings in the plot by the number of trees plus one, in the
plot, to give an indication of persistence of GDE vegetation
at a site); or canopy fullness/density of indicator species
(e.g. overstorey species) to give an indication of the health of
GDE vegetation.
Population-level parameters: indicator species
Population dynamics of appropriate indicator species can be
used to reflect the health and condition of GDE vegetation.
A subset of plants within a GDE can be used as indicators,
110 Australian Journal of Botany D. Eamus et al.
e.g. dominant overstorey species (Pettit and Froend 2001).
Overstorey species have many important ecological functions
within plant communities and as such are suitable indicators
for monitoring wetland health (Pettit 1997). Given the
relatively long ‘lag’ response of overstorey species to changes
in groundwater levels, they are useful long-term indicators of
environmental condition and health.
An example is the drought-sensitive holly-leaf banksia
(Banksia ilicifolia), which has been identified as an important
indicator of long- and short-term changes in groundwater
levels on the Gnangara mound and on other shallow aquifers
on the Swan Coastal Plain. Groundwater drawdown has a
documented negative impact on population size and vigour of
this species (Groom et al. 2001). B. ilicifolia is useful as a key
indicator species as it is an overstorey species, easy to record
and identify, and has been used to define the response of its
associated vegetation community to changes in groundwater
depth (Groom et al. 2001).
Abundance, character and condition of species response
to water regimes can also be measured at a population
level to describe response to water regime (Froend and
Zencich 2001).
The size of a local population is a measure of species
persistence and resilience. Recruitment potential can be
determined by the size and distribution of mature individuals
and is also affected by density. Higher-density stands of a
species require greater water availability. The distribution
of a population along a water-availability gradient reflects
the water requirements of a species and changes in water
regime that exceed the tolerance limits of individuals may
lead to a gradual change in species distribution. Distribution
of indicator plant species along a gradient can be used to
give an indication of the groundwater-availability gradient
as specific species are associated with specific depth-to-
groundwater ranges.
The size (height) and age structure of population can
be measured to give an indication of drawdown effects.
Populations affected by drawdown may demonstrate a lack
of recruitment as water availability is insufficient to support
successful establishment, and are characterised by few,
mature individuals and no new recruits. In contrast, dynamic,
resilient populations are characterised by many cohorts of
different ages, particularly young individuals. Froend and
Zencich (2001) noted an exception, however, when mature
plants succumb to drawdown events, leaving only younger
members of the population, tolerant of the ‘new’ water
regime. Longevity (and therefore persistence) and resilience
of a local population may also be significantly affected by
groundwater drawdown that exceeds the tolerance limits of
the population.
Population vigour, or the appearance of a species, can
be assessed by using indices of canopy vigour (e.g. canopy
condition index (fullness/density, presence/absence of
dead braches and epicormic growth), leaf area index).
Other population measures of vigour include incidence
of juvenile, mature, reproductive, flowering and senescent
individuals. Regeneration potential is also reflective of
condition or vigour of a population. Lack of regeneration
potential (lack of seedlings, no juveniles or saplings,
only senescent trees present) represents a loss of vigour.
However, to use regeneration potential as a measure of
drawdown impact requires other factors such as grazing
or fire management to be ruled out. Parameters used to
quantify regeneration potential include recruitment rate,
seedling survivorship and seed bank viability (Pettit and
Froend 2001).
Individual-level parameters
Individual plant response to water regimes can only be
measured and quantified in terms of condition. As such,
ecophysiological techniques that directly reflect the vigour of
a plant are used when measuring the condition of individuals
(Froend and Zencich 2001). These include the following
(Table 2):
(1) Measurement of plant water relations via pre-dawn
water potentials and gas exchange. These parameters
reflect plant response over time to water availability,
although provide only an indication of potential water
source use.
(2) Measurement of water flux via sap-flow techniques.
This provides information on the rates and timing of
consumptive use by a species. Can also be used to assess
changes in water use when water availability changes.
(3) Measurement of water sources used via isotopic tracers.
This allows water sources accessed by plants to be
identified and the contributions of potential water sources
to be recognised. This knowledge is critical for estimating
and modelling community-scale water balance.
The above parameters provide a more defined
relationship between groundwater levels and the condition
of phreatophytic vegetation than do community- and
population-level parameters; however, they are quite labour
intensive and costly but can (depending on sampling
frequency) have a very short lag time and therefore have
efficient early warning capabilities.
Environmental variables
When monitoring vegetation it is important to measure
environmental variables that will influence vegetation
communities. However, deciding which environmental
variables to measure will, in part, depend on the
Management of groundwater-dependent ecosystems: functional methodology Australian Journal of Botany 111
Table 2. A range of techniques is available to measure aspects of ecosystem function that might be expected to respond to a change
in hydrologic regime
Comparison with control (reference) sites greatly improves the utility of the measurements. These should be close to the experimental sites with
the same species and soil types present as at the experimental site
Measurement Technique Commentary Reference
Leaf water potential Pressure bomb Simple, cheap, but a nearby reference (control) Myers et al. (1997)
site is required because it responds to climate
(rainfall; temperature, solar radiation, vapour pressure
deficit) rapidly. Access to canopy for leaves
can be difficult.
Stomatal Leaf diffusion Simple, cheap, rapid, but a nearby reference Thomas et al. (2000)
conductance porometer (control) site is required because it responds to
climate (rainfall; temperature, solar radiation,
vapour pressure deficit) rapidly. Access to canopy
for leaves can be difficult.
Transpiration—tree Sapflow sensors Technically difficult and time-consuming; Zeppel et al. (2004)
or canopy scale for trees; eddy not cheap. A nearby reference (control) site is
covariance or Bowen required because it responds to climate
ratio for canopies (rainfall; temperature, solar radiation,
vapour pressure deficit) rapidly.
Leaf area index; Visual assessment; Visual assessment easy, rapid and cheap. Leaf area O’Grady et al. (2000)
canopy fullness; hemispherical index responds seasonally. Hemispherical
crown health photographs; LAI photographs technically difficult to analyse. Leaf
analyser, or remote area index analyser not suitable for all vegetation
sensing structures. Remote sensing increasingly available
but expensive and not available for all sites and
best suited to long-term (2+years) studies.
Growth rate Band dendrometers Simple, cheap and can be left in the field for years. Prior et al. (2004)
Seedling Fixed plots measured
establishment over time
Cover and Fixed plots measured Valuable quantitative assessment of population and Froend and McComb (1994),
abundance of over time (indicated) community response over the long-term Froend and Zencich (2001),
indicator plant (3+years). Simple and relatively cheap. Used as Froend et al. (1993, 2004)
species ground-truthing for remote sensing.
Community Visual assessment in Simple and relatively cheap. Provides a measure of Froend and Loomes (2004)
distribution/zonation fixed plots measured community response to shifting water availability
change over time gradient. Most often applied where distinctive
gradients in groundwater use are evident,
e.g. wetland fringes. Requires repeated-measures
over the long-term (3+years).
objectives of the monitoring program. Variables relevant to
assessing the response of vegetation to altered groundwater
levels include
(1) groundwater levels and fluctuating water regimes
(duration of wet/dry phases, seasonality);
(2) water quality (nutrient concentrations, salinity,
(3) soil water-retention capacity and soil stratigraphy (water-
retention layers above water table);
(4) climatic information (rainfall and maximum
temperatures during summer/early autumn) can
be useful in determining the cause of changes
to vegetation;
(5) records of past fires (as these may have strong impact
on vegetation composition and can compound effects of
other environmental factors, such as water regime, on
wetland vegetation;
(6) depth to the saturated zone of water (usually the capillary
fringe above the water table);
(7) rate of groundwater flow rate to the site;
(8) the hydraulic head within the aquifer; and
(9) location of discharge.
Two points bear noting about the choice of measurements
to make. First, there is generally a lag between a decline
in groundwater availability and an observable vegetation
response (Fig. 1). Once the watertable and associated
capillary fringe are beyond the root zone, changes in stomatal
conductance may occur within a short time (a few days), but
changes in leaf area index may require months to be evident
and changes in recruitment to the adult population of trees
112 Australian Journal of Botany D. Eamus et al.
will not be observed for years. The capacity for the soil profile
to store water and buffer the immediate impact of changes
in vegetation cover and watertable recharge was discussed
by Jolly and Cook (2002). Second, field measurements of
any variable will be ‘noisy’. Variation among different trees,
different species at a site, or among sites, can be very large
and consequently replication of measurements at all scales
should be designed such as to ensure an adequate signal to
noise ratio is achieved to allow detection of the impact of the
change in groundwater availability.
Concluding comments: putting GDE protection
into practice
The Council of Australian Governments endorsed water
resource-management reforms in 1994 to achieve a
sustainable water industry that included allocations for the
environment and greater environmental accountability of
water-resource developments. The National Principles for
the Provision of Water for Ecosystems produced by the
Agricultural and Resource Management Council of Australia
and New Zealand and the Australian and New Zealand
Conservation Council provided the basis for considering
ecological water requirements as part of water allocation
decisions by water resource managers.
In response to the Council of Australian Governments
and Agricultural and Resource Management Council of
Australia and New Zealand and the Australian and
New Zealand Conservation Council agreements, a variety of
approaches has been developed nationally to determine the
water requirements of dependent ecosystems. The majority
of these, however, have focussed on the requirements
of ecosystems dominated by surface-water flows with
fewer approaches directed entirely towards GDEs (Murray
et al. 2003). Research and water-resource planning in
Western Australia during the last decade (reviewed in
Froend et al. 2004) has identified several issues relevant
to a functional approach to determining ecosystem
groundwater requirements.
(1) There needs to be acknowledgment of variability in
groundwater dependency within a GDE; e.g. variability
in groundwater dependency of phreatophytic vegetation
relative to the depth of the water table and hydrological
ranges (tolerances) of wetland vegetation. Failure to do
so leads to insufficient awareness of biological/ecological
variability and incorrect interpretation of Ecological
Water Requirements as absolute ‘thresholds’ of
(2) Simplification of water requirements into minimum
water-table depths without recognition of other
hydrological variables important to the ecology
of the system can lead to an underestimation of
water requirements; e.g. duration, timing and rate of
seasonal flooding/drying and the episodicity of extreme
flooding/drying events need to be considered in a
addition to just groundwater level.
(3) The lag-response in the ecology of the system should
be considered particularly where GDEs have a history
of progressive decline in water tables. The cumulative
effects of reduced groundwater availability are important.
(4) The resilience of GDEs to drawdown impacts should
also be considered. Ecosystems affected by change in
groundwater availability may recover to an acceptable
(5) Consideration should be given to GDEs as part of a
system/catchment approach towards identifying water
requirements and possible impacts.
(6) An application of a risk (of impact) assessment which
incorporates variability in current vulnerabilities (water
requirements and drought stress) and potential degree
of change/impact will enhance the utility of the risk
assessment to managers of GDEs.
(7) Management (environmental compliance) criteria based
on simplified minimum ‘threshold’ water-table levels
without consideration of acceptable changes to ecological
values should not be developed because the definition
of acceptable change to ecological value is important to
stakeholders of the ecosystem being managed.
(8) Inaccurate assessment of groundwater levels/wetland
surface-water level relative to GDE ecology; e.g. no
groundwater monitoring at vegetation-monitoring sites.
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reflect the requirements of the ecosystem, often resulting
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... Identifying the relationships between changes in aquifer hydraulic head and spring discharge rates can be difficult as spring discharge rates are dependent on the fault conductance (see Section 2.2), which is difficult to accurately measure. Nevertheless, establishing relationships between water level patterns, spring discharge and the associated habitat characteristics and functions supporting sensitive species have been shown to be important for effectively managing these systems (Eamus et al., 2006;Fensham et al., 2010). Detailed field investigations are especially important in faultcontrolled spring systems as faulting and other geological heterogeneity can make the discharge pathways, and thus appropriate monitoring indicators difficult to establish (e.g., Wu et al., 2011). ...
... Many studies aimed at managing groundwater-dependent ecosystems have identified the importance of baseline and ongoing assessments that integrate knowledge from remote sensing, field-based monitoring, modelling and expert knowledge (Eamus et al., 2006;Humphreys, 2009;Doody et al., 2017). In addition to data obtained using traditional scientific measurement and sampling techniques, information about long-term hydrological and ecological functioning and variability of springs can be gained from historical records, indigenous knowledge, and archaeological evidence (Florek, 1993;Fensham et al., 2016a;Brake et al., 2019;Silcock et al., 2020). ...
... This suggested subsurface seepage may be occurring in these areas. This type of nearsurface discharge has been acknowledged as critical for certain groundwater-dependent ecosystems (see type 3 GDEs in Eamus et al. (2006)) and warrants further investigation as it may be as (if not more) sensitive to water level changes than surface discharge. ...
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Springs sustain groundwater-dependent ecosystems and provide freshwater for human use. Springs often occur because faults modify groundwater flow pathways leading to discharge from aquifers with sufficiently high pressure. This study reviews the key characteristics and physical processes, field investigation techniques, modelling approaches and management strategies for fault-controlled spring systems. Field investigation techniques suitable for quantifying spring discharge and fault characteristics are often restricted by high values of spring ecosystems, requiring mainly non-invasive techniques. Numerical models of fault-controlled spring systems can be divided into local-scale, process-based models that allow the damage zone and fault core to be distinguished, and regional-scale models that usually adopt highly simplified representations of both the fault and the spring. Water resources management relating to fault-controlled spring systems often involves ad hoc applications of trigger levels, even though more sophisticated management strategies are available. Major gaps in the understanding of fault-controlled spring systems create substantial risks of degradation from human activities, arising from limited data and process understanding, and simplified representations in models. Thus, further studies are needed to improve the understanding of hydrogeological processes, including through detailed field studies, physics-based modelling, and by quantifying the effects of groundwater withdrawals on spring discharge.
... GDEs require access to groundwater on a permanent or intermittent basis to meet their water requirements (Eamus et al., 2006), and limitations exist in identifying and mapping GDEs and determining their water needs. Groundwater extraction from unconfined aquifers with shallow water tables poses substantial risks to GDEs and can lead to ecosystem degradation, loss of ecosystem services and environmental damage (Tomlinson and Boulton, 2008;Eamus et al., 2015). ...
... Many approaches have been undertaken to locally identify and map GDEs (Eamus et al., 2006;Doody et al., 2017). While intensive field techniques are valuable, they are mostly restricted to research studies and are considered too costly for widespread application. ...
... This map provides a good first step for water managers, but coverage is incomplete and there can be issues of scale and resolution (Brim Box et al., 2022). Of course, the functional dependence of ecosystems on groundwater, as defined by the degree or seasonality of groundwater dependence (Eamus et al., 2006), is difficult to infer from remote sensing analyses which only provide information on 'potential' GDEs. Some jurisdictions have also undertaken their own GDE mapping, using a combination of remote sensing, and mapping of vegetation community structure and depth to groundwater (Kuginis et al., 2016). ...
Study region Australia Study focus Our incomplete knowledge of groundwater systems and processes imposes barriers in attempting to manage groundwater sustainably. Challenges also arise through complex institutional arrangements and decision-making processes, and the difficulty in involving stakeholders. In some areas, these difficulties have led to water table decline and impacts on groundwater users and groundwater-dependent ecosystems. However, there is potential to improve the sustainable use of groundwater resources through improvements in management practices. We discuss some of the challenges, and present survey results of research, government, and industry professionals across the groundwater sector in Australia. New hydrological insights for the region The highest-ranked challenge identified in the survey was the difficulty in determining regional-scale volumetric water extraction limits. This is surprising given the criticism in the international literature of volumetric based approaches for groundwater management, and the decreased reliance on this approach in Australia and elsewhere in recent years. Other major challenges are the difficulty in determining and implementing maximum drawdown criteria for groundwater levels, determining water needs of ecosystems, and managing groundwater impacts on surface water. Notwithstanding these gaps in technical understanding and tools and a lack of resources for groundwater studies, improvements in stakeholder communication should enable more effective decision-making and improve compliance with regulations designed to protect groundwater and dependent ecosystems.
... In arid areas, GDEs can be observed around natural water pans supported by groundwater [15]. Around natural water pans, the groundwater level is often high and can be easily accessed by groundwater-dependent vegetation (GDV) [16,17]. Ideally, areas around natural water pans are expected to be highly diverse when compared to environments further away from natural water pans where the groundwater level is low. ...
... Ideally, areas around natural water pans are expected to be highly diverse when compared to environments further away from natural water pans where the groundwater level is low. To assess the vegetation diversity around natural water pans, most studies have used field techniques due to their reliability and accuracy [16,18]. However, these techniques are laborious when working with extensive regions, such as transboundary aquifers, and may be costly in some resource-constrained environments, such as developing nations [19]. ...
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Groundwater-Dependent Ecosystems (GDEs) are under threat from groundwater over-abstraction, which significantly impacts their conservation and sustainable management. Although the socio-economic significance of GDEs is understood, their ecosystem services and ecological significance (e.g., biodiversity hotspots) in arid environments remains understudied. Therefore, under the United Nations Sustainable Development Goal (SDG) 15, characterizing or identifying biodiversity hotspots in GDEs improves their management and conservation. In this study, we present the first attempt towards the spatial characterization of vegetation diversity in GDEs within the Khakea-Bray Transboundary Aquifer. Following the Spectral Variation Hypothesis (SVH), we used multispectral remotely sensed data (i.e., Sentinel-2 MSI) to characterize the vegetation diversity. This involved the use of the Rao’s Q to measure spectral diversity from several measures of spectral variation and validating the Rao’s Q using field-measured data on vegetation diversity (i.e., effective number of species). We observed that the Rao’s Q has the potential of spatially characterizing vegetation diversity of GDEs in the Khakea-Bray Transboundary Aquifer. Specifically, we discovered that the Rao’s Q was related to field-measured vegetation diversity (R2 = 0.61 and p = 0.00), and the coefficient of variation (CV) was the best measure to derive the Rao’s Q. Vegetation diversity was also used as a proxy for identifying priority conservation areas and biodiversity hotspots. Vegetation diversity was more concentrated around natural pans and along roads, fence lines, and rivers. In addition, vegetation diversity was observed to decrease with an increasing distance (>35 m) from natural pans and simulated an inverse piosphere (i.e., minimal utilization around the natural water pans). We provide baseline information necessary for identifying priority conservation areas within the Khakea-Bray Transboundary Aquifer. Furthermore, this work provides a pathway for resource managers to achieve SDG 15 as well as national and regional Aichi biodiversity targets.
... From the different types of GDEs, this study considers wetlands and terrestrial ecosystems. The authors applied an inferential approach (Eamus et al. 2006) where, depth to groundwater and the water access of plant roots from the saturated zone are considered as decisive attributes. Monthly groundwater connection is identified either through (1) hydrologically active roots that reach the capillary fringe (GW dep1 ; Eqs. 12) or (2) through shallow groundwater (GW dep2 ; Eq. 15). ...
... Ecosystems' degree of groundwater dependence (Based on:Hatton and Evans 1998;Froend and Loomes 2006;Eamus et al. ...
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Water provision and distribution are subject to conflicts between users worldwide, with agriculture as a major driver of discords. Water sensitive ecosystems and their services are often impaired by man-made water shortage. Nevertheless, they are not sufficiently included in sustainability or risk assessments and neglected when it comes to distribution of available water resources. The herein presented contribution to the Sustainable Development Goals Clean Water and Sanitation (SDG 6) and Life on Land (SDG 15) is the Ecological Sustainability Assessment of Water distribution (ESAW-tool). The ESAW-tool introduces a watershed sustainability assessment that evaluates the sustainability of the water supply-demand ratio on basin level, where domestic water use and the water requirements of ecosystems are considered as most important water users. An ecological risk assessment estimates potential impacts of agricultural depletion of renewable water resources on (ground)water-dependent ecosystems. The ESAW-tool works in standard GIS applications and is applicable in basins worldwide with a set of broadly available input data. The ESAW-tool is tested in the Danube river basin through combination of high-resolution hydro-agroecological model data (hydrological land surface process model PROMET and groundwater model OpenGeoSys) and further freely available data (water use, biodiversity and wetlands maps). Based on the results, measures for more sustainable water management can be deduced, such as increase of rainfed agriculture near vulnerable ecosystems or change of certain crops. The tool can support decision making of authorities from local to national level as well as private enterprises who want to improve the sustainability of their supply chains.
... One commonly used method in Europe of quantifying the relationship between indicator plant species and various environmental variables, for example, is the Ellenberg index (Ellenberg, 1988;Hill et al.,1999). The type of vegetation in wetlands and its location is intimately entwined with the hydrological conditions, and hence, these ecohydrological relationships need to be properly understood and evaluated, from which healthy envelopes/thresholds can then be defined for the different key wetland habitats (Eamus et al., 2006;Bertassello et al., 2019;Bhatnagar et al., 2021a). Such metrics can be used in order to assess the ecohydrological status of wetlands and monitor and manage their current and future existence. ...
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Calcareous fens are peat-wetlands fed mainly by groundwater, located in topographic hollows and served by springs or seepages of water derived by contact with base-rich mineral ground. In the European Union they are protected habitats requiring special conservation measures. To better understand the environmental supporting conditions of alkaline fen habitat found in Ireland, a 2-years hydrological and hydrochemical monitoring programme was carried out on four contrasting fen sites that covered an ecohydrological gradient representing intact to highly degraded conditions. This paper presents a methodology (and its limitations) of how the evaluation of the temporal and spatial variability of the quality and hydrological dynamics associated with different representative vegetation communities within the four fens can be used to develop ecohydrological metrics for fen habitat. The conditions supporting fen habitat in “good” ecological condition, indicative of a functioning alkaline habitat, were determined and contrasted with conditions found in habitats deemed to be in “poor” ecological condition. Results indicate that an annual water level always above the ground surface within a threshold depth envelope of between 0.030 and 0.28 m is required for at least 60% of the year for healthy fen vegetation. In terms of hydrochemistry, higher concentrations of nutrients were found in the sediments at depth and the groundwater feed compared to the phreatic water table. This suggests the fens may be incorporating the incoming nutrients into their internal biogeochemical cycles and that there is a net accumulation of nutrients within the wetlands. Although envelopes of nutrient concentrations in the surface water of the fen associated with habitat in good ecological condition can be defined (e.g., dissolved reactive phosphorus concentrations of between 6 and 37 µg-P/l for PF1 habitat), these levels should be regarded as the water quality after the fen vegetation has effectively interacted with the higher incoming nutrient levels in the groundwater. Moreover, the local hydrogeological conditions that define groundwater pathways (and associated water quality impacts) in such calcareous fens, allied to the localized nature of these measurements in these field studies, makes it difficult to define groundwater water quality threshold values with any confidence.
Grassland degradation in the semiarid area of China is often attributed to overgrazing and expansion of irrigated agriculture without the consideration of groundwater. However, falling groundwater levels in the region have caused many problems for grassland ecology. In this paper,the typical sandy grassland- the Tongliao part of Horqin grassland was selected as the study area to study the spatial distribution of normalized difference vegetation index (NDVI) and the relationship between the NDVI and groundwater depth (GD) of each grassland pixel was studied. The temporal variation of the NDVI and its correlations with precipitation were investigated. The relationship between GD and vegetation species diversity in grassland was analyzed based on the vegetation survey. The results of the relation between GD and NDVI and vegetation species showed that GD can buffer the impact of increase in drought events on the vegetation because GD affected the composition and diversity of vegetation species. As the GD increases, the mean NDVI showed a trend of increasing first and then decreasing, and basically remained unchanged when the GD was greater than 3.5 m. The species diversity of grassland decreased with increasing GD and was not related to GD when GD was greater than 3.5 ∼ 4 m. Falling groundwater levels will cause great changes in the structure of grassland; specifically, azonal vegetation will gradually die and 3.5 m may be the critical GD needed to maintain the incomplete extinction of grassland azonal vegetation in the aeolian sandy soil. Correlation analysis proved that the correlation between NDVI and precipitation was significantly enhanced when GD increased sharply from 3 m to 6 m after 1999. This research indicates that the control of grassland degradation should pay attention to controlling the groundwater exploitation to recover a reasonable groundwater level, which was found to be necessary in the support of diverse and climate resilient of vegetation.
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Springs represent relevant habitats which support high levels of biodiversity and productivity, providing refugia to both plants and animals together with essential ecosystem services. However, springs are often overlooked or even destroyed by human activity. Locating, inventorying, and monitoring springs is, therefore, becoming of increasing importance. This study aims at developing a multidisciplinary approach to detect ephemeral springs, which can only be discovered by hydrogeological surveys when discharge occurs. We suggest potential cooperation between hydrogeologists and botanists based on the use of a plant species as an indicator of the occurrence of ephemeral springs fed by ophiolitic perched groundwater aquifers. In this study, the grass Molinia arundinacea was used as a bioindicator, observed in periods when there was no discharge and whose occurrence on ophiolite bodies was revealed by previous studies in the northern Apennines (Italy). A total of twenty ophiolitic bodies were explored and grassland stands including Molinia populations were discovered in 86 springs (15 reported for the first time). Detailed geomorphological and hydrogeological sampling was also performed in ten springs, recording the occupancy area, cover and functional traits of Molinia at two permanent and eight ephemeral springs and along their streambeds. Molinia not only colonised the spring outlets, but also occurred along the streambeds fed by ephemeral springs, with a higher occupancy area where the rocky outcrops were covered by detrital deposits and at the base of the morphological incision. Additionally, the cover of Molina populations was correlated with the average water discharge, with the highest functional trait values found in locations with high soil moisture. Results confirmed that while Molinia performs best on permanently moist soils, it is also able to grow on soils with a fluctuating water-table or even in dry conditions, thus representing an excellent indicator of springs fed by temporary perched aquifers.
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The imbalance between exploitation and recharge and the yearly decline in water tables is a common water resource problem in semi-arid areas in China. To ensure the sustainable use of groundwater and the stability of the regional hydrological cycle, a reasonable groundwater control table need to be determined to achieve a balance between groundwater exploitation and recharge. Based on the characteristics of the vertical hydrological cycle in semi-arid irrigation areas, this paper used a combination of in situ field experiments and software simulations to calculate the groundwater infiltration recharge under three types of subsurface conditions, namely, drip irrigation under mulch, border irrigation, and bare area, and to analyze the relationship between groundwater infiltration recharge and groundwater table. Based on the relationship between groundwater recharge and discharge, the critical groundwater depth of maintaining exploitation and recharge balance (CGDM) for drip irrigation under mulch was calculated to be 131.52-187.15 cm, and the critical groundwater depth of disrupting exploitation and recharge balance (CGDD) was 307.66-363.67 cm. For border irrigation, the CGDM was 80.00-84.34 cm and the CGDD was 198.87-248.44 cm. It also proposed a groundwater table management strategy to address the risk of imbalance in regional groundwater exploitation and recharge. The critical groundwater depth of controlling risk management (CGDC) can be used for regional groundwater management and guidance for agricultural irrigation, providing technical support for achieving regional groundwater exploitation and recharge balance and maintaining a stable hydrological cycle.
Groundwater supports the sustenance of major ecosystem functions including maintenance of native vegetation. Water availability to the groundwater-dependent ecosystems is directly influenced by the water table depth. Overexploitation and overextraction of groundwater resources have led to water scarcity and impending water crises all across the globe. In the present decade, much focus has been given to the rapidly depleting water resources and their impact on future generations. But the impact of the water table dropdown on the ecosystems it supports has been overshadowed by prioritizing human water needs and ways to replenish the water sources to fulfil population demand. Groundwater-dependent ecosystems support a wide range of ecosystem services and are critical habitats that need to be included in the watershed-level policies and water resources management initiatives, in order to build a sustainable ecosystem with the precise allocation of water resources. The present chapter discusses the Groundwater-Dependent Ecosystems (GDE) and their classification, briefly discussing the methods for their identification, along with the global advances in GDEs mapping and groundwater allocation trends. The chapter highlights the need for GDE assessment especially for developing and underdeveloped countries that are currently facing water stress.KeywordsVegetation water demandWater stressEcohydrologyStable isotopesRemote sensing
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Climate change, land cover change and the over–abstraction of groundwater threaten the existence of Groundwater-Dependent Ecosystems (GDE), despite these environments being regarded as biodiversity hotspots. The vegetation heterogeneity in GDEs requires routine monitoring in order to conserve and preserve the ecosystem services in these environments. However, the in–situ monitoring of vegetation heterogeneity in extensive, or transboundary, groundwater resources remain a challenge. Inherently, the Spectral Variation Hypothesis (SVH) and remotely-sensed data provide a unique way to monitor the response of GDEs to seasonal or intra–annual environmental stressors, which is the key for achieving the national and regional biodiversity targets. This study presents the first attempt at monitoring the intra–annual, spatio–temporal variations in vegetation heterogeneity in the Khakea–Bray Transboundary Aquifer, which is located between Botswana and South Africa, by using the coefficient of variation derived from the Landsat 8 OLI Operational Land Imager (OLI). The coefficient of variation was used to measure spectral heterogeneity, which is a function of environmental heterogeneity. Heterogenous environments are more diverse, compared to homogenous environments, and the vegetation heterogeneity can be inferred from the heterogeneity of a landscape. The coefficient of variation was used to calculate the α- and β measures of vegetation heterogeneity (the Shannon–Weiner Index and the Rao's Q, respectively), whilst the monotonic trends in the spatio–temporal variation (January–December) of vegetation heterogeneity were derived by using the Mann–Kendall non–parametric test. Lastly, to explain the spatio–temporal variations of vegetation heterogeneity, a set of environmental variables were used, along with a machine-learning algorithm (random forest). The vegetation heterogeneity was observed to be relatively high during the wet season and low during the dry season, and these changes were mainly driven by landcover- and climate–related variables. More specifically, significant changes in vegetation heterogeneity were observed around natural water pans, along roads and rivers, as well as in cropping areas. Furthermore, these changes were better predicted by the Rao's Q (MAE = 5.81, RMSE = 6.63 and %RMSE = 42.41), than by the Shannon–Weiner Index (MAE = 30.37, RMSE = 33.25 and %RMSE = 63.94). These observations on the drivers and changes in vegetation heterogeneity provide new insights into the possible effects of future landcover changes and climate variability on GDEs. This information is imperative, considering that these environments are biodiversity hotspots that are capable of supporting many livelihoods. More importantly, this work provides a spatially explicit framework on how GDEs can be monitored to achieve Sustainable Development Goal (SDG) Number 15.
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An increasing amount of information is being collected on the ecological and socio-economic value of goods and services provided by natural and semi-natural ecosystems. However, much of this information appears scattered throughout a disciplinary academic literature, unpublished government agency reports, and across the World Wide Web. In addition, data on ecosystem goods and services often appears at incompatible scales of analysis and is classified differently by different authors. In order to make comparative ecological economic analysis possible, a standardized framework for the comprehensive assessment of ecosystem functions, goods and services is needed. In response to this challenge, this paper presents a conceptual framework and typology for describing, classifying and valuing ecosystem functions, goods and services in a clear and consistent manner. In the following analysis, a classification is given for the fullest possible range of 23 ecosystem functions that provide a much larger number of goods and services. In the second part of the paper, a checklist and matrix is provided, linking these ecosystem functions to the main ecological, socio–cultural and economic valuation methods.
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Woody riparian vegetation in western North American riparian ecosystems is commonly dependent on alluvial groundwater. Various natural and anthropogenic mechanisms can cause groundwater declines that stress riparian vegetation, but little quantitative information exists on the nature of plant response to different magnitudes, rates, and durations of groundwater decline. We observed groundwater dynamics and the response of Populus fremontii, Salix gooddingii, and Tamarix ramosissima saplings at 3 sites between 1995 and 1997 along the Bill Williams River, Arizona. At a site where the lowest observed groundwater level in 1996 (-1.97 m) was 1.11 m lower than that in 1995 (-0.86 m), 97-100% of Populus and Salix saplings died, whereas 0-13% of Tamarix stems died. A site with greater absolute water table depths in 1996 (-2.55 m), but less change from the 1995 condition (0.55 m), showed less Populus and Salix mortality and increased basal area. Excavations of sapling roots suggest that root distribution is related to groundwater history. Therefore, a decline in water table relative to the condition under which roots developed may strand plant roots where they cannot obtain sufficient moisture. Plant response is likely mediated by other factors such as soil texture and stratigraphy, availability of precipitation-derived soil moisture, physiological and morphological adaptations to water stress, and tree age. An understanding of the relationships between water table declines and plant response may enable land and water managers to avoid activities that are likely to stress desirable riparian vegetation.
"Field screening" indicates field analytical tools, and (quick) methods and strategies for on-site or in-situ environmental analysis and assessment of contamination. "Field screening" includes not only field analytical methods, such as mobile laboratories, portable analyses, detectors, sensors, or noninvasive techniques, but also reconnaissance strategies and problems of measurement in heterogeneous media, using, among others, new geotechnical and geophysical instruments. This volume contains both oral and poster contributions to the Second International Conference on Strategies and Techniques for the Investigation and Monitoring of Contaminated Sites, "Field Screening Europe 2001", held in Karlsruhe, May 14 - May 16, 2001. As an integrated study of environmental contamination, "field screening" has become a more and more important part of environmental monitoring and the assessment of chemical contaminations. Recent developments are presented in these proceedings. Audience: Environmental engineers, geo-scientists, chemists, biologists, soil scientists, hydrologists and geophysicists.
Monitoring Ecological Impacts provides the tools needed by professional ecologists, scientists, engineers, planners and managers to design assessment programs that can reliably monitor, detect and allow management of human impacts on the natural environment. The procedures described are well grounded in inferential logic, and the statistical models needed to analyse complex data are given. Step-by-step guidelines and flow diagrams provide the reader with clear and useable protocols, which can be applied in any region of the world and to a wide range of human impacts. In addition, real examples are used to show how the theory can be put into practice. Although the context of this book is flowing water environments, especially rivers and streams, the advice for designing assessment programs can be applied to any ecosystem.
The role of groundwater in controlling ecosystems in Australia is poorly understood. The findings of a report prepared for the Land and Water Resources Research and Development Corporation (LWRRDC) entitled 'Dependence of Australian Ecosystems on Groundwater' are summarized. Four ecosystems were considered: terrestrial vegetation, river base flow systems, aquifer and cave ecosystems and wetlands. Criteria used to assess dependence on groundwater are outlined. The ecosystems were classified according to their dependence on groundwater. Case studies illustrating the complexities in evaluating the significance of ecosystem dependence on groundwater are presented.
Long-term changes in vigour and distribution of the dominant Banksia (5 species) and Melaleuca (1 species) overstorey species were examined within four vegetation transects overlying the Gnangara Groundwater Mound, a superficial unconfined shallow aquifer on the northern Swan Coastal Plain, Western Australia. All transects were positioned along topographical gradients and monitored over a 20-30 year period. The two co-dominant overstorey species (Banksia attenuata and B. menziesii) inhabited a range of topographical positions within the landscape, from dune crest to low lying areas, with only B. attenuata increasing its distribution (moving further downslope) within the transects over time. Both species displayed a reduction in vigour, as indicated by foliage condition, during the monitored period. Species commonly inhabiting low-lying winter-wet areas (e.g. Banksia littoralis, Melaleuca preissiana) showed the greatest loss of tree vigour in response to declining groundwater levels, with B. littoralis replaced by the more drought tolerant B. prionotes. M. preissiana populations were overall more resilient to altered groundwater regimes, responding over a much greater time period (many decades) than B. littoralis (<10 years). Overall, changes in species distribution and vigour were primarily caused by long-term declines in groundwater levels resulting from the cumulative effects of abstraction and below average annual rainfall (low groundwater recharge). Long-term distribution trends and overall observed reductions in population vigour within the transects may be a function of the species' dependency on groundwater to fulfil its water requirements. This may explain declining vigour and tree numbers of B. ilicifolia on the Gnangara Groundwater Mound, as this species is considered an important indicator of significant long- and short-term reductions in groundwater levels.
River damming and flow regulation can alter disturbance and stress regimes that structure riparian ecosystems. We studied the Bill Williams River in western Arizona, USA, to understand dam-induced changes in channel width and in the areal extent, structure, species composition, and dynamics of woody riparian vegetation. We conducted parallel studies along a reference system, the Santa Maria River, an unregulated major tributary of the Bill Williams River. Flood magnitude on the Bill Williams River has been dramatically reduced since the closure of Alamo Dam in 1968: the 10-yr recurrence interval flood in the pre-dam era was 1397 m(3)/s vs. 148 m(3)/s post-dam. Post-dam average annual flows were higher due to increased precipitation in a few years, but increases in post-dam May-September flows are largely attributable to dam operation. An analysis of a time series of aerial photographs showed that channels along the Bill Williams River narrowed an average of 111 m (71%) between 1953 and 1987, with most narrowing occurring after dam closure. Multiple regression analysis revealed significant relationships among flood power, summer flows, intermittency (independent variables), and channel width (dependent variable). The pattern of channel width change along the unregulated Santa Maria River was different, with less narrowing between 1953 and 1987 and considerable widening between 1987 and 1992. Woody vegetation along the Bill Williams River was denser than that along the Santa Maria River (27 737 stems/ha vs. 7559 stems/ha, P = 0.005), though basal areas were similar (14.3 M(2)/ha vs. 10.7 M(2)/ha, P = 0.42). Patches dominated by the exotic Tamarix ramosissima were marginally (P = 0.05) more abundant along the Bill Williams River than along the Santa Maria River, whereas the abundance of patches dominated by the native Populus fremontii or Salix gooddingii was similar across rivers (P = 0.30). Relative to Populus and Salix, Tamarix dominates floodplain vegetation along the Bill Williams River (P < 0.0001). Most stands of the dominant pioneer trees on both rivers became established in the 1970s and 1980s. Recent seedling establishment occurred in wider bands along the Santa Maria River (15.3 m wide vs. 5.4 m wide on the Bill Williams River, P = 0.0009), likely due to larger floods and associated seedbed formation along the Santa Maria River. Seedling survival rates were generally higher along the Bill Williams River, perhaps due to higher summer flows.
On gravel bars of lower Dry Creek, Sonoma Co., California, species became established from May-July. Seedling establishment was correlated with sediment texture. Salix spp. established preferentially on areas where surface sediment size was c0.2 cm. Populus fremontii established more densely on areas of intermediate and large-sized sediments (0.2-1.0 cm). Baccharis viminea dominated on the larger sediment sizes. Drought-induced mortality of seedlings was highest on bars where the stream had dried up completely prior to 1 September 1981. The high winter flows of the 1981-1982 season scoured the sediments and killed the vast majority of remaining seedlings. Seedlings which established in areas protected from the swiftest current were, however, able to withstand the winter flows. Changes of gravel bar landforms resulted in significant losses of established trees as well as young seedlings and saplings.-from Authors